Next Article in Journal
Bacteria as Cloud Condensation Nuclei (CCN) in the Atmosphere
Next Article in Special Issue
Levels and Sources of Atmospheric Particle-Bound Mercury in Atmospheric Particulate Matter (PM10) at Several Sites of an Atlantic Coastal European Region
Previous Article in Journal
Radar Remote Sensing of Precipitation in High Mountains: Detection and Characterization of Melting Layer in the Grenoble Valley, French Alps
Previous Article in Special Issue
Urban Aerosol Particle Size Characterization in Eastern Mediterranean Conditions
Order Article Reprints
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:

Household Dust: Loadings and PM10-Bound Plasticizers and Polycyclic Aromatic Hydrocarbons

Centre for Environmental and Marine Studies (CESAM), Department of Environment, University of Aveiro, 3810-193 Aveiro, Portugal
Department of Physics, IMARENAB University of León, 24071 León, Spain
Institute of Environmental Assessment and Water Research (IDAEA-CSIC), 08034 Barcelona, Spain
Author to whom correspondence should be addressed.
Atmosphere 2019, 10(12), 785;
Received: 28 October 2019 / Revised: 3 December 2019 / Accepted: 3 December 2019 / Published: 6 December 2019
(This article belongs to the Special Issue Urban Atmospheric Aerosols: Sources, Analysis and Effects)


Residential dust is recognized as a major source of environmental contaminants, including polycyclic aromatic hydrocarbons (PAHs) and plasticizers, such as phthalic acid esters (PAEs). A sampling campaign was carried out to characterize the dust fraction of particulate matter with an aerodynamic diameter smaller than 10 µm (PM10), using an in situ resuspension chamber in three rooms (kitchen, living room, and bedroom) of four Spanish houses. Two samples per room were collected with, at least, a one-week interval. The PM10 samples were analyzed for their carbonaceous content by a thermo-optical technique and, after solvent extraction, for 20 PAHs, 8 PAEs and one non-phthalate plasticizer (DEHA) by gas chromatography-mass spectrometry. In general, higher dust loads were observed for parquet flooring as compared with tile. The highest dust loads were obtained for rugs. Total carbon accounted for 9.3 to 51 wt% of the PM10 mass. Plasticizer mass fractions varied from 5 µg g−1 to 17 mg g−1 PM10, whereas lower contributions were registered for PAHs (0.98 to 116 µg g−1). The plasticizer and PAH daily intakes for children and adults via dust ingestion were estimated to be three to four orders of magnitude higher than those via inhalation and dermal contact. The thoracic fraction of household dust was estimated to contribute to an excess of 7.2 to 14 per million people new cancer cases, which exceeds the acceptable risk of one per million.

Graphical Abstract

1. Introduction

In industrialized nations, people spend most of their time in closed environments, especially at home [1]. Walking induced particle resuspension has been reported to be an important source of indoor particulate matter [2,3,4]. Several factors affect resuspension of particles, including relative humidity, flooring type, and dust loadings [3]. Household dust is a complex mixture of particles of both indoor and outdoor origin, including organic, inorganic, and biological components, many of which are toxicants, carcinogens or allergens [5,6]. Its composition depends on numerous conditions, such as environmental and seasonal factors, ventilation and air filtration, homeowner activities, and in- and outdoor sources [7]. Several studies have shown that inhalation of dust particles is linked to an increased risk of a range of health hazards, spanning from asthma symptoms in susceptible adults and children [8,9,10,11] to cancer and fertility problems [12].
Residential dust is recognized as a major source of environmental contaminants, including polycyclic aromatic hydrocarbons (PAHs) and phthalic acid esters (PAEs) [13,14,15]. PAHs are primarily byproducts of incomplete combustion of fossil fuels and biomass and pyrosynthesis of organic materials [16]. Dust, especially in carpeted floor, can be a permanent reservoir for these chemicals, which may be inhaled through resuspension into air, ingested accidentally by children or absorbed through the skin [13,17,18]. Because of their widespread sources and strong carcinogenicity, cytotoxicity, mutagenicity, endocrine disrupting, and other hazardous properties, PAHs have been the focus of extensive attention by scientists and governmental organizations [7,18,19,20,21,22,23,24]. Benzo[a]pyrene (BaP), the most extensively studied carcinogenic PAH, is classified by the International Agency for Research on Cancer (IARC) as a Group 1 or known human carcinogen [25]. Four of the top ten priority pollutants, nominated by the Agency for Toxic Substances and Disease Registry (ATSDR) in 2011, are single PAHs or PAH mixtures (PAHs, BaP, benzo[b]fluoranthene, and dibenzo[a,h]anthracene) [26].
PAEs (also called phthalates) are used as plasticizers in several consumer products, commodities, and building materials [27]. Therefore, phthalates are ubiquitous in residential and occupational environments, where they are present in high concentrations, both in air and in dust [14,27,28,29,30,31,32,33,34,35,36,37,38,39,40,41]. Comparisons of mass fractions reported in different studies indicate that, for measurements conducted over the past decade, the levels of PAEs in indoor dust tend to be three to five orders of magnitude higher than those of PAHs [14]. Recent toxicological studies have proven the potential of PAEs to disturb the human hormonal system and human sexual development and reproduction [42,43]. Moreover, PAEs are suspected to trigger asthma and dermal diseases in children ([44], and references therein). An EU risk assessment classified bis(2-ethylhexyl)phthalate (DEHP), dibutyl phthalate (DBP), and benzyl butyl-phthalate (BBP) as hazardous substances in 2005, and has issued a directive to ban these materials from products, particularly toys and cosmetics [45].
Indoor aerosol sources can significantly contribute to the daily dose of particles deposited into the human respiratory system [46]. Since indoor dust contributes to human exposure, its resuspension rates and chemical composition should be evaluated. The selection of appropriate techniques to assess household dust loadings and composition is a major challenge since several methodologies have been employed, including passive (dust settling) and active techniques (surface wiping, press sampling, sweeping, and vacuuming) [6,47].
Studies on the concentration of contaminants in household dust have been focused on the analysis of the total mass or of sieved fractions, involving sizes of tens or hundreds of µm [17,29,30,32,35,37,39,40]. A summary of average PAH concentrations in settled house dust by country and year (data from 35 studies) can be found in a recent review article [48]. A major issue encountered when comparing these studies was the variability in both the sampling methods employed and the dust particle size fractions subjected to analysis. In 21 out of the 35 studies reviewed, the particle size cut-off points were either 150 μm or 63 μm. It has been suggested that particles >150 μm do not easily and efficiently adhere to hands or skin. Therefore, these sizes are less relevant when evaluating exposure via ingestion or dermal pathways [49]. Liu et al. [24] collected and sieved indoor dust into six size fractions from office and public microenvironments in Nanjing, China. Higher PAH concentrations, and consequently more health risks, were observed for the smallest particles (< 43 μm). In general, the common sampling procedures are affected by the loss of fine particles owing basically to the difficulties of collecting all deposited material and to the electrostatic adhesion of particles to brush hairs, vacuum cleaner bags, and sieve meshes. Considering the possibility that household dust is resuspendable and can become airborne, most methodologies have drawbacks when assessing human inhalation exposure. Such an approach requires measurements of the contaminant concentration in smaller particle sizes. The main goal of this study was to determine household dust loadings and the respective PAH and plasticizer levels in the thoracic fraction (<10 µm) of resuspendable material on the floor, which deposit anywhere within the lung airways. The in situ resuspension chamber was previously devised and successfully applied to collect the deposited particulate matter with an aerodynamic diameter smaller than 10 µm (PM10) from different road pavements [50,51], but it was the first time this active sampling methodology was used to collect settled thoracic particles directly from the floor in indoor environments. For road dust, this sampling technique was previously compared to the USEPA methodology (AP-42 documents), which is based on vacuuming or sweeping of surfaces. Then, the collected material passes through a 200-mesh sieve to determine surface silt loadings, and finally an empirical formula is used to estimate PM10 emission factors. Both methodologies provided very comparable results for resuspendable PM10 dust [52]. Toxic substances attached to inhaled particles capable of passing beyond the larynx, i.e., PM10, could lead to a series of respiratory and cardiovascular diseases, and increase the risks of cancer [53]. Thus, sampling the total concentration or coarser dust fractions most likely provides only a rough estimate of exposure. In addition to household dust loadings and PM10-bound chemical components, this paper also provides a relatively exhaustive review of literature values to show not only the order of magnitude of the levels, but also the difficulties in comparing results due to the lack of standard methodologies.

2. Methodologies

2.1. Sampling

To determine and characterize dust loadings, a sampling campaign was carried out in four different houses located in the Spanish city of León (Table 1). In each housing unit, three rooms were investigated, including the kitchen, the living room and a bedroom. In each room, two samples were collected with, at least, a one-week interval. Samples in each house were taken one or two days before weekly cleaning. For dust collection, an in situ resuspension chamber operating at an air flow rate of 25 L min−1 was used [51]. After vacuuming, PM10 was separated from the total dust through a Negretti stainless steel elutriation filter and collected onto 47 mm quartz fiber filters (Pallflex®, Ann Arbor, USA), while particles with aerodynamic diameter >10 μm were deposited in the methacrylate chamber and along the elutriation filter. Electrostatic adhesion could cause some losses of particles <10 µm. However, this loss is likely to be negligible with respect to losses of traditional sampling procedures (i.e., sweeping). In a previous work, the chamber sediments were brought to a laboratory, dried for 48 h at room temperature, sieved and, then, analyzed by means of an optical particle sizer with the aim of verifying the granulometry selection of the sampling system. Results showed that the fraction <10 µm was, on average, only 0.6% and 0.1% (in volume) of samples previously sieved at 250 µm and 63 µm, respectively [50].
Sampling was performed in surface areas of 1 m2 for 30 min. Two to three different square meters were sampled using the same filter in order to ensure enough particulate mass for the subsequent gravimetric and chemical analyzes. Since some compounds could derive from the resuspension chamber itself or from ambient air that enters the system and passes through the filter during dust collection from the floor, air samples were vacuumed through the same system after sampling in each room. These background air filters were also sampled for 30 min.

