A review of polychlorinated biphenyls (PCBs) pollution in indoor air environment

ABSTRACT

Polychlorinated biphenyls (PCBs) were widely used in industrial production due to the unique physical and chemical properties. As a kind of persistent organic pollutants, the PCBs would lead to environment pollution and cause serious problems for human health. Thus, they have been banned since the 1980s due to the environment pollution in the past years. Indoor air is the most direct and important environment medium to human beings; thus, the PCBs pollution research in indoor air is important for the protection of human health. This paper introduces the industrial application and potential harm of PCBs, summarizes the sampling, extracting, and analytical methods of environment monitoring, and compares the indoor air levels of urban areas with those of industrial areas in different countries according to various reports. This paper can provide a basic summary for PCBs pollution control in the indoor air environment.

Implications: The review of PCBs pollution in indoor air in China is still limited. In this paper, we introduce the industrial application and potential harm of PCBs, summarize the sampling, extracting, and analytical methods of environment monitoring, and compare the indoor air levels of urban areas with industrial areas in different countries according to various reports.

Introduction

The application and harm of PCBs

Polychlorinated biphenyls (PCBs) comprise a light yellow or deep yellow oily liquid with properties of insulating ability, thermal stability, resistance to acids, oxidation, hydrolysis, and flame resistance. Due to these unique physical and chemical properties, PCBs have been widely used in many products, especially in transformers and power capacitors (Fitzgerald et al., 2007). Until 1971, 61% of PCBs were used in closed electrical systems, 13% were used in nominally closed systems, and 26% were used in open-end applications. After 1971, almost 100% of all PCBs produced were used in closed electrical systems (Erdal et al., 2016). The chemical formula of PCBs is C₁₂H_(0-9)Cl_(1-10). Theoretically, there are 209 kinds of different congeners based on the different position of chlorine atoms in the benzene ring. Actually, the number of PCBs congeners is just about 130 (World Health Organization [WHO]/International Program on Chemical Safety [IPCS], 1992). Generally, there are 41 main monitoring congeners in IUPAC.

PCBs have several trade names in different countries. For example, they are called Aroclor, Pyranol, or Pyroclor in the United States, Phenoclor and Pyralene in France, Clophen and Elaol in Germany, Kanechlor and Santotherm in Japan, Fenchlor and Apirolio in Italy, and Sovol in the Soviet Union. According to previous reports, there were about 1,326,000 tons PCBs produced and used since 1930; the production details are shown in Table 1 (Breivik et al., 2007; Fitzgerald et al., 2007).

Table 1. Total PCBs production as reported in the literature (in tons).

Manufacturer	Country	Producing Time	Total (t)	Percent (%)
Monsanto	United States	1930–1977	641,246	48.4
Bayer AG	West Germany	1930–1983	159,062	l2
orgsteklo	Former Soviet Union	1939–1990	141,800	10.7
Prodelec	France	1930–1984	134,654	10.2
Monsanto	United Kingdom	1954–1977	66,542	5
Kanegafuchi	Japan	1954–1972	56,326	4.2
orgsintez	Former Soviet Union	1972–1993	32,000	2.4
Caffaro	Italy	1958–1983	31,092	2.3
S.A.Cros	Spain	1955–1984	29,012	2.2
Chemko	Czechoslovakia	1959–1984	21,482	1.6
Xi An	China	1960–1979	8,000	0.6
Mitsubishi	Japan	1969–1972	2,461	0.2
Electrochemical Company	Poland	1966–1970	1,000	<0.1
Zaklady Azotowe	Poland	1974–1977	679	<0.1
Geneva Industries	United States	1971–1973	454	<0.1
Total	Whole world	1930–1993	1,325,810	100

During the past 60 years, in total 1,325,810 tons PCBs were manufactured and almost 50% of them came from America. The main PCB congeners in all industrial products were PCBs containing two to seven chlorine atoms. Though the production of PCBs has been stopped since the 1980s, their considerable usage during the past two decades has led to their ongoing detection in humans, wildlife and many environment media (Knobeloch et al., 2009; Sofuoglu et al., 2013). Every year, there is about 1000 tons of PCBs entering into the environment, mainly by leaking in the process of transporting and improper disposal of PCBs production, together with the combustion of solid waste and fossil fuels (Choi et al., 2008a; Kim et al., 2011; Mari et al., 2008; Xing et al., 2005). These concerns rise not only due to the large amounts of PCBs released into the environment, but also because of their developmental toxicity, cancer risk, and dangers to human and wildlife health (Di and Chang, 2011). Because they are lipophilic, hydrophobic, and persistent, it is easy for PCBs to gather in fatty tissue through the food chain (Fitzgerald et al., 2007). Under a long time of exposure to PCBs, the human skin and liver are most vulnerable to damage, as well as the gastrointestinal tract, immune system, and nervous system (Mukerjee 1998). According to the experiments on animals, some PCB congeners have carcinogenicity and this would be enhanced in cooperation with other chemicals. In 1968 and 1979, there were two serious PCBs pollution events: PCBs in a heat exchanger leaked into edible oil and resulted in 10,000 people poisoned (Tsukimori et al., 2008). Of the 209 PCBs congeners, the 12 dioxin-like and 6 indicator PCBs were usually of concern due to their high toxicity and high concentrations in the environment. In recent years, 3,30-dichlorobiphenyl (PCB-11), a non-Aroclor PCB congener, has attracted great attention due to its wide presence and potential emission sources associated with pigment yellow production (Choi et al., 2008b, Lehmler et al., 2010, Ross et al., 2008, Shang et al., 2014).

