Science of the Total Environment 470–471 (2014) 527–535

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Emissions of two phthalate esters and BDE 209 to indoor air and their impact on urban air quality Anna Palm Cousins a,⁎, Tomas Holmgren b,1, Mikael Remberger a a b

IVL Swedish Environmental Research Institute Ltd, Box 21060, SE-10031 Stockholm, Sweden Umeå University, Department of Chemistry, SE-901 87 Umeå, Sweden

H I G H L I G H T S • • • •

Emissions of phthalates from PVC materials and urban atmospheric occurrence were measured. The SMURF model produced concentration estimates within a factor of 1.5–10. Indoor sources have limited impact on the occurrence of phthalates in urban air. Indoor releases of BDE 209 are estimated to contribute 38% to the mass entering the city with inflowing air.

a r t i c l e

i n f o

Article history: Received 24 May 2013 Received in revised form 8 October 2013 Accepted 8 October 2013 Available online 26 October 2013 Editor: Adrian Covaci Keywords: Emission Polyvinylchloride Urban model Exposure Ventilation Fate

a b s t r a c t Estimated emissions of decabrominated diphenyl ether (BDE 209) and the two phthalate esters diethylhexyl phthalate (DEHP) and diisononyl phthalate (DINP) to indoor air in the Stockholm conurbation, Sweden were used to assess the contribution of chemical outflows from the indoor environment to urban outdoor air pollution for these substances, by applying the recently developed Stockholm MUltimedia URban fate (SMURF) model. Emission rates of DINP from PVC materials were measured and published emission rates of DEHP were adapted to Swedish conditions. These were used as input to the model, as well as recently reported estimates of BDE 209 emissions to indoor air in Stockholm. Model predicted concentrations were compared to empirical monitoring data obtained from the literature and from additional measurements of phthalates in ventilation outlets and urban air performed in the current study. The predicted concentrations of the phthalates DINP and DEHP in indoor air and dust were within a factor of 1.5–10 of the measured concentrations. For BDE 209, predicted indoor concentrations were within the measured ranges, but measured concentrations showed a much larger variability. An adjusted emission scenario to better fit observed concentrations indoors was employed for DEHP and final outcomes resulted in estimated indoor emissions of 250 (50–1250), 2.9 (0.58–15), and 0.068 (0.014–0.34) kg year−1 for DEHP, DINP and BDE 209. These emissions could not explain the observed concentrations in urban air of the phthalates, suggesting an underestimation of background inflow or existence of additional sources in the outdoor environment. For BDE 209, the assessment indicates that the Stockholm indoor environment contributes about 25% to the air pollution load in inflowing background air, but additional monitoring data in urban air are needed to confirm this conclusion. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Indoor air quality and chemistry and its effect on human health have been studied for more than two decades (Weschler, 2011). Studies concerning emissions to indoor air have targeted volatile organic chemicals (VOCs) from e.g. adhesives, consumer- and building products (Brown, 2009; Girman et al., 1986; Knöppel and Schauenburg, 1989; Seaman et al., 2007; Wallace et al., 1987) and several studies have been performed monitoring the concentrations of volatile substances ⁎ Corresponding author. Tel.: +46 8 598 56 300; fax: +46 8 598 56 390. E-mail address: [email protected] (A.P. Cousins). 1 Present address: National CBRN Defence Centre, SE-90176 Umeå, Sweden. 0048-9697/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.scitotenv.2013.10.023

in indoor air (Gustafson et al., 2004; Loh et al., 2006; Wu et al., 2011). Bennett and Furtaw (2004) developed a multimedia fugacity model to study the indoor fate of volatile pesticides. During the last decade, an increasing number of studies have focused on the occurrence and fate of semivolatile organic chemicals (SVOCs) indoors. SVOCs occur in a variety of materials and consumer articles used for indoor applications, including plastic materials used in floors, cables, toys, outer casings for electronic equipment, as well as textiles and upholstered furniture (Schripp and Wensing, 2009; Wensing et al., 2005). Many of these substances are not chemically bound into the polymeric material in which they occur, and can be released from the products throughout their lifetime (Afshari et al., 2004; Kemmlein et al., 2003; Schripp and Wensing, 2009). In the indoor environment,

