PCBs and organochlorine pesticides in indoor environments - A comparison of indoor contamination in Canada and Czech Republic

Accepted Manuscript PCBs and organochlorine pesticides in indoor environments - A comparison of indoor contamination in Canada and Czech Republic Ondrej Audy, Lisa Melymuk, Marta Venier, Simon Vojta, Jitka Becanova, Kevin Romanak, Martina Vykoukalova, Roman Prokes, Petr Kukucka, Miriam L. Diamond, Jana Klanova PII:

S0045-6535(18)30858-0

DOI:

10.1016/j.chemosphere.2018.05.016

Reference:

CHEM 21344

To appear in:

ECSN

Received Date: 3 January 2018 Revised Date:

1 May 2018

Accepted Date: 2 May 2018

Please cite this article as: Audy, O., Melymuk, L., Venier, M., Vojta, S., Becanova, J., Romanak, K., Vykoukalova, M., Prokes, R., Kukucka, P., Diamond, M.L., Klanova, J., PCBs and organochlorine pesticides in indoor environments - A comparison of indoor contamination in Canada and Czech Republic, Chemosphere (2018), doi: 10.1016/j.chemosphere.2018.05.016. 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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PCBs and organochlorine pesticides in indoor environments - a comparison of indoor contamination in Canada and Czech Republic

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Ondrej Audy,1 Lisa Melymuk,1,* Marta Venier,2 Simon Vojta,1,+ Jitka Becanova,1,+ Kevin Romanak,2 Martina Vykoukalova, 1 Roman Prokes,1 Petr Kukucka,1 Miriam L. Diamond,3 Jana Klanova1

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6 RECETOX, Masaryk University, Kamenice 753/5, pavilion A29, 62500 Brno, Czech Republic

School of Public and Environmental Affairs, Indiana University, 702 Walnut Grove Avenue, Bloomington, IN, 47405 United States

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Department of Earth Sciences, University of Toronto, 22 Russell Street, Toronto, Canada M5S 3B1

11 * Corresponding author: Lisa Melymuk

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Email: [email protected]

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Phone: +420 549 493 995

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Address: RECETOX, Masaryk University, Kamenice 753/5, pavilion A29, 62500 Brno, Czech Republic

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Present address of S. Vojta and J. Becanova: Graduate School of Oceanography, University of Rhode Island, Narragansett, Rhode Island 02882-1197, USA

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Keywords: polychlorinated biphenyls; organochlorine pesticides; DDT; HCB; HCH; country differences

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ACCEPTED MANUSCRIPT ABSTRACT

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Polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) are restricted compounds

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that are ubiquitously detected in the environment, including in indoor matrices such as air and

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residential dust. We report concentrations of PCBs and selected OCPs in indoor air and dust from

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homes in Canada (23 homes) and Czech Republic (20 homes). Indoor air concentrations of PCBs and

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OCPs were ~10 times higher than that outdoors. PCB concentrations of ~450 ng/m3 were similar in

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both countries, higher in homes built before the restrictions on PCBs, and had congener profiles

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consistent with PCB mixtures manufactured or used in each country. All OCP air concentrations were

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higher in the Czech Republic than in the Canadian samples, suggesting greater indoor use of, for

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example, DDT and HCH. These data emphasize the persistence of these organochlorine compounds

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indoors and their presence in homes even after new usage was prohibited. Indoor levels of these

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legacy POPs remain at similar levels to compounds of current concern, such as brominated flame

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retardants and perfluorinated alkyl substances, emphasizing that they deserve ongoing attention in

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view of knowledge of OCP toxicity.

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INTRODUCTION

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In 1995, the United Nations Environment Program requested that international controls be placed on

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12 persistent organic pollutants or POPs (the “Dirty Dozen”), including polychlorinated biphenyls

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(PCBs) and nine organochlorine pesticides (OCPs). The impetus came from decades of research that

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documented their persistence and ability to cause harm to human health and the environment. These

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concerns led to the Stockholm Convention, which entered into force in 2004 with 152 countries as

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signatories (Stockholm Convention, 2008). The Convention called on the parties to “prohibit and/or

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eliminate the production and use, as well as the import and export” of the 12 POPs. For PCBs, this

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came some 30 years after regulations were passed in several countries to control production and new

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uses (Diamond et al., 2010). Still, today measurements continue to document human and

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environmental exposure to, and harm from, many of the “Dirty Dozen” (Diamond, 2017; Lyall et al.,

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2016; Rosenbaum et al., 2017).

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The indoor environment is an important pathway for human exposure to semi-volatile organic

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compounds (SVOCs), including the “Dirty Dozen”, via air and dust inhalation and dust ingestion

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(Booij et al., 2016; Dirtu et al., 2012; Harrad et al., 2006; Marek et al., 2017). PCBs and OCPs have

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been found in indoor air and dust due to past uses (i.e., primary sources; Booij et al., 2016;

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Frederiksen et al., 2012; Rudel et al., 2008), the subsequent contamination of and release from indoor

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materials (i.e., secondary sources; Marek et al., 2017), from current poorly-quantified sources such as

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impurities in other chemicals (Vorkamp, 2016), and emissions from new building materials and

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consumer products (Herkert et al., 2018; Hu and Hornbuckle, 2010). Levels of PCBs and often OCPs

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in indoor air are higher than in outdoor air (Harrad et al., 2006; Melymuk et al., 2016a; Rudel et al.,

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2010), and there is no recent evidence of declines in the indoor levels (Harrad et al., 2006). This

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points to POPs used or released indoors being more persistent than outdoors as fewer opportunities

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exist for losses through air flow and photolytic reactions (Bennett et al., 2014; Shin et al., 2013).

