Characteristics of Indoor and Outdoor Fine Phthalates during Different Seasons and Haze Periods in Beijing

Received: April 28, 2018
Revised: September 18, 2018
Accepted: September 20, 2018

Download Citation: RIS|BibTeX|https://doi.org/10.4209/aaqr.2018.03.0114  

Cite this article:

Wang, Y., Ding, D., Shu, M., Wei, Z., Wang, T., Zhang, Q., Ji, X., Zhou, P. and Dan, M. (2019). Characteristics of Indoor and Outdoor Fine Phthalates during Different Seasons and Haze Periods in Beijing. Aerosol Air Qual. Res. 19: 364-374. https://doi.org/10.4209/aaqr.2018.03.0114

HIGHLIGHTS

• The indoor phthalate concentration was generally higher than that outdoors.
• Much higher concentrations of fine phthalates were found during the haze days.
• As age increased, the risk of DEHP carcinogenesis gradually declined.
• DEHP inhalation and skin exposure were highest in 1–2-year-old children.

Keywords: PM2.5; Indoor air; Exposure.

INTRODUCTION 

Phthalic acid esters (PAEs, also called phthalates) are an important class of organic compounds and are widely used as plasticizers to improve the flexibility and workability of polymeric materials such as those in toys, food packaging, building materials, medical devices, electronics, vinyl flooring and wallpaper. Lower molecular weight PAEs are also used as solvents and/or fragrance in cleaning agents, lubricants, personal care products (e.g., nail polish, hair spray, soap, and shampoo), repellents, and hundreds of other products (Guo and Kannan, 2013; Škrbic et al., 2016). Growing concern about phthalates is related to their toxicity, in particular to their endocrine disrupting activity. Toxicological researches in both animal and human investigated that the most sensitive health outcomes following exposure to some PAEs were the adverse reproductive development effects, which included shortened anogenital distance observed in newborn boys, shortened pregnancy, destroyed testicles, lower sex and thyroid hormones, and reduced sperm quality observed in adults (Hauser et al., 2006; Axelsson et al., 2015; Al-Saleh et al., 2017). Exposure to some PAEs has been associated with the allergic symptom, asthma and neurodevelopmental problems in children (U.S. EPA, 2012; Bamai et al., 2016). Besides, the International Agency for Research on Cancer (IARC) (IARC, 1982) has listed DEHP as a possible carcinogen to humans (Group 2B; 1982). Due to the adverse health effects, many countries and organizations list some PAEs as priority control pollutants. For instant, the U.S. Environmental Protection Agency (U.S. EPA) has listed dimethyl phthalate (DMP), diethyl phthalate (DEP), dibutyl phthalate (DBP), butyl benzyl phthalate (BBP), phthalic acid (2-ethylhexyl) phthalate (DEHP), and dioctyl phthalate (DnOP) among 129 priority pollutants. DMP, DBP, DnOP are on the Chinese environmental priority pollutants blacklist, and DBP, DnOP, DEHP and BBP are also on the list of priority pollutant of the European Community (EC, 1997).

Because they are not chemically bound to polymeric matrices and thus easily released into the environment, PAEs have become the world’s most widely distributed class of persistent environmental pollutants. Some studies have reported PAEs as a ubiquitous, even dominant class of organic compounds in outdoor airborne particulates worldwide—for example, in cities in India (Fu et al., 2010), France (Teil et al., 2006), the Netherlands (Peijnenburg and Struijs, 2006), and China (Wang et al., 2006; Wang et al., 2008; Kong et al., 2013; Wang et al., 2018)—as well as in indoor gases, aerosols and home dusts (Wensing et al., 2005; Teil et al., 2006; Weschler et al., 2008; Pei et al., 2013; Zhang et al., 2014; Wang et al., 2014; Song et al., 2015).