2.2. Household Dust Characterization

Due to the loss of a small fragment after sampling, the kitchen filter from house 2 obtained in the 2nd week was discarded because weighing led to a negative dust mass. After gravimetric determination, two punches (9 mm) of each filter were analyzed by a thermo-optical transmission technique to obtain the PM10 carbonaceous content (organic and elemental carbon, OC and EC). This method is based on the quantification of the CO2 released from the volatilization and oxidation of different carbon fractions under controlled heating by a non-dispersive infrared (NDIR) analyzer. The blackening of the filter is monitored using a laser beam and a photodetector, which enables separating the EC formed by pyrolysis [54]. The remaining portion of each filter was extracted by sonication for 15 min with three aliquots (25 mL each) of dichloromethane. After filtration, the solvent was concentrated in a TurboVap system from Biotage and evaporated to dryness by a gentle nitrogen stream. All the extracts were analyzed by gas chromatography-mass spectrometry (GC-MS) in a Shimadzu QP5050A equipped with a TRB-5MS 30 m × 0.25 mm × 0.25 μm column. The quantitative analysis was performed by single ion monitoring (SIM). Background air filters were analyzed in the same way as the samples to obtain blank-corrected results. Data were acquired in the electron impact (EI) mode (70 eV). The oven temperature program was as follows: 60 °C (1 min), 60 to 150 °C (10 °C min−1), 150 to 290 °C (5 °C min−1), 290 °C (30 min) and using helium as carrier gas at 1.2 mL min−1. For the quantification of PAHs, the following mixture of deuterated internal standards (IS) was used: 1,4-dichlorobenzene-d4, naphtalene-d8, acenaphthene-d10, phenanthrene-d10, chrysene-d12, and perylene-d12 (Supelco, St. Louis, USA). In the case of plasticizers, deuterated diethyl phthalate-3,4,5,6-d4 and bis(2-ethylhexyl)phthalate-3,4,5,6-d4 (Supelco, St. Louis, USA) were used as IS. Calibrations were performed with authentic standards (Sigma-Aldrich, St. Louis, USA) in eight different concentration levels.

2.3. Extraction Recoveries

To assess recoveries during extraction, prebaked blank filters were spiked with known amounts of standards, covering the concentration range commonly reported in the literature for household dust. The solutions of known concentrations were applied to the entire surface of each filter with Pasteur pipettes. The solvent was allowed to evaporate keeping the filters in a desiccator overnight. The filters were subjected to the same methodology of extraction and analysis used for the samples. Six prebaked blank filters were also extracted and analyzed. Four distinct concentrations were tested, in triplicate, and each extract was injected 3 times. The following overall recoveries (%) were obtained: diethyl phthalate 38.4 ± 8.05, di-n-butyl phthalate 113 ± 17.4, benzyl butyl phthalate 70.2 ± 12.7, bis(2-ethylhexyl) adipate 70.4 ± 12.9, bis(2-ethylhexyl) phthalate 108 ± 10.9, phenanthrene 53.0 ± 10.6, anthracene 67.1 ± 11.5, fluoranthene 84.7 ± 11.2, pyrene 81.9 ± 11.0, benzo[a]anthracene 97.2 ± 11.4, chrysene 81.9 ± 9.7, benzo[b]fluoranthene 106 ± 10.0, benzo[k]fluoranthene 108 ± 10.6, benzo[a]pyrene 93.4 ± 13.3, indeno [1,2,3-cd]pyrene 90.6 ± 15.8, dibenzo[a,h]anthracene 87.2 ± 16.7, and benzo[g,h,i]perylene 84.7 ± 16.5. The concentrations shown throughout the manuscript were not adjusted with these percent recoveries. Due to its high volatility and concentration variability in both dust and background air samples, naphthalene was excluded from quantification. Due to its vapor pressure of 0.087 mm Hg at 25 °C, naphthalene is sometimes considered a “borderline” volatile/semivolatile compound, since it may often be detected in both gas and particulate phases. Because of its tendency to sublimate, in ambient air, naphthalene is known to mainly exist in the vapor phase [55].

2.4. Health Risk Evaluation

Human exposure to plasticizers occurs via ingestion, inhalation, and dermal contact. The daily intake (DI) through different pathways were estimated by the following equations ([39], and references therein), where DIing, DIinh and, DIder are the daily intakes (ng kg−1 day−1) via dust ingestion, inhalation, and dermal contact, respectively:
DI ing   =   C dust   ×   IngR   ×   EF   ×   ED   ×   CF BW   ×   AT
DI inh   =   C dust   ×   InhR   ×   EF   ×   ED BW   ×   AT   ×   PEF
DI der   =   C dust   ×   SA   ×   AF dust   ×   ABS   ×   EF   ×   ED   ×   CF BW   ×   AT
Cdust represent the mean concentrations in dust (ng g−1); IngR is the ingestion rate of indoor dust (200 mg day−1 for children, 100 mg day−1 for adults); EF is the exposure frequency (180 days per year for both children and adults); ED is the exposure duration (6 years for children, 24 years for adults); CF is an unit conversion factor (10−3 g mg−1); BW is the body weight (15 kg for children, 70 kg for adults); AT is the average time (2190 days for children, 8760 days for adults); InhR is the inhalation rate (7.6 m3 day−1 for children, 12.8 m3 day−1 for adults); PEF is the particulate emission factor (1.36 × 106 m3 g−1); SA is the dermal exposure area (1150 cm2 for children, 2145 cm2 for adults); AFdust is the dust adherence factor (0.2 mg cm−2 day−1 for children, 0.07 mg cm−2 day−1 for adults); and ABS is the dermal adsorption fraction (0.001 for both children and adults, dimensionless).
Like plasticizers, PAHs can enter the body through ingestion (swallowing), inhalation (breathing), and skin contact. The carcinogenic potency of PAHs in indoor dust can be evaluated by comparing the carcinogenic potency of each PAH to that of BaP. The BaP carcinogenic equivalent concentration (BaPTEQ) is defined as follows:
where Ci is the concentration of each individual PAH and BaPTEF are toxic equivalency factors [24]. The incremental lifetime cancer risk (ILCR) is generally used to quantitatively estimate the exposure risks from the three exposure routes [24,56]:
ILCR ing   =   CS   ×   ( CSF ing BW 70 3 ) ×   IngR   ×   EF   ×   ED   ×   CF BW × AT
ILCR inh   =   CS   ×   ( CSF inh × BW 70 3 ) InhR   ×   EF   ×   ED BW × AT × PEF
ILCR der   =   CS   ×   ( CSF der × BW 70 3 ) SA   ×   AF dust   ×   ABS   ×   EF   ×   ED   ×   CF BW   ×   AT
where CS is the BaPTEQ concentration (mg kg−1); CSFing, CSFinh, and CSFder are carcinogenic slope factors of 7.3, 3.85, and 25 (mg kg−1 day−1)−1, respectively.
The risk associated with non-carcinogenic PAHs in household PM10 dust was estimated through the hazard quotient (HQ):
HQ = DInc/RfD
where DInc is the daily intake via ingestion of non-carcinogenic PAHs, obtained by Equation (1), and RfD is the reference dose. The oral ingestion RfD values were taken from Iwegbue et al. [57].

3. Results and Discussion

3.1. Dust Loadings

Huge differences in dust loadings between rugs and hard floorings were registered (Figure 1). Rugs represented the surface with the highest dust loadings. In general, higher dust loadings were observed for parquet flooring as compared with tile. Among the four dwellings, the highest dust loadings were obtained in the living room of a suburban family house with an open fireplace. Several factors can affect the surface dust loadings, such as walking, cleaning frequency, household materials, and indoor particle sources. Thus, the presence or absence of occupants can have a marked effect on resuspension levels. In house 1, for example, a very significant decrease in dust loadings from the first to the second week was registered, possibly due to the absence of owners for a few days. The highest mean value was observed in bedrooms (1745 µg PM10 m−2), followed by living rooms (785 µg PM10 m−2) and, finally, kitchens (297 µg PM10 m−2). The size distribution of aerosols emitted from cooking activities has been reported to be dominated by ultrafine particles, with modes generally in the range of 20–100 nm [58]. Thus, although cooking emissions can be high and contribute to the deposition and subsequent resuspension of dust, nanometer sizes represent a small fraction of the particulate mass. Dust loadings in the kitchen of house 3 were two to seven times higher than in the kitchens of other houses. This may be due to a higher utilization rate. In the other three residences the owners, in general, only prepare dinner. A higher mean value for bedrooms may be related to the presence of specific active sources or activities in this space of the house, such as making and unmaking the bed, dressing and undressing (which potentiates skin desquamation), existence of various rugs, use of cosmetics, etc. [59].
Using a mechanical resuspension device, Tian et al. [3] characterized walking-induced particle resuspension as a function of flooring type. Results showed that for particles at 0.4 to 3.0 µm, the difference in resuspension fraction between carpets and hard floorings was not significant. It was also found that for fine particles (0.4 to 3.0 µm), the difference in resuspension caused by flooring type is negligible, while for coarse particles (3.0 to 10 µm) carpets are associated with two to four times higher resuspended concentration in comparison with hard floorings. In fact, carpeted floors may contribute to significantly higher surface dust loadings and allergen concentrations than hard floors [60,61,62,63,64]. Roberts et al. [65] used a high-volume surface sampler to measure surface dust in carpets. Dust loadings ranged from 0.32 to 14.4 g m−2. Adgate et al. [66] collected bare floor and carpet dust samples in 216 Jersey City, New Jersey, homes using quantitative wipe and vacuum sampling techniques. Dust loadings varied from 0.05 and 7.0 g m−2 and from 0.3 and 99 g m−2 in the wipe floor and vacuum samples, respectively. It should be noted that dissimilar indoor dust sampling strategies (e.g., wipe versus vacuum methods) are used to measure loadings and the amounts of toxicants per unit area, which renders comparisons between studies difficult. Lioy et al. [67] found that while loadings were substantially greater with wipe sampling, metal concentrations within the dust samples were similar for both methods of sampling. Vacuum cleaner sampling has its own series of problems, especially the variability in design and efficiency, and likely does not retain particles below 10 µm [47]. Bai et al. [68] evaluated the following five methods of sampling lead-contaminated dust on carpets: (i) wipe, (ii) adhesive label, (iii) C18 sheet, (iv) vacuum, and (v) hand rinse. The wipe and vacuum methods showed the best reproducibility and correlation with other sampling techniques. The authors concluded that surface wipe sampling was the best method to measure accessible lead from carpets for exposure assessment, while vacuum sampling was most effective for providing information on total lead accumulation (long-term concentrations). In their review paper, Lioy et al. [6] stated, “although we have come a long way in determining the uses of house dust to identify sources of indoor contamination and to provide improved estimates of residential human exposure, one of the challenges still lies in the reliability of sampling techniques.” The instrument used in this study has the advantage of being a device capable of collecting dust particles below 10 µm, for gravimetric and chemical analysis. Sampling and analysis of higher dust particle sizes may only provide a crude estimate of inhalation exposure. However, validation and intercomparison studies are still needed, which is not easy, as the panoply of established methodologies does not match inhalable sizes.