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Potential risks arise from exposure to POPs through indoor pathways as people in developed countries

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spend >90% of their time indoors (Klepeis et al., 2001; Matz et al., 2014), and inhalation exposure

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ACCEPTED MANUSCRIPT from indoor environments can contribute up to one-third of total exposure for lower-chlorinated PCB

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congeners (Ampleman et al., 2015).

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PCBs were high production volume chemicals used worldwide from the 1940s to 1980s, with global

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production of ~1.4 million tonnes. 650 000 tonnes were produced in USA from 1930 to 1977 by

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Monsanto as Aroclor mixtures, and 21 000 tonnes in Slovakia (former Czechoslovakia) from 1959-

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1984 by Chemko Stražské as Delor mixtures (Breivik et al. 2002). PCBs used in Canada were

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imported Aroclors from USA. PCBs were used as coolants, insulating fluids in transformers and

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capacitors, stabilizing additives in PVC, and as plasticizers in paint and building sealants, as well as in

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numerous other applications (WHO/IPCS, 1993). Ongoing primary use continues in older building

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materials and electrical equipment in countries that are not signatories to the Stockholm Convention

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and thus have not regulated their removal and destruction, notably USA (US EPA, 2017). Even for

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signatories of the Convention, regulations may specify only partial removal and destruction. For

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example, Canada mandated removal by 2009 of equipment containing PCBs <500 mg/kg or >50 to

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<500 mg PCB/kg if it is located in a sensitive location such as a child care facility or drinking water

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treatment plant and removal of most uses (e.g., PCBs in light ballasts, pole-top electrical

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transformers) by 2025 (Canadian Environmental Protection Act, 2015). The Czech Republic required

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removal of materials with >500 mg/kg PCB content by 2010 (National Centre for Toxic Compounds,

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2017). These regulations may not cover all PCB-containing materials, such as building sealants, for

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which concentrations can be less than these “action” levels (Diamond et al., 2010; Robson et al.,

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2010) or exemptions (e.g., in colouring pigments, electrical capacitors that are “an integral part of a

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consumer product”) (Canadian Environmental Protection Act, 2015). Nonetheless, these PCB-

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containing materials may be a concern in many indoor environments as they can lead to elevated

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indoor levels of PCBs and thus higher exposure (Frederiksen et al., 2012; Herrick et al., 2016; Marek

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et al., 2017).

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DDT was used as a potent insecticide in both indoor and outdoor applications. Although new uses

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were prohibited when it was added to the Stockholm Convention in 2004, three countries currently

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ACCEPTED MANUSCRIPT produce DDT and it is used in about 15 countries, primarily to control mosquito vectors of malaria

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(Stockholm Convention, 2009; van den Berg, 2009). DDT use was permitted in Canada until the mid-

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1970s, but a total ban on all uses only took place in 1985 “with the understanding that existing stocks

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would be sold, used or disposed of by December 31, 1990” (Environment Canada, 2013a). DDT was

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registered for use from 1951 to 1973 in Czechoslovakia; remaining unused stocks were destroyed in

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the 1990s (Holoubek et al., 2006; Turusov et al., 2002). Past indoor uses included direct pest control

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in homes and agricultural buildings, and in combination with PCBs, as a wood preservative (Abb et

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al., 2010). DDT undergoes degradation to DDD and subsequently to the more stable DDE.

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HCH was first produced as “technical HCH” which contained about 14% γ-HCH, the only conformer

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with insecticidal properties, 65–70% α-HCH, 7–10% β-HCH, and about 7% δ-HCH (Willett et al.,

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1998). The use of technical HCH was banned in Canada in 1971. Technical HCH was replaced on the

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market by purified (90%) γ-HCH, also called lindane. The major use of γ-HCH was to treat crops,

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seeds, soils and livestock, with additional uses as a pharmaceutical. Combined with DDT, it was also

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used to treat wood, including various art objects (Holt et al., 2017; Marcotte et al., 2014; Schieweck et

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al., 2007). Agricultural use of γ-HCH is restricted in the Czech Republic (since 1977; Holoubek et al.,

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2006) and Canada (since 2004; Health Canada, 2009), and the Czech Republic and Canada stopped

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the sale of all HCH-containing pharmaceuticals in 2008 (Scott and Chosidow, 2011) and 2011

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(Environment Canada, 2013b), respectively.

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Hexachlorobenzene (HCB) was a fungicide produced to treat seeds and preserve wood (WHO/IPCS,

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1997). Pentachlorobenzene (PeCB) is an impurity originating during the synthesis of several

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chlorinated pesticides, including HCB (Environment Canada, 2005). A major current source of both

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PeCB and HCB is incomplete combustion of organic material, e.g. solid waste and biomass burning

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(Bailey et al., 2009; Environment Canada, 2005; Nielsen et al., 2013). HCB was used primarily as a

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soil treatment in Czech Republic, with production ceasing in 1968 and use in 1977 (Holoubek et al.,

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2006). Canada first regulated HCB in 2003 and then strengthened these regulations in 2012

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(Environment Canada, 2017).