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where humans spend about 95% of their time (Schweizer et al., 2006), releases from products may contribute significantly to human exposure to SVOCs (Harrad et al., 2010). Andersson and Rännar (2009) and Rännar and Andersson (2010) illustrated a method to identify combinations of chemicals and articles of potential concern for human and/or environmental exposure, and found diisononyl phthalate (DINP) in polyvinylchloride (PVC) flooring to be one combination of potential concern. DINP is used as a plasticizer and is a replacement for diethylhexyl phthalate (DEHP). Both DINP and DEHP have mainly been used as additives in PVC materials. Floors and vinyl walls account for a large part of the indoor PVC applications. Another group of chemicals commonly present in products used indoors is the polybrominated diphenyl ethers (PBDEs), which are flame retardants applied in plastics, electronics, some textiles and upholstered furniture. PentaBDE and octaBDE were banned from use in the EU in 2004 (EU, 2003) and were included in the Stockholm Convention in 2009 (UNEP, 2009) and are thus subject to a global ban. DecaBDE was about to be phased out from electrical and electronical (E&E) equipment under the Restriction of Hazardous Substances Directive (RoHS) in 2006 but was exempted from the directive (EU, 2005). However, due to a decision in the European Court of Justice it was banned for use in E&E products in 2008 (EU, 2008). As a result, the only BDE mixture that is still allowed is the decabrominated product, where the congener BDE 209 represents about 98% of the content and only in products other than E&Es. Despite the restrictions, several other BDE congeners are also still present in consumer products within the technosphere because of the long life-times of the products. Apart from contributing to human exposure, chemicals emitted in the indoor environment may also be transported further to the outdoor environment via ventilation or sewage systems (Cousins, 2012). This makes the indoor environment a central mediator for chemical transport into the outdoor environment. It is of particular importance in densely populated urban areas where the stock of consumer articles is high. A recent study measured outgoing amounts of polybrominated diphenyl ethers (PBDEs) with ventilation outlets and concluded that the indoor environment may be an important source of PBDEs to urban air (Björklund et al., 2012b). The impact of indoor releases on the concentration of organic chemicals in the urban environment was investigated in a recent study by Cousins (2012) using a chemical fate model that included the indoor environment (the SMURF model). The model was applied to four organic chemicals with widely differing partitioning properties using illustrative emissions. The simulations indicated that the steady state mass and the environmental residence time of the four chemicals in urban Stockholm increased by a factor of

1.1–22 when emissions were assigned to indoor air rather than to outdoor air. In addition, emissions indoors were shown to result in lower outdoor environmental concentrations of semivolatile, hydrophobic organic chemicals compared to if these emissions had occurred to outdoor air, because of chemical removal pathways in the indoor environment. The aims of this study were to investigate whether emissions from certain consumer products can explain the observed concentrations in indoor air and dust of DEHP, DINP and BDE 209, and to assess the impact of indoor emissions on urban air quality for these chemicals. To this end, emission rates of DINP from PVC wall and flooring material were measured and used as input to the SMURF model (Cousins, 2012), together with emissions for DEHP and BDE 209 based on literature data (see Section 2.2). Monitoring data from the literature were used to estimate background concentrations in inflowing air to urban Stockholm as a boundary condition for the model. Concentrations of DEHP and DINP in urban air and ventilation outlets were measured to provide additional data for model evaluation. Model predictions of chemical concentrations in indoor air and dust and outdoor air were compared to monitoring data. An uncertainty analysis was conducted to assess the influence of individual input parameters on the variance in model output. 2. Material and methods 2.1. The SMURF model The SMURF model (Cousins, 2012) is an eight-compartment (air (1), water (2), soil (3), sediment (4), urban surfaces (5), indoor air (6), indoor vertical surfaces plus ceilings (7) and indoor upward facing horizontal surfaces (i.e. floors) (8)), steady-state non-equilibrium chemical fate model parameterized to the municipality of Stockholm, Sweden. It comprises 46 environmental variables, 7 physical–chemical property input variables, 23 kinetic variables to describe transport between environmental compartments and five variables describing environmental degradation. Degradation in both indoor and outdoor air is described by a second order OH radical reaction rate constant applied to different concentrations of OH radicals for indoor and outdoor air. The indoor surface compartments consist of a 10 nm thick organic film layer covering all surfaces as suggested by Zhang et al (2009). In addition to the film, the horizontal compartment also includes an additional dry dust layer of 8.5 × 10−5 kg m−2 as suggested by Bennett and Furtaw (2004), whereas vertical surfaces and ceilings are represented by the film only. Degradation on indoor surfaces is

Table 1 Physical–chemical properties of target substances and confidence factors assigned. Property −1

Molecular weight (g mol ) Water solubility (g m−3) Vapour pressure (Pa) LogKOW Melting point (°C) Enthalpy of solution (J mol−1) Enthalpy of vaporization (J mol−1)c Degradation ratesd Second order atmospheric rate constant, kOH (cm3 molecule−1 s−1) Water half-life (h) Soil half-life (h) Sediment half-life (h) Film/air reactivity factore Concentration in ambient air (ng m−3)f a b c d e f

DEHPa

DINPa

BDE209b

Assigned CF

390.6 2.5 × 10−3 2.5 × 10−5 7.7 −46 −10000 108000

418.6 3.1 × 10−4 6.8 × 10−6 8.6 −48 −10000 113000

959.2 2.7 × 10−4 3.2 × 10−9 9.1 300 −25000 140000

1 5 5 1.1 1.05 1.1 1.1

2.2 × 10−11 360 720 3240 45 0.84

2.3 × 10−11 900 1800 8100 45 0.25

3.4 × 10−14 4320 8640 38900 45 0.00012

3 5 5 5 5 5

From Cousins et al. (2003) unless otherwise noted. From Cousins and Palm (2003) unless otherwise noted. Estimated from vapour pressure using Trouton's rule according to MacLeod et al. (2007). Estimated using EPISuite Estimation software (USEPA, 2011). (Cousins, 2012). Weighted median from remote air samples presented in Fig. 2.