3.2. Carbonaceous, Plasticizer, and PAH Particulate Mass Fractions

Total carbon accounted for 9.3 to 51 wt% of PM10 with the highest mass fractions recorded in dust samples collected in the city center apartment (Figure 2). More than 80% of the total carbonaceous matter was composed of OC, whereas in many samples the EC was too low or undetectable. The highest percentages of EC (10% to 20% of total carbon) were observed in kitchens, where there are sources of combustion. The high proportion of OC reflects the importance of a multitude of indoor sources that contribute to the organic carbonaceous component of household dust, including bacteria, skin flakes, cosmetics, cleaning products, cooking, paper and clothing fibers, microscopic specks of plastics, environmental contaminants brought on the soles of our shoes, etc. Our finding that a large portion of PM10 from indoor dust is composed of OC agrees with the results of Polidori et al. [69], who measured particulate OC and EC concentrations at 173 homes in the USA. These authors demonstrated that part of the OC can be secondarily formed in the indoor environment as a result of reactions involving gas-phase organic compounds emitted by cleaning products, air fresheners, and other sources.
Eight phthalate plasticizers (PAEs) and one non-phthalate plasticizer [bis(2-ethylhexyl) adipate, DEHA] were quantified in PM10 from settled house dust. PAEs included dimethyl phthalate (DMP), dimethylpropyl phthalate (DMPP), diethyl phthalate (DEP), diisobutyl phthalate (DIBP), di-n-butyl phthalate (DBP), di-n-hexylphthalate (DNHP), benzyl butyl phthalate (BBP), dicyclohexyl phthalate (DCHP), bis(2-ethylhexyl) phthalate (DEHP), di-n-octyl phthalate (DNOP), di-isononylphthalate (DINP), and di-isodecylphthalate (DIDP). The total mass fractions varied from 5 µg g−1 to 17 mg g−1 PM10 (Figure 3). As observed with dust loads, huge differences were observed from week to week and from home to home. The highest values were registered in the bedroom of the city center apartment and in the living room of the rural house. In addition to variability in sources, the concentrations of these compounds depend on the same factors already mentioned for the dust loads (e.g., ventilation, domestic routines, and cleaning activities). Widely scattered concentration levels were also documented in many previous works. For example, Kubwabo et al. [33] analyzed 17 PAEs in 126 Canadian household dust samples, evidencing the huge variability in spatial and temporal distribution of these compounds across different areas of the home, and thus the difficulty in predicting potential household exposures.
Most likely due to its high volatility, DMP was the compound with the lowest mass fractions (Table 2). On the other hand, DEHP, DNOP, and DBP were the major PAEs in household dust. While high DBP values were observed in all parquet floor bedrooms, only samples from two living rooms showed detectable masses. Bamai et al. [70] also associated higher DBP levels in floor dust with compressed wooden floor. This type of flooring is usually composed of thin pieces, which are glued together and covered with wax, paint, and sometimes flame retardants. The surface applied products (gloss agents, plastic additives, paint, and varnish) contain DBP [70]. From quantitative and qualitative emission data on phthalates from different materials, Afshari et al. [71] reported that polyolefin covered with wax for floor polishing increased DBP concentration in chamber air by two-fold. DBP is also employed as a coalescing aid in latex adhesives, as well as a plasticizer in cellulose plastics and a solvent for dyes [70]. Furthermore, DBP has been reported to be largely present in cosmetic and personal care products [72].
Although the concentration of DNOP in floor dust has been reported in only a very limited number of publications, the mass fractions of this study are higher than those described in the literature. There were no appreciable differences between the amounts found in PM10 of the various rooms of the houses. DNOP is used in carpet back coating, floor tile, and adhesives. It is also employed in cosmetics and pesticides [73]. DEHP was present at higher concentrations in the bedroom and living room samples. Although DEHP has been consistently described as one of the most abundant PAEs in settled dust, the levels have decreased over time, reflecting its phase-out in the EU [28]. In Europe, the use of DEHP decreased drastically in 2001 and has to a large extent been replaced by DINP and DIDP, with their longer chains and lower volatility [74]. DEHP has been used in numerous consumer products, children toys, medical devices, and building materials (e.g., vinyl flooring, furniture, paints, cables, wires, wall coverings, and packaging materials) [75]. Bamai et al. [70] and Bornehag et al. [34] associated polyvinyl chloride (PVC) flooring with DEHP levels in house dust. Kolarik et al. [76] documented higher concentrations of BBP, DNOP, and DEHP in indoor dust in homes where polishing agents were employed as compared with homes where such products were infrequently used or not used at all. As PVC flooring was not present in any of the houses for this study, other emission sources may have contributed to the DEHP levels. It has been shown that the source characteristics (surface area and material phase concentration of DEHP), as well as the external mass-transfer coefficient and ventilation rate, are important variables that influence the steady-state DEHP concentration and the resulting exposure [77]. DEHP and other PAEs are strongly sorbed to surfaces. A relatively small gas-phase concentration, such as 0.1 ppb, is enough for significant vapor transport of a PAE and its subsequent partitioning between the gas phase and indoor surfaces, including airborne particles and settled dust [78].
BBP was present in all samples at relatively similar concentrations regardless of room type. The values fall within the range reported for household dust from other regions. BBP is commonly employed as a plasticizer for vinyl foams, which are often used as floor tiles. Other uses are in artificial leather, paints, and adhesives. BBP was classified as toxic by the European Chemical Bureau, and therefore its use has decayed rapidly in the last decade [79].
No appreciable differences were found between DINP concentrations in samples from the various rooms of the houses. DIPD, instead, was present in higher amounts in living rooms and at lower levels in kitchens. The values observed in this study for these two compounds seem to be lower than those documented by Santillo et al. [80] for dust samples of houses in several European cities, although the ranges reported by these researchers are very wide. On the basis of a risk assessment, in 2013, the European Chemicals Agency (ECHA) concluded that there is no evidence that would justify a re-examination of the existing restriction on DINP and DIDP in toys and childcare articles which can be placed in the mouth by children [81]. About 95% of DINP is used in PVC applications. The other 5% is employed in non-PVC applications such as rubbers, adhesives, sealants, paints and lacquers, and lubricants. For DIDP, non-PVC applications are comparatively small, but comprise use in anticorrosion and antifouling paints, sealing compounds and textile inks ([81], and references therein).
As compared with plasticizer compounds, PAHs accounted for a much smaller mass of household dust (Σ20PAHs 0.98 to 116 µg g−1 PM10). Given the small number of samples and the variability in concentrations, it is difficult to infer patterns between house or room typologies (Figure 4). The values obtained in this study fall into the broad range of values reported in the literature. Wang et al. [22] analyzed 15 PAHs in settled house dust of urban dwellings with preschool-aged children in Nanjing, China. ∑15PAHs ranged from 1.2 to 280.4 µg g−1, averaging 11.1 µg g−1. Yadav et al. [85] investigated the contamination level of EPA′s priority PAHs in indoor dust from residential, educational, commercial, public places, and office premises, in four major cities of Nepal. Concentrations of ∑16PAHs ranged from 747 to 4910 ng g−1 (median 1320 ng g−1). In Palermo, Italy, Mannino and Orecchio [86] collected indoor dust samples by brushing from surfaces at a height of 1.5 to 2.0 m above ground level in bedrooms, living rooms, kitchens, laboratories, offices, in a market, and in a car. The ∑16PAH concentrations were within a broad interval (36 to 34,453 µg g−1), with an average of 5111 µg g−1, indicating heterogeneous levels of contamination in the investigated microenvironments. Organic extracts of sieved vacuum cleaner dust from 51 homes in Canada were examined for the presence of 13 PAHs [87]. Total concentrations varied between 1.5 and 325 µg g−1 with a geometric mean of 12.9 µg g−1. These values were found to be comparable to those documented in a previous review in which the total PAHs in samples collected from urban, rural, and suburban homes ranged between 0.4–544 µg g−1 with a geometric mean of 4.5 µg g−1 [7]. High concentrations of ∑16PAHs were also observed in indoor dust samples collected across China from 45 private domiciles and 36 public buildings (1.00 to 470 µg g−1, mean value of 30.9 µg g−1) [88]. It must be noted, once again, that comparability between results of various works should be made with caution, as they concern different surfaces, particle sizes, sampling and analytical methodologies, and list of compounds.
Low molecular weight-PAHs (LMW, two and three rings) were less abundant than high molecular weight-PAHs (HMW, four and six rings), suggesting the dominance of pyrogenic sources. Regardless of the microenvironment, the median LMW/HMW ratios were always in the range from 0.3 to 0.5. The main PAHs in the thoracic fraction of resuspended dust were pyrene, retene, and indeno[1,2,3-cd]pyrene, reaching concentrations up to 27, 21, and 13 µg g−1, respectively (Table 3). While the medians of the latter two compounds were higher in the PM10 sampled in the living rooms, pyrene showed higher levels in the kitchens. PAH levels and speciation are highly dependent on the cooking methods [58], biomass burning appliances and operating conditions [89], traffic fleet and meteorology in the outdoor surrounding environment [90], among other factors.