Research range and status of PCBs

As PCBs have been detected from widely environment and biological media, the pollution studies have become a research hotspot. From searching “indoor air PCBs” as “key words” in SCIENCEDIRECT system, the search results are shown in Figure 1. In the recent decade, the number of research papers has been increasing (searching on April 16, 2016).

Figure 1. Indoor air PCBs papers published.

According to the present studies, the research on indoor air PCBs mainly focuses on the levels of PCBs in air samples, passive sampling comparison with active sampling, human exposure, potential sources of PCBs, spatial distribution, and the effects of home characteristics on PCBs (Asante et al., 2011; Heinzow et al., 2007; Heo and Lee 2014; Klees et al., 2015; Li et al., 2012; Lu et al., 2015; Lyng et al., 2015; Moussaoui et al., 2012; Tasdemir et al., 2012; Wang et al., 2013).

Sampling extracting and analytical methods

Sampling

There are two sampling methods for indoor air PCBs analysis: passive air sampling and active air sampling. A high-volume air sampler acts as the main tool of active air sampling; it can collect a large volume of air sample in a short time. However, this kind of sampler needs electrical power, which would cause much noise during sampling. Also, the price of a high-volume sampling machine is relative high. (Tuduri et al., 2012) The schematic of the high-volume air sampler is shown in Figure 2. High-volume air sampling media are mounted on tripods and high enough from the ground to keep the sampling height in the human breathing area. The sampling media consist of a glass fiber filter (GFF) and polyurethane foam (PUF) plug. The GFF is used to capture particulate-bound PCBs and the PUF plug is used to capture vapor-phase PCBs. Passive air samplers adsorb atmospheric PCBs on a polyurethane foam (PUF) disk, and have been developed in recent years as complementary tools for monitoring of atmospheric POPs (Hogarh et al., 2012). Compared with high-volume active air samplers, they are economical and do not require electricity. Also, passive air samplers could provide the average concentrations of PCBs during the sampling time because they could work for a long time without noise. The passive air sampler is also a good tool to monitor and evaluate point-source pollution (Hazrati and Harrad, 2006; Klánová et al., 2006). The schematic of a passive air sampler is shown in Figure 3.

Figure 2. Schematic of high-volume active air sampler.

Figure 3. Schematic of PUF disks passive air sampler.

Samples extraction

Due to the low PCBs concentrations in environment samples and interference from the matrix, the samples must be pretreated and combined with advanced analysis methods. The main extraction technologies for PCBs are solvent extraction (SE), solid-phase extraction (SPE), solid-phase microextraction (SPME), ultrasonic extraction (USE), microwave extraction (MAE), and accelerated solvent extraction (ASE).

SE could totally extract the target compound, but it takes too much time. Also, during the extraction progress, it is easy for the semivolatile substance to run off. SPE has high separation and recovery efficiency. It is also easy to control with low cost. SPE has already been applied in the samples treatment of food, water, soil, and plasma (Beyer and Biziuk 2008; Centi et al., 2007). SPME is a multifunctional technology that can sample from environment as well as extract and preconcentrate the samples. SPME has the advantage in the analysis of liquid samples without an organic solvents multiple purification process, so it can reduce the environment pollution and the damage on analysts. But SPME hasn’t been included by the U.S. EPA because of the low sample recovery and poor repeatability (Picó et al., 2007; Popp et al., 2005). When a large number of samples need to be treated at the same time, USE is a good choice. However, after extracting, the samples still need to be further centrifuged with an organic phase, so it would cost more time and solvent (EPA, 2007). The MAE system can heat different components in samples and there are many alternative solvents because of the affinity of solvents limits. In recent years, MAE was applied to PCBs analysis in blubber, soil, sediments, and mussels (Fujita et al., 2009). ASE is the method that solvent extracts solid and semisolid samples under high temperature (50–200ºC) and high pressure (6.89–20.68 MPa). This technology can remove some fatty compounds and separate target components by adding solid-phase sorbents into the extraction pool, so it can achieve extraction and meanwhile purification. However, the equipment costs too much (Garrido Frenich et al., 2005).