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ACCEPTED MANUSCRIPT PCBs and OCPs continue to be detected in indoor air (Blanchard et al., 2014; Bohlin et al., 2008;

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Harrad et al., 2006; Laborie et al., 2016; Tue et al., 2013; Wilson et al., 2001; Zhang et al., 2011) and

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in dust from homes, offices, and public buildings (Ali et al., 2012; Bräuner et al., 2011; Dirtu et al.,

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2012; Harrad et al., 2009; Tan et al., 2007). Although the major uses of PCBs and OCPs were in

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agriculture or in closed industrial applications, indoor concentrations are frequently found to be higher

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than outdoor levels, confirming their indoor persistence and the importance of indoor

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microenvironments in the total exposure to these chemicals decades after restrictions on use

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(Melymuk et al., 2016a; Vorhees et al., 1997; Wilson et al., 2011).

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In this study we report levels of PCBs and OCPs in indoor air and dust from homes in Czech Republic

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and Canada. We compare concentrations in the context of past production, usage and regulation, and

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evaluate the relative significance of legacy indoor pollutants relative to newer compounds, such as

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flame retardants. We also interpret the data in terms of effectiveness of regulatory controls since

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PCBs and OCPs were among the first chemicals to come under control through the Stockholm

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Convention.

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MATERIALS AND METHODS

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Sampling

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Samples were taken from 43 houses and apartments from two locations: 23 homes in Toronto, Canada

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(CAN) and 20 homes in Brno, Czech Republic (CZ) (Table S1) in June-August 2013. These locations

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were not influenced by local industrial emissions and were representative of domestic chemical use

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patterns. At least one room was sampled in each home and a second room was sampled in ~10 houses

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from each country, resulting in 35 samples from CAN and 30 from CZ. Participation was voluntary

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with written consent given and no compensation. Air samples were collected by 28-day deployments

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of polyurethane foam passive air samplers (PUF-PAS, Figure S1) and floor dust samples were

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collected with a household vacuum cleaner. All details regarding study locations, sampling and

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analysis are described by Venier et al. (2016) and are summarized briefly herein.

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ACCEPTED MANUSCRIPT All sampling materials were pre-cleaned before sampling by Soxhlet extraction (8 h in acetone, then 8

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h in toluene). On day one, PUF-PAS were deployed in either single bowl (CAN) or double bowl (CZ)

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configurations. After 28 days, PUF-PAS were collected, along with floor dust. Floor dust was

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collected using a polyester sock inserted into the hose of a household vacuum cleaner, and the largest

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possible accessible area in each room was sampled. All sampled media were wrapped in clean

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aluminum foil, sealed, labeled and subsequently stored at -20°C until analysis.

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Sample analysis

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In this paper we focus on seven indicator PCB congeners (∑7PCB is the sum of PCB 28, 52, 101, 118,

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138, 153, and 180), p,p’-DDT and related compounds (DDX is the sum of o,p’-DDT, p,p’-DDD, o,p’-

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DDD, p,p’-DDE, and o,p’-DDE), HCB, PeCB and HCH isomers (HCHs defined as the sum of α-, ß-,

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and γ-HCH) (Table S2).

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Extraction and fractionation of samples is described in detail in Venier et al. (2016). Before

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extraction, all samples were spiked with known amounts of recovery standards (PCB 30 and 185

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(Absolute Standards, Hamden, CT, USA) for CAN and CZ samples. Socks with dust were weighed,

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the dust was sieved to <500 µm, approximately 100 mg were separated for analysis, and the excess

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dust was archived. The sock was rinsed with 30 mL hexane in acetone (1:1 v:v), and the solvent was

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combined with weighed dust. Dust was sonicated in 30 mL of 1:1 acetone:hexane (v:v); left to settle

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for 30 minutes, and the supernatant was decanted. The procedure was repeated 2 additional times with

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10 mL of fresh solvent, and the extracts were combined. PUFs were extracted with 150 mL of

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dichloromethane (DCM) using automated warm Soxhlet extraction (Büchi B-811, Switzerland), the

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volume of the combined extracts was reduced under a N2 stream and split 70:30 by weight. The 70%

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aliquot was treated by sulfuric acid-modified silica and this extract was used for analysis of PCBs and

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OCPs. The remaining 30% of the extract was used for other analyses (Venier et al., 2016;

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Vykoukalová et al., 2017).

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Final extracts were analyzed for PCBs and OCPs by gas chromatography-tandem mass spectrometry

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(GC-MS/MS) using an Agilent 7890B GC coupled with an Agilent 7000B QQQ MS-MS. Three µL 7

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ACCEPTED MANUSCRIPT was injected in pulsed splitless mode. Compounds were separated on a 60 m x 250 µm x 0.25 µm

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HT8 capillary column (SGE, Australia) and analysed in electron impact (EI) mode. For each

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compound two selected multiple reaction monitoring (MRM) transitions (quantitative and qualifier

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ion) and their respective ratios were monitored to ensure correct compound identification and

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quantification. Samples were quantified using an internal standard added prior to instrumental

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analysis (PCB-121, Absolute Standards, Hamden, CT, USA). Further GC-MS/MS parameters are

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given in Tables S3 and S4.

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Masses collected by passive air samplers were converted to concentrations based on the sampling

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rates determined in a dedicated calibration study (details in SI). Sampling rates of 1.1 m3/day were

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used for the single-bowl sampler (CAN) and 0.63 m3/day were used for the double bowl sampler

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(CZ). These rates are comparable to indoor sampling rates previously determined for PUF-PAS

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samplers (Bohlin et al., 2014; Hazrati and Harrad, 2007; Saini et al., 2015).

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Samples from 20 homes from Bloomington, Indiana, USA (as described by Venier et al. 2016) were

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also quantified for PCBs and OCPs in the same laboratory as the CAN and CZ samples. However,

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these extracts were shipped to CZ for analysis and they suffered significant losses during transport.