3.3. Human Exposure to Plasticizers and PAHs in Resuspended Dust

Ingestion is the main pathway for intake of plasticizers from dust (Table 4). Regardless of the route, household residents are exposed to higher intakes in the bedrooms, whereas the lowest doses are experienced in the kitchens. The daily intakes for children and adults (13 to 29 and 1.4 to 3.1 µg kg−1 day−1, respectively) via dust ingestion were three to four orders of magnitude higher than those via inhalation (0.357 to 0.802 and 0.129 to 0.289 ng kg−1 day−1, respectively) and dermal contact (14.9 to 33.0 and 2.05 to 4.62 ng kg−1 day−1, respectively). Children are at higher risk of exposure to plasticizers than adults. The total DIing, DIinh, and DIder for children were about 9.3, 2.8, and 7.1 times higher than those estimated for adults, respectively. DBP, DNOP, and DEHP contributed the most to daily intakes. The dust ingestion intakes for these compounds were lower than the U.S. EPA maximum acceptable oral doses of 0.1, 0.01, and 0.02 mg kg−1 day−1, respectively. However, the daily intakes of plasticizers through dust ingestion, inhalation, and dermal contact in this study are higher than those estimated for indoor dust from houses of several Chinese regions [39], and childcare facilities, salons, and homes across the USA [29]. Albar et al. [37] assessed human exposure to phthalates via dust ingestion for the worst-case scenario (with 95th percentile levels) for Saudi and Kuwaiti toddlers and adults. The exposure to DEHP, which is cardiotoxic and endocrine disruptor, for Saudi toddlers was estimated to be 37,630 ng kg−1 day−1, while for Kuwaiti toddlers it was 6722 ng kg−1 day−1. Similarly, exposure estimates to other PAEs, such as DBP, DIBP, and DNOP, was also higher for Saudi toddlers. In the case of Saudi and Kuwaiti adults, dust ingestion intakes for DEHP were estimated at 1613 and 288 ng kg−1 day−1, respectively.
BaP, DahA, and IcdP were the compounds that most contributed to the carcinogenic potency, accounting for 45% to 57%, 11% to 29% and 11% to 13% of the total BaPTEQ, respectively. As shown in Table 5, the total cancer risk could be attributed almost entirely to ingestion and did not vary much with the microenvironment. Therefore, inhalation of resuspended particles through the mouth and nose or via dermal contact was almost negligible as compared with the ingestion route. Under most regulatory programs, an ILCR between 10−6 and 10−4 indicates potential risk, an ILCR of 10−6 or less is considered insignificant, and an ILCR ≥ 10−4 is taken as high risk [93,94]. In the present study, the total risk of adult and children exposure to PAHs in dust via the three pathways ranged from 7.2 × 10−6 to 1.4 × 10−5. This means that the resuspendable thoracic fraction of household dust can contribute to an estimated excess of 7.2 to 14 per million people new cancer cases. One cancer case per million people is usually used as a baseline level of acceptable risk.
To assess the potential health risk of foliar dust in Nanjing, China, Zha et al. [56] analyzed the contents of 16 priority PAHs. Total concentrations in dust ranged from 1.77 to 19.02 µg g−1, with an average value of 6.98 µg g−1. The cancer risk levels via dermal contact and ingestion varied from 10−8 to 10−6 in all the dust samples, while the mean cancer risk via inhalation was 10−10 to 10−12, about 104 to 107 times lower than through ingestion and dermal contact.
ILCR values due to human exposure to PAHs in indoor dust in city, town, village, and orefield of Guizhou province, China, were 6.14 × 10−6, 5.00 × 10−6, 3.08 × 10−6, and 6.02 × 10−6 for children and 5.92 × 10−6, 4.83 × 10−6, 2.97 × 10−6, and 5.81 × 10−6 for adults, respectively [16]. As noted in our study, inhalation of resuspended particles through mouth and nose was very small as compared with the ingestion pathway. Maertens et al. [87] estimated higher excess lifetime cancer risks, ranging from 1 × 10−6 and 1 × 10−4, associated with nondietary ingestion of PAHs in settled dust from homes in Ottawa, Canada, during preschool years. However, the researchers observed that the level of risk varies substantially according to the ingestion rates adopted to perform the estimates.
Since ingestion was found to be the dominant exposure pathway, the risk associated with seven non-carcinogenic PAHs (ACY, ACE, FLU, PHE, ANT, FLUA, and PYR) in household PM10 dust was estimated through the hazard quotient (HQ). Under most programs, if the HQ value is greater than one, the exposed population is likely to experience considerable non-carcinogenic effects. The highest daily intakes were obtained for PYR (24 to 31 ng kg−1 day−1 for children and 2.6 to 3.3 ng kg−1 day−1 for adults) and ACY (12 to 31 ng kg−1 day−1 for children and 1.2 to 3.3 ng kg−1 day−1 for adults). These values are higher than those documented for dust samples collected from three different microenvironments (cars, air conditioner filters, and household floor dust) of Jeddah, Saudi Arabia, and Kuwait [92]. Nevertheless, the HQ values of our study were less than one, indicating that there was no considerable non-carcinogenic risk arising from ingestion of PAHs in resuspended PM10. Although the risk remains very low for both age groups, higher HQ values for children (1.51 × 10−3 to 1.89 × 10−3) than those obtained for adults (1.32 × 10−4 to 2.03 × 10−4), demonstrate greater susceptibility of the younger ones. It should be noted that HQ values are underestimated as naphthalene was not included in the quantifications. Despite its volatility, naphthalene is normally quite abundant in the particulate phase.

4. Conclusions

This preliminary study provides a first insight on the occurrence of plasticizers and PAHs in PM10 from resuspended dust samples in Spanish households and adds to the growing evidence that non-dietary exposure contributes to the total body burden. Considering that people spend most of their time indoors, exposure to these pollutants could lead to an increased human health risk. Although no appreciable differences between plasticizer and PAH levels in resuspended dust from the various residential microenvironments were observed, it was concluded that exposure through the ingestion route poses much higher risks as compared to inhalation and dermal contact. This is of particular concern for infants due to their higher dust intake via frequent hand-to-mouth activities. Because of the small number of samples analyzed in this study, it should be noted that exposure estimates are only an indication of the likely range for children and adults within the studied population. More assessments with wider spatial and temporal coverage are needed to better understand the dynamics and possible effects of these pollutants in different indoor microenvironments. As observed in other works, this study on exposure to organic pollutants suggests that measurements only in food and outdoor environments can substantially underestimate exposures to chemicals. Comparisons of the results of this study with those reported in the literature revealed huge amplitudes in the numbers and great difficulty in generalizing conclusions, mainly due to the heterogeneity of applied methodologies. Traditional methodologies have been compartmentalized (settled versus suspended dust) and as a result may have described the environment incompletely. Thus, the scientific community and international agencies should discuss and establish standardized protocols.

Supplementary Materials

The following are available online at, Analytical technique for quantification of the carbonaceous content of PM10, Table S1: Standards from Sigma-Aldrich (and product references) used in chromatographic analysis, Table S2: Dust loading obtained in the various houses, in both sampling periods, mass percentage of carbonaceous material (TC = OC + EC) and mass fractions of plasticizers and polycyclic aromatic hydrocarbons in PM10.

Author Contributions

Conceptualization, C.A., E.D.V., and F.A.; sampling, E.D.V., A.C. (A. Calvo), C.d.B.-A., F.O., A.C. (A. Castro), and R.F.; carbon analyzes, T.N.; filter solvent extractions, E.D.V.; GC-MS analyzes, A.V.; writing of original draft, C.A.; review of original draft, C.A., in collaboration with all co-authors.


The sampling and analytical work was mainly supported by the project POCI-01-0145-FEDER-029574 (SOPRO- Chemical and toxicological SOurce PROfiling of particulate matter in urban air), funded by FEDER, through COMPETE2020 - Programa Operacional Competitividade e Internacionalização (POCI), and by national funds (OE), through FCT/MCTES. The authors would like to express their gratitude to the Portuguese Foundation of Science and Technology (FCT) and to the POHP/FSE funding program for the fellowship SFRH/BD/117993/2016. Ana Vicente was supported by national funds (OE), through FCT, I.P., in the scope of the framework contract foreseen in the numbers 4, 5, and 6 of article 23, of the Decree-Law 57/2016, of August 29, changed by Law 57/2017, of July 19. Some support was received from CESAM (UID/AMB/50017), which was funded by FCT/MEC through national funds, and co-funded by FEDER, within the PT2020 Partnership Agreement and Compete 2020.


We thank the homeowners who allowed sampling in various rooms.

Conflicts of Interest

The authors declare no conflict of interest.