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estimated using an air/film reaction ratio of 45 (Cousins, 2012). The parameterization of the SMURF model was described in detail by Cousins (2012). In addition to environmental parameters and substance property input data, the SMURF model requires data on emissions to one or more of the compartments indoor air, outdoor air, surface water and soil, as well as concentrations in incoming air and water. For the current study, environmental and kinetic variables were the same as in Cousins (2012), and are given in Table S1 in the supporting information (SI) to this paper. Physical–chemical properties for DEHP, DINP and BDE 209 used as input are given in Table 1, together with confidence factors (CFs) assigned to account for uncertainty (see further description in Section 2.1.2). Emissions to indoor air were estimated according to the methodology described in Section 2.2. Since the current study aimed to assess the importance of indoor emissions for outdoor air pollution, the background concentration in inflowing air is a crucial model input parameter. A careful review of available literature on remote concentrations of the target substances was therefore undertaken to determine relevant concentrations in inflowing air (see the SI). Remote locations were selected since semi-urban levels from Stockholm or other Swedish locations were not available, and the body of literature on remote levels is extensive and more comparable than e.g. semi-urban levels from other countries where chemical use patterns may vary substantially. Because of the focus on indoor and outdoor air, background concentration with inflowing water as well as emissions to water and soil were all assumed to be zero. To highlight the contribution from the indoor environment specifically, emissions to outdoor air from other sources within the city were also set to zero. Model predicted concentrations of DEHP, DINP and BDE 209 in indoor air, outdoor air and indoor floor dust were compared to monitoring data derived from literature and from new measurements performed in the current study (Section 2.3). For DEHP and BDE 209, adjusted emission scenarios were also modelled, where predicted concentrations were fitted to indoor monitoring data. 2.1.1. Background concentrations of phthalates and BDE 209 in air A literature review was undertaken, compiling relevant studies on remote concentrations of DEHP, DINP and BDE 209 in the atmosphere. The results are summarized in Fig. 2. The number of samples and the reporting format of the data varied between the studies. Therefore, the median concentrations obtained in the different studies were used to calculate a weighted median concentration to be used as model input, taking into account the number of samples included in each study: xweighted ¼

Xm

Xm

xn= i¼1 i i

n i¼1 i

ð1Þ

where xi is the median concentration in study i, ni is the number of samples included in study i and m is the total number of studies considered. This resulted in median background concentrations of 0.84 ng m−3 and 0.12 pg m−3 for DEHP and BDE 209 respectively. For DINP, only one study (Cousins et al., 2007) has reported remote air concentrations (b 0.25–0.48 ng m−3, n = 3). The median value of 0.25 ngm−3 from that study was used as the background concentration. Because of the scarcity, alternatively, large variability in background concentrations in combination with the uncertainty associated with the applicability of remote concentrations as input on the regional scale a confidence factor of 5 was assumed for all three substances(see further 2.1.2 and the supporting information). More details on the determination of background concentrations are presented in the SI. 2.1.2. Uncertainty analysis A preliminary uncertainty analysis was performed to investigate the influence of uncertainty and/or variability in individual input parameters on the variance in the target model outputs. The true uncertainty and shape of distribution of these input variables are not known and had to be estimated. MacLeod et al. (2002) proposed a

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screening level uncertainty analysis methodology suitable for application to multimedia mass balance models. They recommended that one assumes a log normal distribution of input variables when the true distribution is unknown. The uncertainties in all 83 input variables were described using 95% confidence factors (CFs) ranging from 1 to 5, based on a combination of known and estimated uncertainties. In brief, the uncertainty analysis was performed in two steps: first, the sensitivity of each output Oj to changes in individual input parameters Ii by 0.1% was monitored:   Sij ¼ ΔO j =O j =ðΔIi =Ii Þ:

ð2Þ

Then, for each output, confidence factors due to uncertainty in all n input parameters were calculated as suggested by MacLeod et al. (2002):   1 2 2 2 2 2 1=2 : ð3Þ CFO ¼ exp SI1 ðLn CFI1 Þ þ SI2 ðLn CFI2 Þ þ … þ SIn ðLn CFIn Þ These confidence factors were applied to the model results to illustrate the variance in model outputs due to the assumed variability in all inputs. Finally, the contribution of individual input parameters i to the variance in output was calculated:   2 2 n 2 2 ðLnCFIi Þ SIi =∑k¼1 ðLnCFIk Þ SIk :