Samples analysis

Generally, PCBs can be analyzed by gas chromatography (GC) and gas chromatography–mass spectroscopy (GC-MS) methods. PCBs are often measured by gas chromatography–electron capture detector (GC-ECD) with a high-resolution quartz capillary column. It has the advantages of high sensitivity and low detection limit as well as low costs (Wiberg et al., 2007). However, we can’t know the information of molecular structure because the chromatography just conducts qualitative analysis according to the retention time of the test substance. Besides, the ECD detector requires purity of samples, as the sulfide in samples would do harm to ECD (Valsamaki et al., 2006). In recent years, comprehensive two-dimensional gas chromatography (GC-GC) has been applied in the analysis of PCBs; the separation capacity increased a lot and more complex compounds also can be analyzed. Compared with GC-ECD, GC-MS is more accurate. GC-MS can get the characteristic ions of PCBs and the abundance information of chlorine atoms, thus providing more reliable data (Barro et al., 2009). Also, GC-MS can be divided into high-resolution (HR) GC–low-resolution (LR) MS and HRGC–HRMS, and the sensitivity and accuracy of HRGC–HRMS are both higher than for HRGC–LRMS. Recently, in order to improve the PCBs analysis accuracy by HRGC–LRMS, the samples usually are combined with an isotope dilution method (Castro-Jiménez et al., 2009).

The analysis accuracy of PCBs has been improved with technology development. In recent years, the method has turned from aggregate analysis to homologue analysis, and then developed to monosomic analysis. Meanwhile, the analysis equipment is also improved. It turns from the GC-ECD to GC-LRMS, and then develops to GC-HRMS. Thus, the sensitivity and accuracy are improved continuously.

Pollution situation

Indoor air PCBs in urban area

Air transmission is the primary route of transmission of PCBs from air emission sources into the terrestrial and aquatic ecosystems. PCBs in the atmosphere have two forms, which are gas-phase PCBs and particulate-phase PCBs. Indoor air PCBs concentrations in urban area reported for various cities around the world are compared in Table 2.

Table 2. Indoor air PCBs concentrations of urban area in different countries.

Site	Number of Congeners	Samples Type	Concentration (pg m^−3)	Remark
Milwaukee (Wethington and Hornbuckle, 2005)	88	Gas Particle	1900 ± 780 50 ± 20	Particulate-phase PCBs account for less than 5% of the total atmospheric concentration.
Yokohama (Kim and Masunaga, 2005)	41	Gas+Particle	62~250	The ratios of tetra-CBs and penta-CBs to the total PCBs concentrations were largest.
Lzmir (Cetin et al., 2007)	41	Gas+Particle	winter (847 ± 610) summer (314 ± 129)	The average gas-phase concentration was about 87% of the total atmospheric PCBs.
Bursa (Sıddık Cindoruk and Tasdemir, 2007)	37	Gas+Particle	492 ± 189	The particulate-phase PCBs accounted for 15% of the total PCBs concentrations.
Lowa (Persoon and Hornbuckle, 2009)	37	Gas+Particle	50~10500	The samples were collected from two laboratories.
New York (Fitzgerald et al., 2011)	12	Gas+Particle	600~233000	The congener with the highest average concentration was PCB-28.
Germany (Don) (Gabrio et al., 2000)	6	Gas+Particle	36200~317400	In Don and Wai the low chlorinated PCB-28 and -52 prevailed and contributed almost 90% to the sum of the six indicator congeners. In Neu the medium chlorinated PCB dominated.
Germany (Wai) (Gabrio et al., 2000)	6	Gas+Particle	612000~2131000
Germany (Neu) (Gabrio et al., 2000)	6	Gas+Particle	15400~202500