  1. Schweizer, C.; Edwards, R.D.; Bayer-Oglesby, L.; Gauderman, W.J.; Ilacqua, V.; Juhani Jantunen, M.; Lai, H.K.; Nieuwenhuijsen, M.; Künzli, N. Indoor time–microenvironment–activity patterns in seven regions of Europe. J. Expo. Sci. Environ. Epidemiol. 2007, 17, 170–181. [Google Scholar] [CrossRef] [PubMed][Green Version]
  2. Qian, J.; Peccia, J.; Ferro, A.R. Walking-induced particle resuspension in indoor environments. Atmos. Environ. 2014, 89, 464–481. [Google Scholar] [CrossRef]
  3. Tian, Y.; Sul, K.; Qian, J.; Mondal, S.; Ferro, A.R. A comparative study of walking-induced dust resuspension using a consistent test mechanism. Indoor Air 2014, 24, 592–603. [Google Scholar] [CrossRef] [PubMed]
  4. Bo, M.; Salizzoni, P.; Clerico, M.; Buccolieri, R. Assessment of indoor-outdoor particulate matter air pollution: A review. Atmosphere 2017, 8, 136. [Google Scholar] [CrossRef][Green Version]
  5. Naspinski, C.; Lingenfelter, R.; He, L.Y.; Cizmas, L.; Naufal, Z.; Islamzadeh, A.; Li, Z.; Li, Z.; Donnelly, K.C.; McDonald, T. A comparison of concentrations of polycyclic aromatic compounds detected in dust samples from various regions of the world. Environ. Int. 2008, 34, 988–993. [Google Scholar] [CrossRef]
  6. Lioy, P.J.; Freeman, N.C.G.; Millette, J.R. Review Dust: A Metric for Use in Residential and Building Exposure Assessment and. Environ. Health Perspect. 2002, 110, 969–983. [Google Scholar] [CrossRef][Green Version]
  7. Maertens, R.M.; Bailey, J.; White, P.A. The mutagenic hazards of settled house dust: A review. Mutat. Res. 2004, 567, 401–425. [Google Scholar] [CrossRef]
  8. Liu, H.Y.; Dunea, D.; Iordache, S.; Pohoata, A. A review of airborne particulate matter effects on young children’s respiratory symptoms and diseases. Atmosphere 2018, 9, 150. [Google Scholar] [CrossRef][Green Version]
  9. Cipriani, F.; Calamelli, E.; Ricci, G. Allergen avoidance in allergic asthma. Front. Pediatr. 2017, 5, 103. [Google Scholar] [CrossRef][Green Version]
  10. Ahluwalia, S.K.; Matsui, E.C. The indoor environment and its effects on childhood asthma. Curr. Opin. Allergy Clin. Immunol. 2011, 11, 137–143. [Google Scholar] [CrossRef]
  11. McCormack, M.C.; Breysse, P.N.; Matsui, E.C.; Hansel, N.N.; Peng, R.D.; Curtin-Brosnan, J.; Williams, D.L.; Wills-Karp, M.; Diette, G.B.; Center for Childhood Asthma in the Urban Environment. Indoor particulate matter increases asthma morbidity in children with non-atopic and atopic asthma. Ann. Allergy. Asthma Immunol. 2011, 106, 308–315. [Google Scholar] [CrossRef] [PubMed][Green Version]
  12. Mitro, S.D.; Dodson, R.E.; Singla, V.; Adamkiewicz, G.; Elmi, A.F.; Tilly, M.K.; Zota, A.R. Consumer Product Chemicals in Indoor Dust: A Quantitative Meta-analysis of U.S. Studies. Environ. Sci. Technol. 2016, 50, 10661–10672. [Google Scholar] [CrossRef] [PubMed]
  13. Roberts, J.W.; Wallace, L.A.; Camann, D.E.; Dickey, P.; Gilbert, S.G.; Lewis, R.G.; Takaro, T.K. Monitoring and Reducing Exposure of Infants to Pollutants in House Dust. In Reviews of Environmental Contamination and Toxicology; Springer: Berlin, Germany, 2009; Volume 201, pp. 1–39. [Google Scholar]
  14. Langer, S.; Weschler, C.J.; Fischer, A.; Bekö, G.; Toftum, J.; Clausen, G. Phthalate and PAH concentrations in dust collected from Danish homes and daycare centers. Atmos. Environ. 2010, 44, 2294–2301. [Google Scholar] [CrossRef]
  15. Dodson, R.E.; Camann, D.E.; Morello-Frosch, R.; Brody, J.G.; Rudel, R.A. Semivolatile organic compounds in homes: strategies for efficient and systematic exposure measurement based on empirical and theoretical factors. Environ. Sci. Technol. 2015, 49, 113–122. [Google Scholar] [CrossRef] [PubMed][Green Version]
  16. Yang, Q.; Chen, H.; Li, B. Polycyclic aromatic hydrocarbons (PAHs) in indoor dusts of Guizhou, southwest of China: Status, sources and potential human health risk. PLoS ONE 2015, 10, e0118141. [Google Scholar] [CrossRef]
  17. Mercier, F.; Glorennec, P.; Thomas, O.; Bot, B. Le Organic contamination of settled house dust, a review for exposure assessment purposes. Environ. Sci. Technol. 2011, 45, 6716–6727. [Google Scholar] [CrossRef]
  18. DellaValle, C.T.; Deziel, N.C.; Jones, R.R.; Colt, J.S.; De Roos, A.J.; Cerhan, J.R.; Cozen, W.; Severson, R.K.; Flory, A.R.; Morton, L.M.; et al. Polycyclic aromatic hydrocarbons: Determinants of residential carpet dust levels and risk of non-Hodgkin lymphoma. Cancer Causes Control 2016, 27, 1–13. [Google Scholar] [CrossRef][Green Version]
  19. Boström, C.E.; Gerde, P.; Hanberg, A.; Jernström, B.; Johansson, C.; Kyrklund, T.; Rannug, A.; Törnqvist, M.; Victorin, K.; Westerholm, R. Cancer risk assessment, indicators, and guidelines for polycyclic aromatic hydrocarbons in the ambient air. Environ. Health Perspect. 2002, 110, 451–488. [Google Scholar]
  20. Kim, K.H.; Jahan, S.A.; Kabir, E.; Brown, R.J.C. A review of airborne polycyclic aromatic hydrocarbons (PAHs) and their human health effects. Environ. Int. 2013, 60, 71–80. [Google Scholar] [CrossRef]
  21. Deziel, N.C.; Rull, R.P.; Colt, J.S.; Reynolds, P.; Whitehead, T.P.; Gunier, R.B.; Month, S.R.; Taggart, D.R.; Buffler, P.; Ward, M.H.; et al. Polycyclic aromatic hydrocarbons in residential dust and risk of childhood acute lymphoblastic leukemia. Environ. Res. 2014, 133, 388–395. [Google Scholar] [CrossRef][Green Version]
  22. Wang, B.L.; Pang, S.T.; Zhang, X.L.; Li, X.L.; Sun, Y.G.; Lu, X.M.; Zhang, Q.; Zhang, Z.D. Levels and neurodevelopmental effects of polycyclic aromatic hydrocarbons in settled house dust of urban dwellings on preschool-aged children in Nanjing, China. Atmos. Pollut. Res. 2014, 5, 292–302. [Google Scholar] [CrossRef][Green Version]
  23. Kang, Y.; Cheung, K.C.; Wong, M.H. Polycyclic aromatic hydrocarbons (PAHs) in different indoor dusts and their potential cytotoxicity based on two human cell lines. Environ. Int. 2010, 36, 542–547. [Google Scholar] [CrossRef] [PubMed]
  24. Liu, R.; He, R.; Cui, X.; Ma, L.Q. Impact of particle size on distribution, bioaccessibility, and cytotoxicity of polycyclic aromatic hydrocarbons in indoor dust. J. Hazard. Mater. 2018, 357, 341–347. [Google Scholar] [CrossRef] [PubMed]
  25. World Health Organization International Agency for Research on Cancer. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans. Some Non-Heterocyclic Polycyclic Aromatic Hydrocarbons and Some Related Exposures; WHO: Lyon, France, 2010. [Google Scholar]
  26. ATSDR. Priority List of Hazardous Substances. Available online: (accessed on 5 December 2019).
  27. Ma, W.-L.; Subedi, B.; Kannan, K. The Occurrence of Bisphenol A, Phthalates, Parabens and Other Environmental Phenolic Compounds in House Dust: A Review. Curr. Org. Chem. 2014, 18, 2182–2199. [Google Scholar] [CrossRef]
  28. Larsson, K.; Lindh, C.H.; Jönsson, B.A.; Giovanoulis, G.; Bibi, M.; Bottai, M.; Bergström, A.; Berglund, M. Phthalates, non-phthalate plasticizers and bisphenols in Swedish preschool dust in relation to children’s exposure. Environ. Int. 2017, 102, 114–124. [Google Scholar] [CrossRef]
  29. Subedi, B.; Sullivan, K.D.; Dhungana, B. Phthalate and non-phthalate plasticizers in indoor dust from childcare facilities, salons, and homes across the USA. Environ. Pollut. 2017, 230, 701–708. [Google Scholar] [CrossRef][Green Version]
  30. Kang, Y.; Man, Y.B.; Cheung, K.C.; Wong, M.H. Risk assessment of human exposure to bioaccessible phthalate esters via indoor dust around the Pearl River Delta. Environ. Sci. Technol. 2012, 46, 8422–8430. [Google Scholar] [CrossRef]
  31. Kweon, D.J.; Kim, M.K.; Zoh, K.D. Distribution of brominated flame retardants and phthalate esters in house dust in Korea. Environ. Eng. Res. 2018, 23, 354–363. [Google Scholar] [CrossRef]
  32. Abb, M.; Heinrich, T.; Sorkau, E.; Lorenz, W. Phthalates in house dust. Environ. Int. 2009, 35, 965–970. [Google Scholar] [CrossRef]
  33. Kubwabo, C.; Rasmussen, P.E.; Fan, X.; Kosarac, I.; Wu, F.; Zidek, A.; Kuchta, S.L. Analysis of selected phthalates in Canadian indoor dust collected using household vacuum and standardized sampling techniques. Indoor Air 2013, 23, 506–514. [Google Scholar] [CrossRef]