ð4Þ

Details about the methodology, its limitations, and the individual CFs applied are given in the SI. 2.2. Emission estimates 2.2.1. BDE 209 The use of BDE 209 in Sweden was estimated based on the approach of Palm et al. (2004), where PBDE consumption was derived from global statistics on bromine production, fraction used for BFRs and individual technical components, geographical market shares and national GDP, taking into account stored amounts in treated technosphere products. 2006 was chosen as a reference year since the majority of the data used for model verification later in this paper were generated in that year. It is assumed that the majority of the stored amounts of decaBDE this year occurred in electrical and electronical (E&E) equipment, considering that this was before the EU ban of the substance in E&E products, and the long life-time of such products, thus emission factors for electronical products (Hirai et al., 2006; Sakai et al., 2006) were used. Emissions were scaled to the Stockholm level through the multiplication of a Stockholm/Sweden population ratio of 0.09. 2.2.2. Phthalate esters Emissions of phthalate esters to indoor air (Ei) were estimated according to:       −1 −1 −1 2 ¼ SERA;i g m h  86:4  A m Ei kg year

ð5Þ

where Ei represents the annual emission of the target chemical, SERA,i is the area specific emission rate and A is the total surface area of material treated with the target chemical, and 86.4 is a unit conversion factor. The area (A) was estimated based on information on the total surface area of DEHP- and DINP-treated floors sold in Sweden between 1990 and 2010, to be 1.3 × 108 (±3.1 × 107) m2 for DEHP and 5.0 × 107 (± 3.0 × 105) m2 for DINP (Wessen (2010), details in the SI). SERA-values of DEHP from PVC floors have previously been determined using different types of emission chamber devices at temperatures ranging from 22 to 61 °C (Afshari et al., 2004; Clausen et al., 2004, 2007, 2010, 2011) for several materials, but with different start concentrations of phthalates than those present in Swedish vinyl materials. Therefore,

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these values were re-calculated using the relevant start concentrations (see equation S1 in the SI). To our knowledge, only one study has measured emissions of DINP from PVC to air (Schossler et al., 2011), with a reported SERA of 0.22μgm−2 h−1 at 23°C. Since the PVC materials available on the Swedish market differ from those tested in Schossler et al. (2011) (different material layers in vinyl flooring and wall paper including e.g. surface treatment with polyurethane), new emission measurements were performed to determine emissions of DINP from PVC that are more representative for Sweden (see 2.2.3). Area specific emission rates were then determined as described in the SI (equation S1). The emissions were scaled to Stockholm by assuming a linear relationship between PVC use and population (i.e. multiplying the Swedish emissions by the Stockholm/Sweden population ratio of 0.09).

an older building with potentially higher PVC content (many apartments equipped with vinyl flooring), while the other, situated in Hammarby Sjöstad just south of the city centre, was a modern building with low content of PVC material according to the local building caretaker. Apart from building materials, no information was available on the presence of other products containing PVC (e.g. toys) in the building. The samples were collected with a high-volume air sampler equipped with a glass fibre filter (pre-cleaned by heating to 400 °C) and a glass column packed with polyurethane foam (PUF) (pre-cleaned through Soxhlet extraction with acetone). After sampling the filters and columns were wrapped in aluminium foil and stored at −18 °C until analysis. Details of the sampling programme are given in the SI. 2.4. Chemical analysis

2.2.3. Emission tests Emissions of DINP from vinyl floors and wallpaper available on the Swedish market (treated with the DINP product CAS 28553-12-0) were measured employing a 1 m3 stainless steel emission chamber, with a loading factor of 1.06 m2 m−3 and an air exchange rate of 0.48 h−1. Prior to the emission tests, the flooring and wall paper material were analysed for their phthalate content to determine the start concentration and to assure their representativeness. The analytical procedures and the details on the emission tests are described in the SI. In brief, vinyl pieces of size 125 × 85 cm were placed into the chamber at 40 °C and emitted amounts were sampled by passing air through a 1 m long cooled glass tube, followed by two PUF-filters. Sampling was carried out for 10 days with 24 h sampling periods at the same flow rate (8 L min−1). The test piece was then removed and the chamber was heated to 300 °C, followed by renewed sampling to collect analytes adsorbed to the chamber walls. Sample analysis is described in 2.4. 2.3. Monitoring of ventilation outlets and outdoor air Outdoor air samples were collected in urban Stockholm in 2008 and 2009 during autumn, winter and spring. Samples were collected at rooftop level and at street level. Ventilation outlets were sampled in two multi-storey apartment buildings of different age. One, which was situated in Häggvik, a northwestern suburb of Stockholm, represented

The methods used to analyse the PVC material, emission chamber air samples and air samples are illustrated in Fig. 1. More details are given in the SI. 2.4.1. Quality control Quality control routines were implemented to minimise the risk for sample contamination. These included i) immediate (b 10 min) sample extraction after emission tests, ii) avoiding the use of personal care products and any plastic materials during sample handling, with the exception of phthalate-free nitrile gloves, iii) heating of all equipment and chemicals used to 300 °C or 400 °C prior to sampling and analysis, iv) protection of the Soxhlet apparatus using an activated charcoal column, v) protection of all sampling jars, tubes and columns using aluminium foil, vi) using only high-quality solvents (see SI). The solvents were pre-checked for phthalate contamination and exclusively employed for phthalate analysis. Method blanks and field blanks (filters and PUFs) were included for each environmental sample batch, and taken before and after the emission tests. The limits of detection (LOD) and quantification (LOQ), which were defined as 3 times and 10 times the standard deviation of the blanks, were 0.5 (LOD) and 1.7 (LOQ) ng m−3 for DEHP and to 2 (LOD) and 6.7 (LOQ) ng m−3 for DINP. The analytical results were blank corrected (Keith, 1991; Miller and Miller, 1993). Details about the quality criteria employed for identification and quantification of the target compounds are described in the SI.