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Therefore, these results are considered only qualitative, and are described in the SI for the purpose of

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comparison with the CAN and CZ indoor levels.

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QA/QC

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Sample recoveries were monitored for individual samples and averaged 80±10% for PCB 30 in air,

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96±8% for PCB 185 in air, 87±11% for PCB 30 in dust, and 99±19% for PCB 185 in dust. Masses for

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each compound were individually adjusted based on the recovery of the closest surrogate eluting in

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the same fraction in an individual sample. Further information on recovery correction is given in the

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SI.

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Solvent blanks were used to check for laboratory contamination. Three field blanks for each matrix

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per country were collected. Field blank averages and standard deviations are given in Tables S5-S8.

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ACCEPTED MANUSCRIPT Masses of target compounds in each sample were then compared to average masses in field blanks

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and treated as follows: if the blank average was <10 % of the amount in sample, no correction was

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used. If the blank average was 10-35 % of the amount in sample, the blank level was subtracted. If the

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blank average was >35 % of the amount in sample, the sample value was reported as below the limit

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of detection (LOD).

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For statistical analysis, all values below the limit of detection (LOD) were substituted with

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(√2/2)*MDL (method detection limit). For the compounds present in field blanks, MDL was

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calculated as average level in field blanks plus three times the standard deviation. For compounds not

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found in field blanks, the instrumental detection limit was used, calculated as the amount giving a

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signal with signal to noise ratio equal to 3 (Table S9).

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Since concentrations were log-normally distributed they were log-transformed for all statistical

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analyses. Statistical analyses and graphing was done using STATISTICA version 13 and Microsoft

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Excel 2010 software.

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RESULTS AND DISCUSSION

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Table 1: Summary statistics for PCBs and OCPs in air and dust by country. n is the number of samples >LOD. SE indicates standard error of the mean. The ANOVA column identifies concentrations that were significantly different from one another at p<0.05 based on ANOVA results for comparison between countries calculated on logtransformed data. Data denoted by “a” were significantly higher than data denoted by “b”. Data for individual compounds are given in Tables S10-11.

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Dust (ng/g)

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Air (pg/m3)

Canada

Σ7PCBs PeCB HCB HCHs DDX Σ7PCBs PeCB HCB HCHs DDX

Mean±SE Median

Range

734±162 455 222±56.3 113 376±23.7 363 514±105 274 329±79.3 99.6 69.1±39.4
109 - 5110 37.8 - 1800 160 - 767 64.1 - 2180 41.4 - 1550
Czech Republic n ANOVA Mean±SE Median 34 34 34 34 34 16 0 0 25 20

a a b b b b b a

Range

661±146 467 139 - 4230 103±11.3 87.1 52 - 314 768±61.3 711 294 - 1530 1500±359 746 206 - 7820 763±222 363 48 - 5650 79.3±20.8 75.1 11.4 - 358 1.2
n ANOVA 28 28 28 28 28 28 1 0 22 30

a b a a a a a a

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Figure 1 Boxplots of ∑7PCBs, HCHs, and DDX in air and dust in each country. The horizontal line shows the median, boxes are 25% and 75% percentiles, and whiskers are minimum and maximum without outliers. The same letters (ab) indicate no significant difference in ANOVA, using Tukey post-hoc test (p<0.05).

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PCBs

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respectively (Table 1). Concentrations of ∑7PCBs in CAN and CZ samples were significantly

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(p=0.0252) lower than those in USA samples from the same study (Table S14). After correcting for

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the number of congeners reported, levels found in this study were comparable with those reported in

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other European (Bohlin et al., 2008; Harrad et al., 2006) and Canadian (Zhang et al., 2011) homes

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(medians 160 – 1800 pg/m3), and higher than those found in Mexico (Bohlin et al., 2008) and

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Vietnam (Tue et al., 2013) (medians 46 and 57 pg/m3) (see Table S15). However, these concentrations

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were significantly lower than those in buildings with PCB-containing sealants (Frederiksen et al.,

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2012; Herrick et al., 2016; Marek et al., 2017). It is likely that indoor levels reported here came from

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building materials such as caulking, cable insulation, capacitors, paints and varnishes, light ballasts

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and/or other primary uses (ATSDR, 2000), or from contaminated outdoor areas. Moreover, the

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similarity to reported concentrations measured over the past 15 years suggests that indoor levels of

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PCBs have remained relatively consistent over this time period, with the differences controlled by

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such sources (e.g., building sealants) rather than temporal changes caused by fate processes (e.g.,

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ventilation, degradation reactions).

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The indoor air concentrations of ∑7PCBs in CZ (median of 467 pg/m3, range 139-4230) were 3-10

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times higher than outdoor air concentrations measured by the MONET passive sampling network,

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which recorded median concentrations of 29-175 pg/m3 for sites in the same region (Table S17,

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Borůvková et al., 2015). Similarly, indoor air concentrations of ∑7PCBs in CAN (median 455 pg/m3,

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range 109-5110 pg/m3) were 10 times higher than ∑7PCBs in outdoor air concentrations measured in

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the same region by the GAPS passive sampling network (Table S18, Borůvková et al., 2015). The

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higher indoor concentrations indicate a significant contribution of indoor sources of PCBs, rather than

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outdoor to indoor transport from external primary and secondary sources.