  34. Bornehag, C.G.; Lundgren, B.; Weschler, C.J.; Sigsgaard, T.; Hagerhed-Engman, L.; Sundell, J. Phthalates in indoor dust and their association with building characteristics. Environ. Health Perspect. 2005, 113, 1399–1404. [Google Scholar] [CrossRef] [PubMed][Green Version]
  35. Al Qasmi, N.N.; Al-Thaiban, H.; Helaleh, M.I.H. Indoor phthalates from household dust in Qatar: Implications for non-dietary human exposure. Environ. Sci. Pollut. Res. Int. 2019, 26, 421–430. [Google Scholar] [CrossRef] [PubMed]
  36. Zhang, Q.; Lu, X.M.; Zhang, X.L.; Sun, Y.G.; Zhu, D.M.; Wang, B.L.; Zhao, R.Z.; Zhang, Z.D. Levels of phthalate esters in settled house dust from urban dwellings with young children in Nanjing, China. Atmos. Environ. 2013, 69, 258–264. [Google Scholar] [CrossRef]
  37. Albar, H.M.S.A.; Ali, N.; Shahzad, K.; Ismail, I.M.I.; Rashid, M.I.; Wang, W.; Ali, L.N.; Eqani, S.A.M.A.S. Phthalate esters in settled dust of different indoor microenvironments; source of non-dietary human exposure. Microchem. J. 2017, 132, 227–232. [Google Scholar]
  38. Kashyap, D.; Agarwal, T. Concentration and factors affecting the distribution of phthalates in the air and dust: A global scenario. Sci. Total Environ. 2018, 635, 817–827. [Google Scholar] [CrossRef]
  39. Zhu, Q.; Jia, J.; Zhang, K.; Zhang, H.; Liao, C.; Jiang, G. Phthalate esters in indoor dust from several regions, China and their implications for human exposure. Sci. Total Environ. 2019, 652, 1187–1194. [Google Scholar] [CrossRef]
  40. Guo, Y.; Kannan, K. Comparative assessment of human exposure to phthalate esters from house dust in China and the United States. Environ. Sci. Technol. 2011, 45, 3788–3794. [Google Scholar] [CrossRef]
  41. Orecchio, S.; Indelicato, R.; Barreca, S. The distribution of phthalate esters in indoor dust of Palermo (Italy). Environ. Geochem. Health 2013, 35, 613–624. [Google Scholar] [CrossRef][Green Version]
  42. Kay, V.R.; Bloom, M.S.; Foster, W.G. Reproductive and developmental effects of phthalate diesters in males. Crit. Rev. Toxicol. 2014, 44, 467–498. [Google Scholar] [CrossRef]
  43. Kay, V.R.; Chambers, C.; Foster, W.G. Reproductive and developmental effects of phthalate diesters in females. Crit. Rev. Toxicol. 2013, 43, 200–219. [Google Scholar] [CrossRef][Green Version]
  44. Wormuth, M.; Scheringer, M.; Vollenweider, M.; Hungerbühler, K. What are the sources of exposure to eight frequently used phthalic acid esters in Europeans? Risk Anal. 2006, 26, 803–824. [Google Scholar] [CrossRef] [PubMed]
  45. Directive 2005/84/EC of the European Parliament and of the Council of 14 December 2005 amending for the 22nd time Council Directive 76/769/EEC on the approximation of the laws, regulations and administrative provisions of the Member States relating to restrictions on the marketing and use of certain dangerous substances and preparations (phthalates in toys and childcare articles). Available online: (accessed on 5 December 2019).
  46. Manigrasso, M.; Guerriero, E.; Avino, P. Ultrafine particles in residential indoors and doses deposited in the human respiratory system. Atmosphere 2015, 6, 1444–1461. [Google Scholar] [CrossRef][Green Version]
  47. Morawska, L.; Salthammer, T. (Eds.) Indoor Environment, Airborne Particles and Settled Dust; Wiley-VCH GmbH & Co. KGaA: Weinheim, Germany, 2003; ISBN 978-352-7305254. [Google Scholar]
  48. Ma, Y.; Harrad, S. Spatiotemporal analysis and human exposure assessment on polycyclic aromatic hydrocarbons in indoor air, settled house dust, and diet: A review. Environ. Int. 2015, 84, 7–16. [Google Scholar] [CrossRef] [PubMed]
  49. Driver, J.H.; Konz, J.J.; Whitmyre, G.K. Soil adherence to human skin. Bull. Environ. Contam. Toxicol. 1989, 43, 814–820. [Google Scholar] [CrossRef] [PubMed]
  50. Amato, F.; Pandolfi, M.; Moreno, T.; Furger, M.; Pey, J.; Alastuey, A.; Bukowiecki, N.; Prevot, A.S.H.; Baltensperger, U.; Querol, X. Sources and variability of inhalable road dust particles in three European cities. Atmos. Environ. 2011, 45, 6777–6787. [Google Scholar] [CrossRef]
  51. Amato, F.; Pandolfi, M.; Viana, M.; Querol, X.; Alastuey, A.; Moreno, T. Spatial and chemical patterns of PM10 in road dust deposited in urban environment. Atmos. Environ. 2009, 43, 1650–1659. [Google Scholar] [CrossRef]
  52. Alves, C.A.; Evtyugina, M.; Vicente, A.M.P.; Vicente, E.D.; Nunes, T.V.; Silva, P.M.A.; Duarte, M.A.C.; Pio, C.A.; Amato, F.; Querol, X. Chemical profiling of PM10 from urban road dust. Sci. Total Environ. 2018, 634, 41–51. [Google Scholar] [CrossRef]
  53. Anderson, J.O.; Thundiyil, J.G.; Stolbach, A. Clearing the Air: A Review of the Effects of Particulate Matter Air Pollution on Human Health. J. Med. Toxicol. 2012, 8, 166–175. [Google Scholar] [CrossRef][Green Version]
  54. Alves, C.A.; Vicente, A.; Monteiro, C.; Gonçalves, C.; Evtyugina, M.; Pio, C. Emission of trace gases and organic components in smoke particles from a wildfire in a mixed-evergreen forest in Portugal. Sci. Total Environ. 2011, 409, 1466–1475. [Google Scholar] [CrossRef]
  55. Fortune, A.; Gendron, L.; Tuday, M. Comparison of naphthalene ambient air sampling & analysis methods at former manufactured gas plant (MGP) remediation sites. Int. J. Soil Sediment Water 2010, 3, 1–15. [Google Scholar]
  56. Zha, Y.; Zhang, Y.L.; Tang, J.; Sun, K. Status, sources, and human health risk assessment of PAHs via foliar dust from different functional areas in Nanjing, China. J. Environ. Sci. Health Part A 2018, 53, 571–582. [Google Scholar] [CrossRef] [PubMed]
  57. Iwegbue, C.M.A.; Obi, G.; Uzoekwe, S.A.; Egobueze, F.E.; Odali, E.W.; Tesi, G.O.; Nwajei, G.E.; Martincigh, B.S. Distribution, sources and risk of exposure to polycyclic aromatic hydrocarbons in indoor dusts from electronic repair workshops in southern Nigeria. Emerg. Contam. 2019, 5, 23–30. [Google Scholar] [CrossRef]
  58. Abdullahi, K.L.; Delgado-Saborit, J.M.; Harrison, R.M. Emissions and indoor concentrations of particulate matter and its specific chemical components from cooking: A review. Atmos. Environ. 2013, 71, 260–294. [Google Scholar] [CrossRef]
  59. Licina, D.; Tian, Y.; Nazaroff, W.W. Emission rates and the personal cloud effect associated with particle release from the perihuman environment. Indoor Air 2017, 27, 791–802. [Google Scholar] [CrossRef] [PubMed][Green Version]
  60. Lewis, R.D.; Ong, K.H.; Emo, B.; Kennedy, J.; Kesavan, J.; Elliot, M. Resuspension of house dust and allergens during walking and vacuum cleaning. J. Occup. Environ. Hyg. 2018, 15, 235–245. [Google Scholar] [CrossRef]
  61. Becher, R.; Øvrevik, J.; Schwarze, P.E.; Nilsen, S.; Hongslo, J.K.; Bakke, J.V. Do carpets impair indoor air quality and cause adverse health outcomes: A review. Int. J. Environ. Res. Public Health 2018, 15, 184. [Google Scholar] [CrossRef][Green Version]
  62. Foarde, K.; Berry, M. Comparison of biocontaminant levels associated with hard vs. carpet floors in nonproblem schools: Results of a year long study. J. Expo. Anal. Environ. Epidemiol. 2004, 14, S41–S48. [Google Scholar] [CrossRef][Green Version]
  63. Tranter, D.C. Indoor allergens in settled school dust: A review of findings and significant factors. Clin. Exp. Allergy 2005, 35, 126–136. [Google Scholar] [CrossRef]
  64. Causer, S.; Shorter, C.; Sercombe, J. Effect of floorcovering construction on content and vertical distribution of house dust mite allergen, Der p I. J. Occup. Environ. Hyg. 2006, 3, 161–168. [Google Scholar] [CrossRef]
  65. Roberts, J.W.; Clifford, W.S.; Glass, G.; Hummer, P.G. Reducing dust, lead, dust mites, bacteria, and fungi in carpets by vacuuming. Arch. Environ. Contam. Toxicol. 1999, 36, 477–484. [Google Scholar]
  66. Adgate, J.L.; Weisel, C.; Wang, Y.; Rhoads, G.G.; Lioy, P.J. Lead in house dust: Relationships between exposure metrics. Environ. Res. 1995, 70, 134–147. [Google Scholar] [CrossRef] [PubMed]