Fig. 1. Sample extraction and clean-up methods for a) vinyl material, b) emission chamber air samples and c) air and ventilation samples.

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3. Results and discussion 3.1. Emission estimates The use of BDE 209 in Sweden in 2006 was estimated to about 300 tonnes, which corresponds to a use of 27 tonnes in Stockholm, scaled based on population. Applying the experimentally determined emission rates of 2 × 10−7–4.8 × 10−6 g g−1 used (Hirai et al., 2006; Sakai et al., 2006), and a confidence factor of 5 (see Section 2.1.2), emissions were estimated to be 0.054–0.13 kg year−1. A mid-point value of that was used as the median emission, i.e. 0.068 kg year−1 to Stockholm indoor air. The average concentration of DINP in the analysed PVC material samples was 13 ± 1.1% (w w−1) (Table 2), which is consistent with the DINP content of 16 ± 3.5% (w w−1) as reported in the floor companies' product declarations (see Section 2.2.2). We therefore consider the tested materials to be representative of Swedish DINPcontaining PVC floors. Since DEHP was replaced by DINP, it is assumed that the DEHP content was similar in PVC floors. Afshari et al. (2004) reported DEHP contents in PVC floors of 17–18% (w w−1), which supports this assumption. The results from the emission tests are presented in Table 2. The empirical SERA-values (i.e. ṁ(t)) obtained from the tests were in the same order of magnitude but slightly higher than the value reported by Schossler et al. (2011), which may be attributable to differences in emission behaviour due to differences in material composition. The fraction sorbed to the chamber surfaces was 76% on the average. The material-air partition coefficients (KMA) for 23 °C determined according to equations S2–S4 in the SI were 8.1 × 1011 for DEHP and 5.8 × 1012 for DINP. Consequently, calculated SERA-values for 23 °C for DEHP and DINP (eq S1) were 0.49 and 0.067 μg m−2 h−1, respectively. Multiplying these SERA-values by the total estimated surface area of DEHP and DINP treated floors in Stockholm in 2010 (see 2.2 and SI) (1.2 × 107 m2 and 4.5 × 106 m2, respectively) generated estimated annual emissions to Stockholm indoor air of 44 (14–130) kg for DEHP and 2.9 (0.97–8.7) kg for DINP. 3.2. Concentrations of phthalates in outdoor urban air and ventilation outlets Fig. 2 shows the summary statistics from the measurements of phthalates in the outflowing air from building ventilation systems and in urban air together with previously published data. The results for the individual samples are presented in Table S3 in the SI. The median concentration in air sampled at street level was 13 ng m−3 for DEHP and 12 ng m−3 for DINP. In air sampled on rooftops, the corresponding median concentrations were 10 ng m−3 and 15 ng m−3. The DEHP concentrations were similar in magnitude to the concentrations observed in Paris, France (5.8–36ngm−3; Teil et al. (2006)), but showed a larger variation, covering about three orders of magnitude. For DINP, no monitoring studies were found for comparison, but the observed concentrations were about 100 times lower than estimated urban air concentrations in the EU risk assessment of DINP (ECB, 2003), and at least a factor of 100 higher than measured concentrations in the few

Table 2 Results from the emission measurements of DINP. Substance

T (K)

C0 (%)

m˙ðt Þ a (μg m−2)

hmb (m s−1)

Density (kg m−3)

KMAc

PVC floor 1 PVC floor 2 PVC wall cover

320.5 320.5 320.5

14 13 11

0.37 0.52 0.62

9.05 × 10−4 7.88 × 10−4 9.05 × 10−4

1500 1500 1500

1.85 × 1012 1.06 × 1012 8.67 × 1011

a

The emitted amounts include amounts sorbed to chamber walls. hm was calculated as described in Holmgren et al. (2012), using the chamber airflow of 8 L min−1. c Calculated according to equation S2 in the supporting information. b