As the capital of China, Beijing has experienced frequent severe haze pollution in recent years. Although extensive studies have investigated the chemical composition, sources, formation mechanisms, and health effect of haze episodes in Beijing (Sun et al., 2013, 2015, 2016), few studies have been conducted on atmospheric PAEs and none is involved in PAEs during haze period to our knowledge. People usually spend 80–90% of their time each day indoors (Castro et al., 2011), and indoor PM_2.5 concentration is strongly affected by outdoor pollution (Zhao et al., 2015). Therefore, studying indoor and outdoor PAE pollution in Beijing is critical. This study investigated the distribution, pollution level, and main likely sources of PAEs in indoor and outdoor airborne fine particulates (fine PAEs) during haze and non-haze period to provide a basis for improving indoor air quality.

EXPERIMENTAL METHODS

 
Sampling

Paired outdoor and indoor PM_2.5 samples were collected simultaneously at the Beijing Municipal Institute of Labor Protection building during the winter of 2015 and spring of 2016 (Fig. 1). This was a typical residential site with urban traffic, adjacent to busy roads to the south and west. The outdoor sampling point was the rooftop of the building, and the indoor sampling point was a room (third floor) in which no staff work. The room was 20 m² and the sampling height 1.2 m, simulating breathing zone height for favorable comparison of indoor and outdoor pollution. Sampling was conducted over 1 day (a period of 23 h) or 2 days (a period of 48 h) when the air quality was excellent using the URG-3000 ABC particle sampler (with PM_2.5 sampling head, 8.35 L min^–1) outdoors and the BUCK VSS-5 personal sampling pump (with PM_2.5 sampling head, 4 L min^–1) indoors. Both blank and sample filters were packed in aluminum foil and stored at –18°C in the dark prior to gas chromatography–mass spectrometry (GC-MS) analysis.

Fig. 1. Sample location map.

Extraction and GC-MS Analysis

Aliquots of approximately 2.1 cm² each were punched from sample filters and spiked with a known amount of perdeuterated tetracosane (n-C₂₄D₅₀) and perdeuterated benzoic acid (C₇HO₂D₅) as recovery standards prior to extraction (The amount of aliquots used was based on the organic carbon (OC) content, to ensure that the total OC mass in the aliquots was more than 1 mg). The aliquots were cut into pieces and extracted using three successive 10 mL solution of dichloromethane/methanol (1:1; v/v) at three 15-minute intervals using ultrasonic extractors. The solventextracts were combined and concentrated to approximately 5 mL using a rotary evaporator and blown to near-dryness with pure nitrogen gas; a 0.4-mL dichloromethane/acetonitrile (3:1, v/v) solution was then added. The solutions were analyzed using GC-MS (6890 Plus GC, 5973N MSD; Agilent, Santa Clara, CA, USA) with a DB-5MS capillary column (30 m × 0.25 mm × 0.25 µm; J&W Scientific, Inc.). The temperature program for GC consisted of a 10-min isothermal hold at 60°C followed by a ramp of 6°C min^–1 to 300°C, and finally a 50-min isothermal hold at 300°C. MS was performed in electron ionization mode, with electron energy of 70 eV and a mass scanning range of 50–550 amu. Scanning modes included total ion current and selected ion monitoring. Data were gathered and processed using ChemStation software. GC-MS response factors were determined using authentic standards (1000 µg mL^–1, Lot 213081263, AccuStandard) for six PAEs, namely, DMP, DEP, DBP, BBP, DEHP, and DnOP. Recovery for most of the samples was in the range 60–120%. Reagent blank, instrument blank, and sampling blank were analyzed along with the samples, and the concentrations of samples in this study were corrected according to the recovery rate and blank.