  67. Lioy, P.J.; Freeman, N.C.; Wainman, T.; Stern, A.H.; Boesch, R.; Howell, T.; Shupack, S.I. Microenvironmental analysis of residential exposure to chromium-laden wastes in and around New Jersey homes. Risk Anal. 1992, 12, 287–299. [Google Scholar] [CrossRef] [PubMed]
  68. Bai, Z.; Yiin, L.M.; Rich, D.Q.; Adgate, J.L.; Ashley, P.J.; Lioy, P.J.; Rhoads, G.G.; Zhang, J. Field evaluation and comparison of five methods of sampling lead dust on carpets. Am. Ind. Hyg. Assoc. J. 2003, 64, 528–532. [Google Scholar] [CrossRef]
  69. Polidori, A.; Turpin, B.; Meng, Q.Y.; Lee, J.H.; Weisel, C.; Morandi, M.; Colome, S.; Stock, T.; Winer, A.; Zhang, J.; et al. Fine organic particulate matter dominates indoor-generated PM2.5 in RIOPA homes. J. Expo. Sci. Environ. Epidemiol. 2006, 16, 321–331. [Google Scholar] [CrossRef] [PubMed][Green Version]
  70. Bamai, Y.A.; Araki, A.; Kawai, T.; Tsuboi, T.; Saito, I.; Yoshioka, E.; Kanazawa, A.; Tajima, S.; Shi, C.; Tamakoshi, A.; et al. Associations of phthalate concentrations in floor dust and multi-surface dust with the interior materials in Japanese dwellings. Sci. Total Environ. 2014, 468, 147–157. [Google Scholar] [CrossRef]
  71. Afshari, A.; Gunnarsen, L.; Clausen, P.A.; Hansen, V. Emission of phthalates from PVC and other materials. Indoor Air 2004, 14, 120–128. [Google Scholar] [CrossRef]
  72. Koniecki, D.; Wang, R.; Moody, R.P.; Zhu, J. Phthalates in cosmetic and personal care products: Concentrations and possible dermal exposure. Environ. Res. 2011, 111, 329–336. [Google Scholar] [CrossRef]
  73. U.S. National Library of Medicine. National Center for Biotechnology Information. PubChem, 2019. Available online: (accessed on 5 December 2019).
  74. Luongo, G.; Östman, C. Organophosphate and phthalate esters in settled dust from apartment buildings in Stockholm. Indoor Air 2016, 26, 414–425. [Google Scholar] [CrossRef]
  75. Brouwere, K.; Standaert, A.; Torfs, R. Integrated Exposure for Risk Assessment in Indoor Environment (INTERA); Final Report; INTERA: Boeretang, Belgium, 2012. [Google Scholar]
  76. Kolarik, B.; Bornehag, C.G.; Naydenov, K.; Sundell, J.; Stavova, P.; Nielsen, O.F. The concentrations of phthalates in settled dust in Bulgarian homes in relation to building characteristic and cleaning habits in the family. Atmos. Environ. 2008, 42, 8553–8559. [Google Scholar] [CrossRef]
  77. Xu, Y.; Cohen Hubal, E.A.; Little, J.C. Predicting residential exposure to phthalate plasticizer emitted from vinyl flooring: Sensitivity, uncertainty, and implications for biomonitoring. Environ. Health Perspect. 2010, 118, 253–258. [Google Scholar] [CrossRef]
  78. Weschler, C.J. Indoor/outdoor connections exemplified by processes that depend on an organic compound’s saturation vapor pressure. In Proceedings of the Atmospheric Environment; Elsevier: Amsterdam, The Netherlands, 2003; Volume 37, pp. 5455–5465. [Google Scholar]
  79. Institute for Health and Consumer Protection, Toxicology and Chemical Substances. Benzyl Butyl Phthalate. In Summary Risk Assessment Report; EUR 22773 EN/2; Institute for Health and Consumer Protection, Toxicology and Chemical Substances, European Chemicals Bureau: Ispra, Italy, 2008. [Google Scholar]
  80. Santillo, D.; Labunska, I.; Davidson, H.; Johnston, P.; Strutt, M.; Knowles, O. Consuming Chemicals–Hazardous Chemicals in House Dust as an Indicator of Chemical Exposure in the Home; Greenpeace Research Laboratories: Exeter, UK, 2003. [Google Scholar]
  81. European Chemicals Agency. Evaluation of New Scientific Evidence Concerning DINP and DIDP in Relation to Entry 52 of Annex XVII to Reach Regulation (EC) No 1907/2006; ECHA: Helsinki, Finland, 2013. [Google Scholar]
  82. Fromme, H.; Lahrz, T.; Piloty, M.; Gebhart, H.; Oddoy, A.; Rüden, H. Occurrence of phthalates and musk fragrances in indoor air and dust from apartments and kindergartens in Berlin (Germany). Indoor Air 2004, 14, 188–195. [Google Scholar] [CrossRef] [PubMed]
  83. Gevao, B.; Al-Ghadban, A.N.; Bahloul, M.; Uddin, S.; Zafar, J. Phthalates in indoor dust in Kuwait: Implications for non-dietary human exposure. Indoor Air 2013, 23, 126–133. [Google Scholar] [CrossRef] [PubMed]
  84. Kanazawa, A.; Saito, I.; Araki, A.; Takeda, M.; Ma, M.; Saijo, Y.; Kishi, R. Association between indoor exposure to semi-volatile organic compounds and building-related symptoms among the occupants of residential dwellings. Indoor Air 2010, 20, 72–84. [Google Scholar] [CrossRef] [PubMed]
  85. Yadav, I.C.; Devi, N.L.; Li, J.; Zhang, G. Polycyclic aromatic hydrocarbons in house dust and surface soil in major urban regions of Nepal: Implication on source apportionment and toxicological effect. Sci. Total Environ. 2018, 616, 223–235. [Google Scholar] [CrossRef]
  86. Mannino, M.R.; Orecchio, S. Polycyclic aromatic hydrocarbons (PAHs) in indoor dust matter of Palermo (Italy) area: Extraction, GC-MS analysis, distribution and sources. Atmos. Environ. 2008, 42, 1801–1817. [Google Scholar] [CrossRef]
  87. Maertens, R.M.; Yang, X.; Zhu, J.; Gagne, R.W.; Douglas, G.R.; White, P.A. Mutagenic and carcinogenic hazards of settled house dust I: Polycyclic aromatic hydrocarbon content and excess lifetime cancer risk from preschool exposure. Environ. Sci. Technol. 2008, 42, 1747–1753. [Google Scholar] [CrossRef][Green Version]
  88. Qi, H.; Li, W.L.; Zhu, N.Z.; Ma, W.L.; Liu, L.Y.; Zhang, F.; Li, Y.F. Concentrations and sources of polycyclic aromatic hydrocarbons in indoor dust in China. Sci. Total Environ. 2014, 491, 100–107. [Google Scholar] [CrossRef]
  89. Vicente, E.D.; Alves, C.A. An overview of particulate emissions from residential biomass combustion. Atmos. Res. 2018, 199, 159–185. [Google Scholar] [CrossRef]
  90. Cheruyiot, N.K.; Lee, W.J.; Mwangi, J.K.; Wang, L.C.; Lin, N.H.; Lin, Y.C.; Cao, J.; Zhang, R.; Chang-Chien, G.P. An overview: Polycyclic aromatic hydrocarbon emissions from the stationary and mobile sources and in the ambient air. Aerosol Air Qual. Res. 2015, 15, 2730–2762. [Google Scholar] [CrossRef][Green Version]
  91. Ong, S.; Ayoko, G.; Kokot, S.; Morawska, L. Polycyclic aromatic hydrocarbons in house dust samples: Source identification and apportionment. In Proceedings of the 14th International IUAPPA World Congress, Brisbane, QLD, Australia, 9–13 September 2007. [Google Scholar]
  92. Ali, N.; Ismail, I.M.I.; Khoder, M.; Shamy, M.; Alghamdi, M.; Costa, M.; Ali, L.N.; Wang, W.; Eqani, S.A.M.A.S. Polycyclic aromatic hydrocarbons (PAHs) in indoor dust samples from Cities of Jeddah and Kuwait: Levels, sources and non-dietary human exposure. Sci. Total Environ. 2016, 573, 1607–1614. [Google Scholar] [CrossRef]
  93. Qu, C.; Qi, S.; Yang, D.; Huang, H.; Zhang, J.; Chena, W.; Yohannes, H.K.; Sandy, E.H.; Yang, J.; Xing, X. Risk assessment and influence factors of organochlorine pesticides (OCPs) in agricultural soils of the hill region: A case study from Ningde, southeast China. J. Geochem. Explor. 2015, 149, 43–51. [Google Scholar] [CrossRef]
  94. Roy, D.; Seo, Y.C.; Sinha, S.; Bhattacharya, A.; Singh, G.; Biswas, P.K. Human health risk exposure with respect to particulate-bound polycyclic aromatic hydrocarbons at mine fire-affected coal mining complex. Environ. Sci. Pollut. Res. 2019, 26, 19119–19135. [Google Scholar] [CrossRef] [PubMed]
Figure 1. Dust loadings in the different rooms of the various houses for the two sampling campaigns.
Figure 1. Dust loadings in the different rooms of the various houses for the two sampling campaigns.
Atmosphere 10 00785 g001
Figure 2. Total carbon (TC = OC + EC) mass fraction in resuspendable PM10 from household dust.
Figure 2. Total carbon (TC = OC + EC) mass fraction in resuspendable PM10 from household dust.
Atmosphere 10 00785 g002
Figure 3. Plasticizer mass fractions in PM10 from resuspended household dust.
Figure 3. Plasticizer mass fractions in PM10 from resuspended household dust.
Atmosphere 10 00785 g003
Figure 4. Polycyclic aromatic hydrocarbon (PAH) mass fractions in PM10 dust samples collected in two different sampling periods.
Figure 4. Polycyclic aromatic hydrocarbon (PAH) mass fractions in PM10 dust samples collected in two different sampling periods.
Atmosphere 10 00785 g004