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samples of remote air from the Swedish west coast. Concentrations in ventilation outlet air were similar in the old and new building, and varied between 2.7 and 23 ng m−3 for DEHP and between b2.0 and 4.6 ng m−3 for DINP. The concentrations for DINP were all below the LOQ. In the case of DEHP, the median concentrations in ventilation outlets were a factor of 10–60 lower than concentrations in indoor air in Stockholm buildings reported by Bergh et al. (2011a, 2011b). The reason for this difference is unclear, but could be due to variation between buildings, or a systematic error in one of the studies. The only data on DINP in indoor air found in the literature were from a study of 39 dwellings in Sapporo, Japan where concentrations ranged from b65 to 192 ngm−3 (Kanazawa et al., 2010), but the high detection limit in their study prevents a fair comparison with our data. 3.3. Fate modelling 3.3.1. Comparison between predicted and measured concentrations Fig. 2 shows the model predicted concentrations of the chemicals in outdoor urban air (COA), indoor air (CIA) and indoor dust (CDUST) (red square markers). For the phthalates, the predicted concentrations in indoor air and dust deviate by less than an order of magnitude from the median measured concentrations. Considering the large variability in the monitoring data, this degree of agreement is considered to be good. Notably, the model predicted indoor air concentration range of DINP (0.4–8.4 ng m−3) agrees with the measured concentrations in ventilation outlets (b 2.0–4.6 ng m−3, see Fig. 2 and SI), whereas the predicted concentration range of DEHP in indoor air (7.2–250 ng m−3) lies between but overlaps with the measured concentration ranges in ventilation outlets (2.7–23 ng m−3, see Fig. 2 and SI) and indoor air (15–890 ng m−3, Fig. 2). Due to the lack of representative indoor air data of DINP it was not possible to evaluate how model predicted concentrations compare with measured concentrations in indoor air. For outdoor air, The SMURF model underpredicts the median concentration of the phthalates by a factor of 2–16, albeit the modelled concentration of DEHP is still within the measured range. Judging from Fig. 2, it appears that urban air levels are highly variable, but the median concentrations from the most extensive studies included in Fig. 2 (the current study and Teil et al. (2006)) are around 10 ng m−3 rather than around 1 ng m−3 as predicted by the SMURF model. The assigned background concentration of DEHP is believed to be representative for remote conditions (see 2.1.1), but regional background concentrations in air just outside the Stockholm borders could potentially be higher. No studies comparing remote, suburban and urban air concentrations of the target chemicals have been found in the literature. Considering the scarcity of monitoring data for DINP in Swedish background air, it is possible that the concentration of this substance in inflowing air is underestimated. Another possibility for the discrepancy between modelled and measured concentrations of phthalates in urban outdoor air may be the existence of additional emission sources in the outdoor environment that could contribute to the atmospheric levels of phthalate esters in Stockholm. Outdoor uses identified in the EU risk assessment document for DEHP include e.g. car undercoating, roofing material, and cables and wires exposed to air (ECB, 2008). PlasticsEurope (2012) reports that building materials account for 50% of the total PVC usage, of which floors and walls account for a fraction. Thus it seems likely that other sources also contribute to the release of phthalate esters in urban areas. To investigate the consequences for the fate of phthalates for an optimized scenario, the indoor emission rates were adjusted to fit the model predictions to the measured concentrations in indoor air and dust for DEHP (Fig. 2, green triangular markers). This resulted in emissions of 250 (83–750) kg year−1 for DEHP or an increase by a factor of 5. As shown in the figure, the effect on outdoor air concentrations is minor but visible (increase by a factor of 1.2). For BDE 209, the predicted concentrations were within the measured ranges both for indoor air (15–55 pg m−3 vs 0.7–23000 pg m−3) and indoor dust (219–1770 ng g−1 vs 50–100000 ng g−1) although the

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Fig. 2. Model predicted and measured concentrations of DEHP, DINP and BDE 209 in indoor air, outdoor air and indoor dust in Stockholm and other locations for comparison. The red square markers represent the original emission scenario, the green triangles represent the adjusted scenario. For the modelled concentrations, the error bars represent the 95% confidence interval. For the measured concentrations, the error bars show the minimum and maximum values. Letters within brackets refer to: (a) Cousins et al. (2007), (b) Xie et al. (2007), (c) Xie et al. (2005), (d) This study, (e) Teil et al. (2006), (f) Bergh et al. (2011b), (g) Bergh et al. (2011a), (h) Santillo et al. (2003), (i) Bornehag et al. (2005), (j) SEPA (2012), (k) Egebäck et al. (2012), (l) Su et al. (2007), (m) Cetin and Odabasi (2008), (n) Castro-Jiménez et al. (2011), (o) Uhl and Gans (2005), (p) Björklund et al. (2012b), (q) Thuresson et al. (2012), (r) Björklund et al. (2012a), and (s) Nagorka et al. (2010).

variability in the measured data exceeded the predicted variance in the model output (Fig. 2). The model predicted flow from indoor to outdoor air was 0.026 (0.009–0.076) kg year−1. Björklund et al (2012b) estimated BDE 209 releases from indoor air to outdoor air in Sweden based on measured concentrations in ventilation outlets in Stockholm

to 0.8–32 kg year−1. Using the Stockholm/Sweden population ratio of 0.09, this corresponds to 0.07–2.9 kg year−1 for Stockholm, which overlaps the model estimated range here. The differences are partly explained by the fact that Björklund et al. (2012b) applied higher ventilation rates to estimate outflows from public buildings (resulting

100% DEHP

DINP

BDE 209

DEHP

DINP

BDE 209

DEHP

DINP

BDE 209

90% 80% 70% 60% 50% 40% 30% 20% 10% 0% Coa

Cia

Cdust

Fig. 3. Percentage contribution of model variables to variance in concentrations in outdoor air (COA), indoor air (CIA) and indoor floor dust (CDUST) for DEHP, DINP and BDE 209.