RESULTS AND DISCUSSION

 
Concentration Levels and Seasonal Distribution of Fine PAEs in Beijing

The average concentration of indoor fine PAEs at the investigated location was 1288.03 ± 379.02 ng m^–3 and 856.59 ± 217.93 ng m^–3 during spring 2016 and winter 2015, respectively, whereas the outdoor fine PAE concentration was 455.82 ± 97.65 ng m^–3 in spring and 815.22 ± 156.47 ng m^–3 in winter. Compared with the concentrations of PAE aerosols in other studies, the concentrations of indoor PAEs in this study were considerably lower than those in indoor Hangzhouair (Pei, 2013), comparable to those in California (Rudel et al., 2003) and Berlin (Fromme et al., 2004), and much higher than those in Paris (Teil et al., 2006) (Table 1). Similarly, the concentrations of outdoor fine phthalates in this study were higher than those in Tianjin (Kong et al., 2013) and Taizhou (Gu et al., 2010). The comparison in Table 1 shows that the concentration of fine PAEs in Beijing was high. Few studies have been conducted on the PAEs present in indoor fine particles, and this is thus a new research avenue. Given the high PAE concentration and the stronger toxic effect of fine particles compared with large particles, more attention should be paid to fine PAEs in Beijing to increase knowledge about them and reduce the associated health risks.

Unlike other organic materials with fine particles, PAEs had an average concentration that was higher indoors than outdoors, especially during spring. This points to typical indoor emissions sources such as volatilization from plastics and paint. This means that phthalates in the indoor environment were primarily from indoor sources. This result was consistent with an analysis of phthalate contaminants in indoor air by Pei (2013), who stated that the indoor concentration was generally higher (being 124.7 times that of the outdoor concentration).

Different seasonal patterns were found regarding indoor and outdoor fine PAE concentrations during winter 2015 and spring 2016. For indoor fine PAEs, the average concentration in spring (1288.03 ± 379.02 ng m^–3) was higher than in winter (856.59 ± 217.93 ng m^–3). The outdoor fine PAE concentration showed the opposite seasonal pattern, being lower in spring (455.82 ± 97.65 ng m^–3) and higher in winter (815.22 ± 156.47 ng m^–3). These different seasonal distribution patterns were likely due to differing sources and dispersion conditions. With the temperature rising in spring, part of the more highly saturated vapor pressure components, such as DMP, evaporated, and the concentration in the particle phase thus decreased. Seasonal changes in the outdoor concentration were mainly due to industrial pollution sources, incineration of plastic waste, and the use of thin films for agricultural greenhouses, combined with adverse atmospheric dispersion conditions in winter, which cause phthalate concentrations to be maintained at a high level. Beijing was often windy in the spring, which was conducive to the spread and dilution of organic pollutants and was also why PAEs were difficult to collect from particles in the spring.

The effect of meteorological conditions was also assessed in Table 2 and Fig. 2. By studying the relationship between temperature and humidity in terms of concentrations of fine PAEs, we discovered that temperature and humidity had no significant correlation with fine PAEs concentration in winter or spring (P > 0.05). Therefore, temperature and humidity had little effect on the concentrations of PAEs in PM_2.5 in this study. Outdoor wind speed was low in winter and significantly decreased during periods of haze (Fig. 3). Convection slowed down, which was not conducive to the dispersion of PAEs. Therefore, outdoor PAE concentration was high in winter. In spring, however, the wind speed was higher and the atmospheric convection more intense, which could also explain why outdoor PAE concentration was higher in winter than in spring.

Fig. 2. The effect of temperature and humidity on mass of phthalate concentrations.

Fig. 3. Mass of phthalate concentrations in indoor and outdoor PM_2.5 during haze and meteorological condition effects.
 