Table 1. Characteristics of the houses where floor dust particulate matter with an aerodynamic diameter smaller than 10 µm (PM10) was sampled.
Table 1. Characteristics of the houses where floor dust particulate matter with an aerodynamic diameter smaller than 10 µm (PM10) was sampled.
1Suburban two-story house with well-ventilated kitchen, two occupantsKitchen
Living room
Living room rug
Bedroom rug
Cut pile carpet/rug
Long threads shag rug
2Single story apartment located in the city center, two occupantsKitchen
Living room
3Rural two-story house with open fireplace in the living room, two occupantsKitchen
Living room
4Single story apartment with small kitchen open to the living room, one occupantKitchen
Living room
Table 2. Comparison of mass fractions of plasticizers (µg g−1) in house dust.
Table 2. Comparison of mass fractions of plasticizers (µg g−1) in house dust.
This study
Median<0.1 <0.1 <0.1 6.4<0.1 <0.14079.0<0.1
This study
Median0.19 <0.1 964 7.07.4 4423167.33.9
Living roomsMean
This study
Median<0.1 16 <0.1 9.3<0.1 2184738.216.6
Living room rug *Mean0.60 21 893 8.5423.4 19978079.823.8This study
Bedroom rug *Mean1.31 5.8 139 3.443.7 921195.39.1This study
Dust samples
(<180 µm) from houses of China collected with brushes
Mean0.693 0.18717.126.40.0030.040 0.0151050.342 [39]
Median0.181 0.1169.3312.90.0010.014 0.00725.60.130
30 household dust samples
(<63 µm) from vacuum cleaner
Median 87.4 15.2 604 12933.6[32]
Dust from 30 apartments in bags of vacuum cleanerMean10.854.644.6 55.6 86.1 776 [82]
Median1.537.56.1 47.0 29.7 703
Dust (<250 µm) from 11 houses, 3 labs and 1 hospital in vacuum cleaner bags, QatarMean0.89 7.9 2885.810611.4[35]
Median0.98 7.0 3953.010111.0
Dust samples
(<250 µm) from homes in Kuwait from vacuum cleaner bags
Mean0.01 1.5 510.26.4 0.3170014 [83]
Median0.03 1.8 450.398.6 2.9225614
House dust from China collected by sweeping the floor and wiping the top of furniture
(<2 mm)
Median0.2 0.417.220.1nd0.2 nd2280.2 [40]
Dust from Albany, USA, from vacuum cleaner bags of several homes
(<2 mm)
Median0.08 2.03.813.10.621.1 nd3040.4 [40]
Saudi floor dust from vacuum cleaner bags
(<250 µm)
Mean1.4 4.233.680.2 1.5 1140102.4 [37]
Median0.6 1.422.133.3 0.8 102026.8
Kuwaiti floor dust from vacuum cleaner bags (<250 µm)Mean0.2 1.3 2202.4 [37]
Median0.1 2.717.21.6 0.8 2402.8
Settled dust collected in child′s room, above the floor level, in Bulgarian homes by vacuum cleanerMean260 350 7850 320 960250 [76]
Median280 340 9930 340 1050300
Dust from rooms in Germany collected with vacuum cleaner (<2 mm)Mean1.32 80.7107.1384 85.8 1026 11413.5[80]
Median1.42 12.936.544.1 82.2 996 113< 0.1
Dust from rooms in Spain (Madrid) collected with vacuum cleaner (<2 mm)Mean0.14 14.0265131 6.55 464 69.510.2[80]
Median< 0.1 7.88201145 5.3 370 <0.1<0.1
Dust from rooms in France (Paris) collected with vacuum cleaner (<2 mm)Mean<0.1 9.8891.4174 227 111 12248.2[80]
Median<0.1 8.7286.165.8 200 356 115<0.1
Dust from rooms in Italy (Rome) collected with vacuum cleaner (<2 mm)Mean0.30 9.6022136.6 89.0 503 76.0142[80]
Median<0.1 6.7818042.8 23.6 434 <0.1<0.1
Dust from rooms in the UK collected with vacuum cleaner (<2 mm)Mean0.12 12.25250.2 56.5 192 48.520.8[80]
Median<0.1 3.543.252.8 24.5 195 <0.1<0.1
Settled dust from apartments in Stockholm from vacuum cleanerMedian0.47 14104103 16 4490.0010656[74]
Dust from the floor surface and from objects within 35 cm above the floor with vacuum cleaner in Japanese dwellingsMedian<0.5 < 2.08.0 1110 139 [70]
Floor dust samples collected with vacuum cleaner from living rooms in 41 dwellings, Sapporo (Japan)Median<0.2 0.332.919.8 4.26.5 880 126 [84]
Dust samples from household vacuum cleaner bags in French dwellings
(<100 µm)
Mean0.26 10.611110.5 25.8 441 158 [17]
Median0.25 4.9209.1 6.1 462 139
House dust from urban dwellings in Nanjing, China, collected with vacuum cleaner (150 µm)Mean0.4 0.9 52.3 2.9 4621.6 [36]
Median0.1 0.2 23.7 1.6 1830.1
Dust from homes across the USA with vacuum cleaner (<1.4 mm)Mean0.05 2.346.244.17 21477.1 97.2144 [29]
Median0.04 0.494.083.54 51.221.8 73.174.9
* only two samples, dimethyl phthalate (DMP), dimethylpropyl phthalate (DMPP), diethyl phthalate (DEP), diisobutyl phthalate (DIBP), di-n-butyl phthalate (DBP), di-n-hexylphthalate (DNHP), benzyl butyl phthalate (BBP), bis(2-ethylhexyl) adipate (DEHA), dicyclohexyl phthalate (DCHP), bis(2-ethylhexyl) phthalate (DEHP), di-n-octyl phthalate (DNOP), di-isononylphthalate (DINP), and di-isodecylphthalate (DIDP); nd, not detected.
Table 3. Comparison of PAH levels (µg g−1) in household dust from different countries.
Table 3. Comparison of PAH levels (µg g−1) in household dust from different countries.
PAHsThis Study[91][85][87][88][92]
KitchensLiving RoomsBedroomsRug BedroomRug Living RoomHouse Vacuum Samples BrisbaneIndoor Dust from Distinct Buildings NepalSettled Dust from Homes in OttawaDust Samples from Private Domiciles and Public Buildings, ChinaSaudi Household Floor DustKuwaiti Household Floor Dust
0.0630.0670.1362020.018 0.0930.0520.1100.1050.0650.060
<0.001< 0.001
0.045<0.0010.19753.60.128 4.261.770.4250.1750.2200.140
0.6040.0031.0677.30.107 2.891.180.3850.1600.0350.017
< 0.0010.960
ACY, acenaphthylene; ACE, acenaphthene; FLU, fluorene; PHE, phenanthrene; ANT, anthracene; FLUA, fluoranthene; PYR, pyrene; CHR, chrysene; PER, perylene; CAR, carbazole; TER, p-terphenyl; RET, retene; BaA, benzo[a]anthracene; BbF, benzo[b]fluoranthene; BkF, benzo[k]fluoranthene; BeP, benzo[e]pyrene; BaP, benzo[a]pyrene; IcdP, indeno[1,2,3-cd]pyrene; DahA, dibenzo[a,h]anthracene; and BghiP, benzo[ghi]perylene.
Table 4. Mean daily intakes (ng kg−1 day−1) of plasticizers via dust ingestion, inhalation, and dermal contact for children and adults.
Table 4. Mean daily intakes (ng kg−1 day−1) of plasticizers via dust ingestion, inhalation, and dermal contact for children and adults.
Living rooms6.12914776573.01258890493818543723,333
Living rooms0.0000.02550.2170.0030.0020.2480.1380.0050.0120.652
Living rooms0.0071.058.930.1440.08410.25.680.2130.50226.8
Living rooms0.65597.983213.47.8295252919.946.82500
Living rooms0.0000.0090.0780.0010.0010.0900.0500.0020.0040.235
Living rooms0.0010.1471.250.0200.01171.430.7940.0300.0703.75
Reference doses for oral exposure, RfD (mg kg−1 day−1) recommended by the United States Environmental Protection Agency: DEP, 0.8; DBP, 0.1; BBP, 0.2; DEHA, 0.6; DEHP, 0.02; and DNOP, 0.01. Reference doses for other compounds or other exposure pathways are not available.
Table 5. Incremental lifetime cancer risk from human exposure to PAHs in PM10 resuspended from household dust via ingestion, inhalation, and dermal absorption.
Table 5. Incremental lifetime cancer risk from human exposure to PAHs in PM10 resuspended from household dust via ingestion, inhalation, and dermal absorption.
Cancer riskKitchensBedroomsLiving Rooms
ILCRingChildren9.2 ×10−67.2 × 10−69.4 × 10−6
Adults1.4 × 10−51.1 × 10−51.4 × 10−5
ILCRinhChildren1.4 × 10−101.1 × 10−101.4 × 10−10
Adults6.9 × 10−105.4 × 10−109.0 × 10−10
ILCRderChildren3.6 × 10−82.8 × 10−83.7 × 10−8
Adults7.1 × 10−85.6 × 10−87.3 × 10−8
Total ILCRChildren9.3 × 10−67.2 × 10−69.4 × 10−6
Adults1.4 × 10−51.1 × 10−51.4 × 10−5

Share and Cite

MDPI and ACS Style

Vicente, E.D.; Vicente, A.; Nunes, T.; Calvo, A.; del Blanco-Alegre, C.; Oduber, F.; Castro, A.; Fraile, R.; Amato, F.; Alves, C. Household Dust: Loadings and PM10-Bound Plasticizers and Polycyclic Aromatic Hydrocarbons. Atmosphere 2019, 10, 785.

AMA Style

Vicente ED, Vicente A, Nunes T, Calvo A, del Blanco-Alegre C, Oduber F, Castro A, Fraile R, Amato F, Alves C. Household Dust: Loadings and PM10-Bound Plasticizers and Polycyclic Aromatic Hydrocarbons. Atmosphere. 2019; 10(12):785.

Chicago/Turabian Style

Vicente, E. D., A. Vicente, T. Nunes, A. Calvo, C. del Blanco-Alegre, F. Oduber, A. Castro, R. Fraile, F. Amato, and C. Alves. 2019. "Household Dust: Loadings and PM10-Bound Plasticizers and Polycyclic Aromatic Hydrocarbons" Atmosphere 10, no. 12: 785.

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Metrics

Back to TopTop