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in higher chemical outflows) and partly by the large variability in indoor air concentrations (see Fig. 2) with higher concentrations in offices and public buildings than those estimated here. The predicted concentration of BDE 209 in outdoor air was 0.15 pg m− 3 i.e. 25% higher than the concentration in the inflowing air. No Swedish urban air data are available for BDE 209 but a recent review on global PBDE air levels suggests that BDE 209 accounted for 57–94% of total PBDEs in European urban air samples, where ∑ PBDE levels ranged from 6 to 158 pg m− 3 (Besis and Samara, 2012), which corresponds to a range of 3.4–148 pg m− 3, which confirms that urban air concentrations are higher than concentrations in remote air. 3.3.2. Uncertainty analysis The percentage contribution of key model inputs to the variance in the target outputs for the three chemicals is shown in Fig. 3. For DEHP and BDE 209, the results for the adjusted emission scenarios are presented. For DINP, the variance in outdoor air concentration (COA) is almost entirely dominated by the inflowing air concentration, implying that the changes in indoor emissions within the specified range make a negligible contribution to the variance in outdoor air concentration. For DEHP, indoor emissions contribute about 6% to the variance in COA, which is a result of the high absolute emission values in the adjusted scenario for DEHP. The most influential input parameters on the variance in indoor air concentration (CIA) are similar for the two phthalates and dominated by emission to indoor air and ventilation flow, followed by the total floor area and the physico-chemical properties vapour pressure, aqueous solubility and logKOW. The variance in the predicted concentration in floor dust (CDUST) is influenced by multiple input parameters, the most obvious being emissions to indoor air. For DEHP this is followed by the physical–chemical properties and ventilation flow, and thereafter by similar contributions from floor area, indoor film thickness, horizontal wet removal rate and film particle density. For DINP, the pattern is similar but with a larger influence of horizontal wet removal rate and smaller influence of ventilation flow and physico-chemical properties compared to DEHP. For BDE 209, indoor emissions contribute about 7% to the variance in the predicted COA. Wind speed, ventilation flow and indoor vertical deposition rate also make small but visible contributions to the variance in COA. The importance of wind speed is explained by the fact that air advection is the dominant removal process of BDE 209 from the urban air compartment; therefore a higher wind speed will dilute the impact of BDE 209 emissions to outdoor air within the model domain. The input parameters with the most influence on the variance in CIA are emission to indoor air, indoor vertical deposition rate, ventilation flow and floor area. The 13% contribution of indoor vertical deposition rate reflects the strong capacity of BDE 209 to partition to particles, which makes particle behaviour central for the fate of BDE 209. LogKOW has negligible influence on the variance in CIA for BDE 209. BDE 209 is sufficiently lipophilic that it is still almost exclusively associated with dust in the model, even when its KOW is decreased somewhat. Since KOW is used in the derivation of other partition coefficients such as KOA, the effect or no-effect of KOW on model output variance may indirectly be a result of the impact of e.g. the KOA, which influences the air-surface partitioning. Similar to the phthalates, the variance in CDUST is influenced by multiple input parameters, the most dominant being emissions to indoor air, followed by wet removal rate, indoor film thickness, and equal contributions from ventilation flow, film particles density, floor area, indoor air particles concentration and mass fraction particles in indoor film. 3.3.3. Contribution of indoor air emissions to overall urban flows The potential influence of indoor releases on the overall flows in the urban environment was evaluated by plotting model predicted mass balances for the indoor environment and the outdoor air compartment