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PCB congener patterns differed between countries. Lower chlorinated congeners (PCB 28 and 52)

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generally dominated the air samples, consistent with their higher vapour pressures. However, while

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CAN samples contained on average 80% of PCB 28 and 52, CZ samples contained a higher fraction 11

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PCBs were detected in all air samples, with ∑7PCB medians of 455 and 467 pg/m3 in CAN and CZ,

ACCEPTED MANUSCRIPT (~50%) of penta- to hepta- congeners, mainly PCB 101 and 153 (Figure 2). These congener profiles

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were consistent with what has previously been observed in homes in CAN (Harrad et al., 2009) and

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CZ (Melymuk et al., 2016a).

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The differences in congener profiles between CAN and CZ may be related to the PCB technical

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mixtures that were in use in the two countries. PCBs used in North America were generally Aroclor

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mixtures produced by Monsanto Corporation (Frame et al., 1996), while those used in Czech Republic

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were Delor technical mixtures manufactured by Chemko Stražské in eastern Slovakia (Taniyasu et al.,

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2003). North American building sealants are believed to have contained highly chlorinated Aroclors,

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in particular 1248, 1254 and 1260 (Erickson and Kaley, 2011; Kohler et al., 2005; Robson et al.,

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2010), which are dominated by penta-, hexa- and hepta- congeners (Frame et al., 1996), while other

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PCB-containing building materials and products were predominantly formed from the lighter PCB

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technical mixtures (e.g., Aroclors 1016, 1221-1254; Erickson and Kaley, 2011). In CZ, Delor

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mixtures spanned similar ranges of chlorination to the Aroclor mixtures, but the mixture used for

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paints, varnishes and adhesives, Delor 106, was dominated by the hexa- and hepta- congeners

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(Holoubek et al., 2006; Taniyasu et al., 2003). The composition of indoor air in Czech homes has

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been found to be similar to this technical mixture (Melymuk et al., 2016a). We hypothesize that the

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difference in the PCB congener composition between CAN and CZ indoor air is related to differences

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(i.e., lighter congeners in CAN than CZ) in the PCB technical mixtures used in indoor applications

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(paints, sealants, adhesives, varnishes), and differences according to PCB manufacturer.

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80%

PCB 153 60%

PCB 138 PCB 118

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40% PCB 101 PCB 52

20% 0% CAN

CZ

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PCB 28

Figure 2: Average composition of PCB congeners in air samples

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In dust, PCBs were found in 45% of CAN samples (median
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median of 75.1 ng/g (Table 1). Levels were similar to those reported in homes from Romania, UK,

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USA and Canada (Dirtu et al., 2012; Harrad et al., 2009) and higher than those found in rural homes

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in Pakistan (Ali et al., 2012) (see Table S16). Because low molecular weight PCBs were generally

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below detection in dust, it was not possible to evaluate the differences in congener profiles in dust.

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The difference between levels in dust are consistent with the different PCB mixtures used between

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CAN and CZ, as noted above.

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Since new uses of PCBs in Canada and Czech Republic were prohibited in 1977 and 1984,

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respectively, we hypothesized that indoor PCB concentrations are related to building age. This

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relationship has previously been observed in Czech homes (Melymuk et al. 2016) and North

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American residences (Colt et al., 2005; Wilson et al., 2011). In our study, roughly half of CAN

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sampled homes were built after the country-specific PCB regulations on new uses and therefore

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should not contain any PCB-containing materials. In case of CZ, only 4 of the 20 homes were built

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after the 1984 PCB regulations. As expected, Σ7PCB air and dust concentrations were lower in post-

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than in pre-PCB regulation homes (Figure 3), with the difference for air being significant (Mann-

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Whitney U Test, p=0.006 for CAN, p=0.020 for CZ). The qualitative measurements in USA homes

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also show this relationship, with an even greater difference between pre-1977 and post-1977 home

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ACCEPTED MANUSCRIPT indoor air concentrations (p<0.001 for USA homes). The lack of significance between dust

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concentrations may have been due to the more heterogeneous nature of the dust matrix and more

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limited detection due to differences in the PCB mixtures used. Differences according to building age

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suggest, as with the congener profiles, that building materials are a major contributor to the indoor

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levels, and that these primary sources (paints, caulking, wiring, etc.) remain in place leading to long-

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term residential contamination by PCBs. It is also interesting to note the presence of PCBs in

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buildings constructed after regulations were passed and hence when no PCB-containing materials

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should have been used, indicating the ubiquity of these persistent compounds.

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Air

292 293 294 295 296

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ln c(Σ7PCB) [pg/m3]

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7

ln c(Σ7PCB) [ng/g]

291

CAN pre-ban CAN post-ban CZ pre-ban CZ post-ban

Dust Matrix

Figure 3 Boxplots of Σ7PCB in air and dust in homes built pre- and post PCB bans on new (but not existing) uses. The horizontal line shows the median, boxes show 25% and 75% percentiles, whiskers are maximum and minimum without outliers. The relevant years for determining whether a house was built post-ban were 1977 (CAN) and 1984 (CZ).

14

ACCEPTED MANUSCRIPT 297 298

DDT

299

medians of 99.6 pg/m3 in CAN, while in CZ they were higher than Σ7PCB, with a median of 363

300

pg/m3. Concentrations of both DDX and individual compounds were significantly higher in CZ than

301

in CAN (p=0.0144) (Table 1, Figure 1), with the exception of p,p’-DDT (see Table S10). Few studies

302

have reported indoor air concentrations of DDX compounds; the concentrations of DDX compounds

303

in CAN were close to those found in homes from the UK, while those in CZ were comparable to the

304

levels reported in Sweden and USA from 1996-1998, nearly 15-20 years earlier than our sampling

305

campaign (Bohlin et al., 2008; Leone et al., 2000; see Table S15). p,p’-DDE was the dominant

306

congener and the p,p’-DDE/p,p’-DDT ratio was >3 in most samples, indicative of old DDT

307

contamination (e.g. Bohlin et al., 2008). The ratio of o,p’-DDT/(o,p’-DDT+p,p’-DDT) was generally

308

0.5-0.6 (Table S10), which may suggest a contribution of dicofol as a source of DDT, as o,p’-DDT

309

exists as an impurity in dicofol (Venier and Hites, 2014).