Comparison of Indoor and Outdoor PAE Concentrations during Haze Periods

The indoor and outdoor phthalate concentrations in different polluted weather conditions on different dates are shown in Fig. 3. During the pre-haze period of winter, the indoor fine phthalates concentration was 915.02 ± 231.40 ng m^–3, whereas it was 470.39 ± 106.00 ng m^–3 outdoors. During the period of haze, the indoor concentration was slightly higher at 940.15 ± 223.46 ng m^–3, whereas the outdoor concentration was considerably higher at 1549.03 ± 284.58 ng m^–3. In the pre-haze period of spring, the indoor and outdoor fine phthalate concentrations were 1235.76 ± 396.78 ng m^–3 and 363.20 ± 80.84 ng m^–3, respectively, whereas during the haze period these values were 1389.68 ± 384.55 ng m^–3 and 574.27 ± 136.67 ng m^–3.

During the entire haze period of spring, the indoor and outdoor PAE concentration trends were extremely similar (Fig. 3). In hazy conditions, indoor and outdoor fine PAE pollution showed a clear upward trend, with a maximum PAE concentration of 1995.56 ng m^–3. Fig. 4 shows that DEHP was the PAE monomer with the most significantly increased concentration of indoors in the spring haze, as for outdoors the increase in PAE concentration was mainly due to DEHP, with a small increase in the DMP concentration.

Fig. 4. Mass concentration of phthalate monomers in indoor and outdoor PM_2.5 during haze periods.

Fine PAE concentrations in the period of winter haze showed a different trend. In the pre-haze period, PAE concentration was generally higher indoors than outdoors, with the concentrations having the same trend. When the haze period began, the outdoor PAE concentration gradually increased, whereas the indoor concentration showed no obvious change. At the end of the haze period, the indoor and outdoor PAE concentrations gradually returned to their pre-haze synchronization trend, with indoor concentration higher than that outdoors. During the winter haze period, DMP, DBP, and DEHP were the PAE monomers with significantly increased concentrations outdoors, whereas DEHP concentration was increased the most significantly indoors.

As shown in Fig. 4, outdoor PAEs in the spring haze were mainly DMP and DEHP, whereas in the winter haze, the main PAEs were DMP, DBP, and DEHP. During the winter haze, the concentrations of pollutants in PM_2.5 were increased due to adverse weather factors, which caused the outdoor PAE concentration to gradually increase. No obvious change in indoor PAE concentration occurred during the winter haze, mainly because people would close their windows on hazy days, which caused poor air circulation in the indoor environment. This indicates that indoor PAE concentration was mainly affected by indoor PAE sources.

 
Molecular Monomer Distributions and Possible Sources of Phthalates in PM_2.5

Concentrations of indoor and outdoor PAE monomers (DMP, DEP, DEBP, BBP, DEHP, and DnOP) are shown in Table 3 and Fig. 5, and significant characteristics of the main monomers are given. DBP and DEHP were the predominant monomers, followed by DMP, with other PAE monomers being relatively minor. In spring, DBP accounted for 72% of indoor and 39% of outdoor PAEs, whereas DEHP accounted for 27% of indoor and 48% of outdoor PAEs. Similarly, DBP accounted for 62% of indoor and 45% of outdoor PAEs in winter, whereas DEHP accounted for 32% of indoor and 30% of outdoor PAEs. The next highest concentrations were those of DMP and DEP, and the concentration of DnOP was minimal. This means that DBP and DEHP are the two main components of PAE pollution, which is likely because these are the two most widely used plasticizers in China.

Table 3. Concentration of PAEs in indoor and outdoor PM_2.5 in different seasons (mean ± standard deviation, unit: ng m⁻³).

Fig. 5. Mass concentration of phthalate monomers in indoor and outdoor PM_2.5 during winter and spring.

Differences in the main PAE monomer pollutants were primarily due to the local major industries, consumer goods, and consumption levels. Outdoor DBP and DEHP were mainly from the plastic products industry and various other sources of fuel emissions. According to a statistical survey of plasticizer-related industries in China, current plasticizer production and consumption are mainly based on DEHP and DBP, with these PAEs accounting for more than 80% of the total. The lower occurrence of DEP, BBP, and DnOP in PM_2.5 samples may be related to their low production and minimal use as plasticizers. Therefore, we further analyzed the major PAE pollution monomers (DMP, DBP, and DEHP).