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for the three substances (Fig. 4). For DEHP, only mass balances for the adjusted scenario are shown. The figure shows that for DINP, the contribution of the indoor environment is insignificant for the outdoor air compartment and urban air outflows (dark blue bar) are predicted to be lower than background inflows (light grey bar). The reasons for this are the relatively high reactivity of DINP in outdoor air and the relatively small contribution from indoor air, implying that Stockholm would act as a sink for DINP. For DEHP the predicted urban air outflow exceeds the background inflow and releases from the indoor environment are predicted to contribute by 21% to the outflows. However, bearing in mind that the model under-predicts the measured concentrations of DEHP in outdoor air, it is likely that the true contribution from the indoor environment is lower. The indoor–outdoor transport is predicted to exceed that of indoor removal for the phthalates. For BDE 209 the predicted urban air outflow clearly exceeds the background inflow, and releases from the indoor environment are predicted to contribute 27% to the outflow. Björklund et al. (2012b) estimated that release from the indoor environment contributes about 81–82% of the total emissions of BDE 209 to outdoor air in Sweden. Background inflow with air due to long-range atmospheric transport was however not included in that study. The removal of chemical via indoor processes is predicted to be higher than the transport to the outdoor environment for BDE 209. These indoor removal processes include removal of dust and indoor organic film via different cleaning processes (e.g. vacuuming and wet mopping). The greater importance of indoor removal for BDE 209 relative to the phthalates is explained by the higher fugacity capacity of indoor surfaces for BDE 209 (due to its high KOA), leading to more efficient removal from air to surfaces, whereas the phthalates have a weaker tendency to partition to surfaces, which enables them to be transported to urban air to a larger extent. 4. Conclusions The current study provides two estimates of emissions of DEHP to indoor air in Stockholm: 44 versus 250 kg year−1, where the latter is believed to be the most accurate. The estimated emission of DINP of 2.9 kg year−1 is uncertain due to the scarcity of data for model evaluation, but is believed to be in the right order of magnitude considering the fair agreement observed with the few data that exist. The study further indicates that indoor emissions alone cannot explain the levels of phthalates found in outdoor urban air. The model would require additional emissions of about 9 tons year−1 and 8 tons year−1 to outdoor air in Stockholm to reproduce the observed concentrations in outdoor air of DEHP and DINP, respectively. This suggests that i) concentrations in background inflowing air have been underestimated or that ii) phthalate containing products in the outdoor environment (e.g. car interiors, PVC roofing materials, tarpaulins, cables and wires etc.) are more important than PVC-products in the indoor environment for the occurrence of these substances in outdoor air. Thuren and Larsson (1990) suggested an annual deposition of DEHP of 130 tons year−1 in Sweden and commented that this corresponded well to the emissions. Scaling from Sweden to Stockholm according to population, this would correspond to 12 tons year−1 in Stockholm. Thus, 9 tons year−1 may not be unrealistic. With such a scenario, the contribution of the indoor releases would be insignificant for DEHP as well. To address the possibility of underestimated inflowing concentrations, it would be desirable to conduct an upwind and downwind atmospheric monitoring study in the region. Furthermore, we suggest that regional background levels be used as input to urban fate models, rather than remote concentrations. In the case of BDE 209, the estimated emissions of 0.07 (0.01–0.34) kg year−1 are judged to be in the right order of magnitude, indicating contribution from the indoor environment to the total urban outflow of approximately 25%. The scarcity of relevant urban air monitoring data hinders a direct comparison, but the uncertainty analysis revealed

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Annual transport (kg)

1.00E+03

DEHP

IN

8.00E+02 6.00E+02 4.00E+02

IN

OUT

2.00E+02

DINP

IN

OUT

2.00E+02 1.50E+02 1.00E+02 5.00E+01

IN

0.00E+00

OUT

0.00E+00

INDOORS 1.40E-01

Annual transport (kg)

2.50E+02

OUT

Annual transport (kg)

534

BDE 209

OUTDOOR AIR IN

INDOORS

OUTDOOR AIR

OUT

1.20E-01 1.00E-01 8.00E-02

IN

OUT

6.00E-02 4.00E-02 2.00E-02 0.00E+00

INDOORS

OUTDOOR AIR

Fig. 4. Predicted annual chemical mass balances for the Stockholm indoor environment and outdoor air for DEHP (top left), DINP (top right) and BDE 209 (bottom). Mass balances for DEHP and BDE 209 represent predicted flows for the adjusted emission scenario.

that emissions indoors contribute about 6% to the variance in outdoor air concentration. Altogether, the current study illustrates that the occurrence of the three target chemicals in outdoor urban air is strongly influenced by the background atmospheric inflow, and that uncertainties associated with the background air concentration contribute most to the variance in outdoor air concentration, other outdoor sources being neglected. Therefore, remote air concentrations are likely insufficient to describe the chemical inflow to urban areas and future modelling studies should aim for the use of regional, sub-urban air concentrations as model input. For BDE 209, the assessment indicates a potential of the Stockholm indoor environment to contribute to the urban air pollution, but additional monitoring data for urban air are required to confirm this conclusion. Acknowledgements The authors would like to thank three floor companies for providing test material for emission studies, Per Wiklund at IVL for sampling and Eva Brorström-Lundén at IVL and Michael McLachlan at Stockholm University for useful comments on the manuscript. The study was financed partly by Miljöfonden, Sveriges Ingenjörer, and partly by the Swedish Environmental Protection Agency through the research programme ChemiTecs and through the chemical screening programme. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2013.10.023. References Afshari A, Gunnarsen L, Clausen PA, Hansen V. Emission of phthalates from PVC and other materials. Indoor Air 2004;14:120–8. Andersson P, Rännar S. A report on the initial procedure for identification of chemical/article/use combinations of concern, including the selected case-study chemicals, Umeå; 2009. Bennett DH, Furtaw EJ. Fugacity-based indoor residential pesticide fate model. Environ Sci Technol 2004;38:2142–52.

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