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DDX air concentrations in North American homes were generally lower than Σ7PCB, with DDX

ΣDDX

1

-1

1950

CAN CZ

2000

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1900

AC C

ln CAIR [pg/m3]

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2

-2

-3

CAN ln C = -0.018*(Yr)+33.8 r2 = 0.41

-4

310 311

CZ ln C = -0.028*(Yr)+54.2 r2 = 0.62

Figure 4: Relationship between building construction age and indoor air concentration of ΣDDX.

15

ACCEPTED MANUSCRIPT Median indoor air concentrations of DDX compounds (363 pg/m3 in CZ and 99.6 pg/m3 in CAN)

313

were consistently higher by 4-100 times than the outdoor concentrations in the same regions (Table

314

S17 and S18; outdoor median 88.6 pg/m3 in CZ and 2.05 pg/m3 in CAN). Concentrations of DDX also

315

had a significant relationship with building age (Figure 4). This suggests, as for PCBs, that the indoor

316

concentrations of DDX compounds are due to past use of DDT inside buildings, e.g., use of building

317

materials treated with DDT during construction or renovation, or application of DDT inside homes for

318

pest control. These sources appear to sustain higher concentrations in older buildings. The higher

319

median concentrations in CZ compared to CAN may be influenced by the higher amount of homes

320

sampled in CZ from the time of DDT use (Table S1); the median construction year of sampled homes

321

was 1972 in CZ and 1986 in CAN, which reflects the average ages of the housing stock in each

322

country (average 1970 in CZ and between 1984 and 1995 in CAN) (Ministry of Regional

323

Development of the CR, 2016; Natural Resources Canada, 2017). In both countries, as with the PCBs,

324

DDX was found at low levels in buildings constructed after use was regulated.

325

In dust, DDX median concentrations were 45.8 and 41.5 ng/g in CAN and CZ, respectively (Table 1).

326

Despite differences in DDX in air between countries, the differences between countries in dust were

327

not significant, suggesting that the heterogeneous nature of indoor dust is greater than country

328

differences. Our results were comparable with those reported previously for indoor dust in Denmark

329

(Bräuner et al., 2011) and Pakistan (Ali et al., 2012), but lower than those from Eastern Romania

330

(Dirtu et al., 2012) (see Table S16).

331 332

HCHs

333

was significantly higher than the median of 274 pg/m3 in CAN (Figure 1, Table 1). HCH

334

concentrations in CAN were comparable with those reported in UK homes (Bohlin et al., 2008) and

335

higher than those found in Swedish or Mexican homes (Bohlin et al., 2008). As with PCBs and DDX,

336

HCHs in indoor air were significantly higher by more than 10-fold than outdoor concentrations in the

337

same regions; median outdoor air concentrations were 50.0 pg/m3 in CZ and 12.7 pg/m3 in CAN

338

(Tables S17 and S18). 16

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HCHs were found in air at similar concentrations to PCBs, with a median of 746 pg/m3 in CZ, which

ACCEPTED MANUSCRIPT Over ~84% of HCHs consisted of γ-HCH. Both γ-HCH and α-HCH were much higher in CZ

340

(p=0.003 and p<0.001 respectively) than in CAN, which may reflect the differences in CZ and CAN

341

housing stock age, as noted for DDT. Unlike γ- and α-HCH, which were detected in all air samples, ß-

342

HCH was detected in 68% of samples, and its levels were not significantly different between two

343

countries (p=0.267). α-HCH and γ-HCH isomers were strongly correlated in CZ homes (Pearson’s r =

344

0.59 and 0.80, respectively), but no correlation was found in CAN samples. This could indicate

345

different sources in CAN homes or reflect more newer homes in CAN than CZ, as the γ-HCH/(α-

346

HCH+γ-HCH) ratio was above 0.7 in all CZ samples, but lower than 0.6 in several CAN homes

347

(Table S10). In air, γ-HCH is resistant to photodegradation and, although hypothesized, isomerization

348

of γ-HCH to α-HCH in outdoor air has not yet been proved (Walker et al., 1999). The more plausible

349

explanation for the lower ratio of γ-HCH/(α-HCH/γ-HCH) is possible past combined usage of both

350

lindane and the technical HCH mixture which contains 55 – 80 % of α-HCH compared to γ-HCH (8 –

351

15 %).

352

HCHs in indoor dust were previously reported in only a few studies. CAN concentrations were

353

comparable with those reported in Singapore (Tan et al., 2007) and Pakistan (Ali et al., 2012),

354

whereas CZ levels were higher than in Singapore and Pakistan, but lower than concentrations found in

355

indoor dust in Romania (Dirtu et al., 2012).