Table 4 shows correlations between different PAE monomers indoors and outdoors. In spring, no obvious correlations (p > 0.05) were found between different PAE monomers either indoors or outdoors, whereas in winter, although there were no significant correlations (p > 0.05) between different PAE monomers indoors, the correlation coefficient of paired outdoor concentrations of DMP and DBP was 0.742 (p < 0.05), and it was 0.954 (p < 0.05) for paired concentrations of DBP and DEHP. After excluding paired concentrations with the maximum or minimum values of DMP and DEHP in outdoor conditions in winter, paired concentrations of DMP and DEHP were also significantly correlated (p < 0.05), with the correlation coefficient changing from 0.662 to 0.812. The significant correlations between DMP, DBP, and DEHP suggest a strong possibility that these PAEs are released from the same sources in winter outdoor conditions, whereas indoor sources of PAEs in Beijing are more diverse and complicated.

Numerous plastic products were present in the room investigated in this study, and the desk surface had a plastic paint veneer. The interior decor also revealed various plastic hot-melt adhesives, coatings, and ink additives, which are major sources of indoor PAEs. 

Exposure Risk Assessment for Phthalates in PM_2.5

PAEs are an endocrine disrupter with an estrogenic function. U.S. studies have shown that trace amounts of DEHP have a major effect on human health and strong carcinogenicity at low concentrations. However, no indoor air phthalate quality standard has yet been defined, either in China or abroad, and the exposure hazards of indoor air phthalates have not been clearly determined. The Integrated Risk Information System in the United States evaluates the inhalation exposure carcinogenicity of phthalates in the air. China’s indoor air phthalate concentration is relatively high; because such pollution represents a potential hazard for the human body, the health risk should be taken seriously.

As a class of anti-androgen endocrine disruptors and B2 carcinogens, phthalates mainly pose anti-androgen effect and carcinogenic risks, with a greater risk of carcinogenesis from exposure to phthalates in indoor air. In this study, carcinogenic risk was estimated for exposure to phthalates in indoor air in Beijing. The carcinogenic risk value R can be calculated using the following equation (U.S. EPA, 2000): 

where R is the phthalates carcinogenic risk value, q is the carcinogenic slope factor and equal to 0.014 (mg kg^–1 d^–1)^–1 for DEHP (U.S. EPA, 1997a), and E is the PAE exposure in mg kg^–1 d^–1).

PAEs enter the human body mainly through inhalation and skin contact; inhalation exposure was calculated using the following equation (U.S. EPA, 1997b):

where IE is inhalation exposure (unit: µg kg^–1 d^–1), C_i is the total concentration of PAEs (µg m^–3), IR is the respiration rate (m³ d^–1), CF is the unit conversion factor (equal to 1/24), ED is exposure time (h d^–1), and BW is the mass of the exposed body (kg).

Skin exposure was calculated using the following equation (U.S. EPA, 1992): 

where DE is skin contact exposure (µg kg^–1 d^–1), C_i is the gas pollutant concentration (µg m^–3), SA is the skin surface area (m²), f_SA is the ratio of skin to air contact (a value of 0.25 is assumed for a person wearing short pants), P is the total permeability coefficient of the skin (P = 580 cm h^–1for DEHP; Xu et al., 2009), and F is the unit conversion factor (the value is 0.01); the remaining parameters are the same as in Eq. (2).

The gas PAEs concentration was estimate using the following equation: 

where C_g is the gas pollutant concentration (µg m^–3), C_p is the fine pollutant concentration (µg m^–3), K is the ratio factor between gas and fine phase of DEHP (Pei, 2013).

The indoor DEHP inhalation and skin exposure for different age groups, calculated using Eqs. (1)–(4), are shown in Tables 5 and 6. Parameters such as age, respiration rate, body weight, exposure time, and skin surface area were referenced from the U.S. EPA’s recommended values.