356

Chlorobenzenes

357

Levels and geographic patterns of HCB were similar to those of HCHs, with a median air

358

concentration of 711 pg/m3 in CZ, significantly higher than the median concentration of 363 pg/m3 in

359

CAN (Figure 1 and Table 1). Air concentrations of PeCB were significantly lower in CZ, with a

360

median of 87.1 pg/m3, compared with 113 pg/m3 in CAN. PeCB concentrations correlated well with

361

HCB (Pearson’s r = 0.59 and 0.73 in CZ and CAN, respectively). PeCB has multiple past and present

362

sources: it was used independently as a fungicide and flame retardant; existed as an impurity in HCB

363

and other pesticides; it is also a degradation product of HCB; was part of a chlorobenzene mixture

364

used to reduce viscosity of PCB products; and is produced unintentionally during combustion and

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17

ACCEPTED MANUSCRIPT industrial processes (Environment Canada, 2005; Stockholm Convention, 2009). The correlation

366

between HCB and PeCB suggests that HCB related sources of PeCB are dominant contributors to

367

indoor levels of PeCB in CAN and CZ.

368

Few studies have reported levels of HCB in indoor air. CAN results were similar to concentrations

369

reported for UK (Bohlin et al., 2008) and France (Laborie et al., 2016), while higher than levels in

370

Sweden and Mexico (Bohlin et al., 2008). One CZ sample contained, to our knowledge, the highest

371

concentration of HCB reported in residential air (1.5 ng/m3).

372

Nearly all dust samples were below detection limits for HCB and PeCB, which is not surprising given

373

their relatively high vapour pressures.

374

Correlation between air and dust concentrations

375

Air and dust concentrations were compared to provide insights into their equilibrium status and hence

376

relative age of these chemicals indoors. This was accomplished by estimating the equilibrium gas-

377

phase air concentration using measured dust concentrations, and comparing with the measured air

378

concentrations (Figure 5). Although PUF-PAS are known to sample bulk (gas+particle) air indoors

379

(Bohlin et al., 2014), here target PCBs and OCPs were largely expected to be in the gas-phase at

380

indoor temperatures. As such, air concentrations obtained from the PUF-PAS were assumed to

381

represent gas-phase air concentrations. Gas-phase air concentrations were estimated using the

382

methods described in Melymuk et al. (2016b), and previously applied by Venier et al. (2016) and

383

Vykoukalová et al. (2017). Details are also given in the SI.

384

Measured air concentrations correlated well with air concentrations estimated from dust for all

385

compounds with detection in >15 samples, except for PCB 180 (Table S20). Spearman correlation

386

coefficients were between 0.34 and 0.76, p<0.05.

387

In general, estimated air concentrations were within one order of magnitude of measured

388

concentrations, as indicated by slopes between 0.22 and 1.1 for the linear regression of estimated vs.

389

measured air concentrations, consistent with equilibrium between air and dust. This suggests stable

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ACCEPTED MANUSCRIPT indoor conditions and hence no rapid changes in either primary or secondary sources, as would be

391

expected for legacy compounds such as PCBs and OCPs. This contrasts with what is seen with

392

compounds for which primary sources change rapidly over time and for higher molecular weight

393

compounds such as halogenated flame retardants with logKOA >11. High values of logKOA bring a

394

kinetic limitation on air-dust exchange, due to the length of time required to reach equilibrium vs.

395

time scales imposed by air exchange rates relative to volatilization from primary sources in indoor

396

environments (Venier et al., 2016; Weschler and Nazaroff, 2010; Zhang et al., 2011). (b)

0.1

0.01

0.001 0.001

0.01

0.1

1

3

Measured air conc. (pg/m )

TE D

PCB 138: Spearman r: 0.45 (p = 0.03) Slope: 0.45 397

10

SC

Predicted air conc. (pg/m3)

1

1

0.1

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Predicted air conc. (pg/m3)

(a)

RI PT

390

0.01

0.001

0.0001 0.01

0.1

1

10 3

Measured air conc. (pg/m )

p,p‘-DDT: Spearman r: 0.73 (p < 0.0001) Slope: 0.92

Figure 5 Relationship between estimated and measured air concentrations for (a) PCB 138 and (b) p,p’-DDT. The line indicates a 1:1 relationship between estimated and measured concentrations, and Spearman r is given for correlation between estimated and measured air.

401 402

Implications for human exposure

403

compared with dietary exposure to evaluate the relative importance of indoor environments as a

404

source of human exposure. Equivalent dietary exposure information was not available for Canada.

405

Details of all input variables, equations and assumptions used to estimate exposure are given in Table

406

S21. Under typical exposure scenarios (i.e., median levels in air and dust and average

407

inhalation/ingestion ratios) exposure to Σ7PCBs, HCHs and DDX via indoor pathways was small,

408

representing <5% of total exposure, or up to 10-17% for HCB. However, under high exposure

409

scenarios experienced by those living in the highest level indoor environments and with 95th

EP

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Indoor exposure via inhalation and dust ingestion was estimated for Czech adults and toddlers, and

19

ACCEPTED MANUSCRIPT 410

percentile intake rates, the contribution could be significant, accounting for 20-60% of total exposure

411

to DDX, PCBs and HCHs.

412 413

Comparison of legacy and current-use SVOCs

414

diphenyl ethers (PBDEs), “novel” halogenated flame retardants (NFRs) in air and dust (Venier et al.,

415

2016), organophosphate ester flame retardants (OPEs) in air and dust (Vykoukalová et al., 2017), and

416

perfluorinated alkyl substances (PFASs) in dust (Karásková et al., 2016). This gave us the opportunity

417

to examine the differences between levels of legacy pollutants for which new uses ceased decades ago

418

with those still in use or only recently banned from new use and production.