The results showed that DEHP inhalation exposure was the highest for children aged 1–2 years at 0.29 µg kg^–1 d^–1 in winter and 0.36 µg kg^–1d^–1 in spring, followed by children aged 2–3, for whom inhalation exposure was 0.25 µg kg^–1 d^–1 in winter and 0.31 µg kg^–1 d^–1 in spring. From the 2–3 to 16–21 year age groups, DEHP inhalation exposure gradually decreased with age. Likewise, children aged 1–2 had the highest skin exposure, which was 0.31 µg kg^–1 d^–1 in winter and 0.38 µg kg^–1 d^–1 in spring, slightly higher than the inhalation exposure; DEHP skin exposure was lower for children aged 2–3. The calculation results showed that the exposure value was lower than the U.S. EPA limit of 20 µg kg^–1 d^–1, but this does not mean that PAEs pose no health hazards. Multiple PAEs may exert synergistic effects on health. Additionally, this study did not consider food, drinking water, and other media through which PAEs could enter the body; nor did it consider the intake of PAEs in children from interaction with toys. Methods of determining exposure to PAEs through these pathways should be further explored. Our results showed that total exposure to DEHP in children was significantly higher than in adults, with the most serious health hazards for children aged 1–2.

A carcinogenic risk value of under 1 × 10^–6 indicates that the risk is small, whereas a risk value of more than 1 × 10^–4 indicates a high risk, requiring careful attention and appropriate measures. In this study, we found that DEHP exposure in winter and spring in Beijing had the strongest effect on children aged 1–2, and their carcinogenic risk in winter and spring was 8.47 × 10^–6 and 10.50 × 10^–6, respectively (Table 7). Thus, there was some risk of cancer from DEHP exposure in 1–2-year-old children. With increasing age, the risk of cancer was discovered to gradually decrease. People 16–21 years old or older than 21 had a small risk of exposure—3.30 × 10^–6 in winter and 4.09 × 10^–6 in spring—but this is still higher than the threshold carcinogenic risk value. Indoor phthalate pollution thus poses a serious carcinogenic risk because it exerts anti-androgen effects on human health; attention should be paid and effective measures proposed to control PAE contamination.

CONCLUSION

(1) The concentration of phthlates in PM_2.5 in Beijing is relatively high and exhibits some seasonal variation. During the time of our study, it was significantly higher both indoors and outdoors during periods of haze than before the haze, reaching a maximum concentration (indoors) of 1995.56 ng m^–3. Furthermore, the concentration indoors was generally higher than outdoors, indicating that the main sources of phthalates in indoor environments were indoor. The primary PAEs found outdoors during spring haze were DMP and DEHP, whereas those during winter haze were DMP, DBP, and DEHP. During winter haze, the amount of phthalates in the PM_2.5 significantly increased, but the indoor concentration did not significantly change, supporting the conclusion of indoor sources.

(2) By calculating the DEHP exposure of children and adults via inhalation and skin contact, we discovered that 1–2-year-old children were the highest risk group in Beijing during winter and spring, with a level of exposure exceeding that associated with the minimum risk. From the age of 2 or 3, the risk of carcinogenesis decreased with age; according to our calculations, the level of exposure was lower than the U.S. EPA limit of 20 µg kg^–1 d^–1. Nevertheless, the carcinogenic risk and anti-androgen effects of PAEs represent a serious risk to human health, and the synergistic effects of multiple PAEs, due to the variety of indoor contact media, also impact the human body. Thus, effective measures are required to control PAE pollution.

ACKNOWLEDGEMENTS

This work was partly supported by the Beijing Municipal Natural Science Foundation (L150001) and the Beijing Academy of Science and Technology’s Green Shoots Plan (201706).

Aerosol Air Qual. Res. 19 :364 -374 . https://doi.org/10.4209/aaqr.2018.03.0114