(a)

(b) 100000

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10000

1000

100

10

1

423

NFRs

10000

PFCs OCPs PCBs

1000

OPEs

100

10

1 CAN

CZ

Figure 6 Medians of legacy and current-use SVOCs in (a) air, and (b) dust, by country. Data for PBDEs and NFRs are from Venier et al. (2016), PFASs are from Karásková et al. (2016) and OPEs from Vykoukalová et al. (2017). PFASs in air are not available.

EP

420 421 422

CZ

PBDEs

AC C

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CAN

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Air concentration (pg/m3)

100000

SC

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Samples used in this study were also analyzed for current-use SVOCs, including polybrominated

424

OPEs had the highest concentrations in both countries and matrices by orders of magnitude (Figure

425

6). This finding points to high usage of OPEs in consumer products and building materials where they

426

are typically added at percentage levels that exceed that of brominated flame retardants (e.g.,

427

Stapleton et al., 2009). However, the air concentrations of the legacy PCBs and OCPs were second to

428

those of OPEs (although two orders of magnitude less), and notably about one order of magnitude

429

higher than the other currently or more recently used compounds (PBDEs, NFRs; Figure 6a), as has

430

been noted previously in the UK (Harrad et al., 2006), Japan (Takigami et al., 2009), and Canada 20

ACCEPTED MANUSCRIPT (Zhang et al., 2011). PBDEs and NFRs were up to two orders of magnitude higher than PCBs and

432

OCPs in dust in CAN, or comparable in CZ (Figure 6b). The PFAS median was similar to those of

433

PCBs and OCPs in both countries.

434

These patterns raise three points. First, higher dust than air concentrations of “newer” PBDEs and

435

NFRs reflects a move towards the use of lower vapour pressure compounds, presumably intended to

436

avoid the high mobility of legacy POPs conferred by their higher vapour pressures (Wania, 2003).

437

However, considering human exposure, this may reflect merely a shift in the primary exposure

438

pathways from inhalation to dust ingestion, a pathway which is particularly relevant for young

439

children (Jones-Otazo et al., 2005; Larsson et al., 2018; Malliari and Kalantzi, 2017).

440

Second, the elevated concentrations of legacy POPs indoors clearly demonstrates their environmental

441

persistence. The elevated levels in indoor air compared to outdoor air strongly support the influence

442

of ongoing primary sources of POPs indoors, decades after restrictions, in building materials (for

443

PCBs), or due to intentional use and incorporation into building materials (for pesticides). We see that

444

building ventilation and dust cleaning have limited ability to clear these chemicals from indoor

445

environments. Moreover, differences in the typical age of the housing stock between countries, e.g.,

446

older homes in CZ, are associated with differences in levels of legacy POPs which could lead to

447

regional differences in residential exposure.

448

Third, these levels shed light on the difficult task of effective chemical management. It is sobering to

449

consider the effectiveness of chemical regulations: nearly 15 years after international controls were

450

implemented through the Stockholm Convention and 40 years after PCBs were first regulated at

451

national levels, the “Dirty Dozen” continue to be found at elevated levels in indoor environments.

452

Higher levels of PCBs in homes built prior to the initial regulations, and continuing exposure to these

453

PCBs, demonstrate the difficulty in enacting regulations that provide for population-wide safety.

454

These results emphasize that regulation of new uses does not mean that the chemical is banned, which

455

gives the impression of “gone”.

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ACCEPTED MANUSCRIPT

CONCLUSIONS

457

PCBs and OCPs were consistently detected in indoor air and dust in homes from Canada and Czech

458

Republic at levels comparable to those reported in similar environments over the past 15 years. Some

459

country-specific differences were discernable from past usage patterns, such as the congener profiles

460

of PCBs between Canada and the Czech Republic, and higher indoor concentrations of DDX and

461

HCH in CZ. Relationships between indoor air concentrations and building age showed higher levels

462

in older buildings and support the hypothesis of continuing active primary indoor sources of legacy

463

POPs as an important contributor to indoor levels. It is also important to note the presence of legacy

464

compounds in buildings constructed after regulations were enacted. Indoor air concentrations were, on

465

average, 10 times higher than those outdoors, suggesting on-going opportunities for human exposure.

466

Decades after national and international restrictions on new uses, air concentrations of PCBs and

467

OCPs remain stubbornly consistent, reflecting their persistence indoors, the importance of both

468

primary and secondary sources, and the challenges faced in managing chemicals after they have been

469

used widely.

470

Acknowledgments

471 472 473 474 475 476 477 478

This work was supported by the Czech-American Scientific Cooperation Program (AMVIS/KONTAKT II, LH12074). RECETOX research infrastructure was supported by the Czech Ministry of Education, Youth and Sports (LM2015051) and the European Structural and Investment Funds (CZ.02.1.01/0.0/0.0/16_013/0001761). Support in Canada was provided by Allergy, Genes and Environment Network (AllerGenNCE) and the Natural Sciences and Engineering Research Council of Canada (NSERC). We thank Joseph Okeme and Amandeep Saini (University of Toronto) for collecting samples in Canada. We also thank all volunteers in Bloomington (IN, USA), Toronto (Canada) and Brno (Czech Republic) for their participation.

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Highlights: PCBs and OCPs were detected air and dust from homes in Canada and Czech Rep.

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Higher concentrations of pesticides were found in Czech homes

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Older buildings contained higher concentrations of PCBs and DDTs

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Composition of PCB congeners differs between Czech and Canadian homes

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Levels of PCBs, OCPs exceed those of in-use compounds >30 years after restrictions

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