In the urban areas, there are a lot of PCB-containing products applied in our daily life, and the indoor air PCBs levels should be measured for human health. As shown in Table 2, the lowest levels were detected in samples from Yokohama (62–250 pg m⁻³) and the highest levels were found in samples from the three schools in Germany (Don: 36,200–317,400 pg m⁻³, Wai: 612,000–2,131,000 pg m⁻³, Neu: 15,400–202,500 pg m⁻³). Surprisingly, the indoor air PCBs concentrations in German schools are so high that they are even hundreds of times higher than domestic levels. The three schools (Don, Wai, and Neu) were constructed between 1960 and 1980 in Germany, containing elastic sealant material with high portions of technical mixtures of PCBs (Benthe et al., 1992). Also, continuous PCB emission from these sources leads to increasing secondary contamination of floors, walls, and ceilings, which causes a steadily increasing indoor air contamination (Gabrio et al., 2000). In Don and Wai, the low chlorinated PCB 28 and 52 prevailed and contributed almost 90% of the total PCBs concentrations. This PCB pattern is in accord with another report about a PCB-contaminated school (Gabrio et al., 2000). The indoor air PCBs concentrations of New York, NY, are also relatively high (600–233,000 pg m⁻³). The reason for the high value may be that New York is a big city and the economic center of the United States, so PCB-containing products were applied more than in other cities. In Milwaukee, WI, the particulate-phase PCBs concentration is much lower than the gas-phase PCBs concentration. The concentration of gas-phase PCBs (average ± standard deviation) is 1900 ± 780 pg m^−3, while the particulate-phase PCBs concentration (average ± standard deviation) is 50 ± 20 pg m⁻³. In Yokohama, the dominant congeners were tetra-CBs and penta-CBs, followed by hexa-CBs, hepta-CBs, and octa-CBs. The ratio of the tetra-CBs ranged from 53% to 76% of the total atmospheric concentration. In the urban area of Lzmir, the PCBs concentration in winter (847 ±610 pg m⁻³) was 2.7 times higher than the summer concentration (314 ± 129 pg m⁻³). This may be a result of a stronger room ventilation in the summer. The weaker room ventilation would lead to higher PCBs concentrations in winter. Besides, tri-CBs to hepta-CBs were detected in the samples and their dominance decreased with increasing halogen number. In Bursa, the total atmospheric PCBs concentration was 492 ± 189 pg m⁻³ and lower molecular weight PCB congeners generally dominated in the samples. Concentration of individual PCB congeners (average ± standard deviation) in the indoor air was 1.78 ± 3.28 pg m⁻³ for particulate-phase PCB and 10.59 ± 20.11 pg m⁻³ for gas-phase PCB. In New York, the total indoor air PCBs concentration ranged from 600 to 233,000 pg m⁻³, and the average values of indoor air were 20–400 times higher than those ambient air samples collected from the same residences. This may indicate that the sources of PCBs come from indoors rather than outdoors, such as ballast from lighting, old electric appliances, caulking, elastic sealants, and grouts. Other sources are several secondary sources from paints, mastics, ceiling tiles, flooring, and wall boards. Fitzgerald et al. (2011) evaluated the association between levels of PCBs in residential indoor air and in the serum of older; they found significant correlations for PCB-28 and PCB-105.

In conclusion, in indoor air of urban areas the particulate-phase PCBs just accounted for a very small proportion of the total atmospheric concentration, and most of the PCBs existed in the gas phase. Also, the dominant congeners were lower molecular weight PCB congeners, such as tetra-CBs and penta-CBs.

Indoor air PCBs in industrial areas

Indoor air PCBs concentrations in industrial areas reported from various cities around the world are compiled in Table 3.

Table 3. Indoor air PCBs concentrations of industrial area in different countries.

Site	Number of Congeners	Samples Type	Concentration (pg m^−3)	Remark
Denmark (Frederiksen et al., 2012)	24	Gas+ Particle	168,000–3,843,000	The samples were collected from apartments with PCB-containing sealants.
Lzmir (Cetin et al., 2007)	36	Gas+ Particle	Winter (1371 ± 642) Summer (3136 ± 824)	The average gas-phase concentration was about 87% of the total atmospheric PCB.
Gui Yu (Xing et al., 2009)	34	Gas Particle	4700 ± 4200 14,800 ± 7900	The sampling sites were beside an electronic waste recycling area.
Pohang (Jin et al., 2011)	27	Gas+ Particle	37–104,048	Air samples were collected from 44 sites related to PCB-containing products or wastes.

As can be seen from Table 3, the lowest levels were detected in samples from Lzmir, and the highest levels were found in samples from Denmark (168,000–3,843,000 pg m⁻³) and specifically Copenhagen (mean = 1,052,000 pg m⁻³). Compared with the samples from the uncontaminated apartments (mean = 6030 pg m⁻³), the indoor air PCBs concentrations of contaminated apartments in Denmark were extraordinarily high. And in the air samples, the dominant congeners were PCB-52 and -28, followed by -66, -74, and -101. Generally, PCBs with more than five chlorines were rarely detected. Moreover, the levels of the lower chlorinated congeners from the elastic sealant samples, typically tri- to penta-chlorinated, were strongly correlated with their values in indoor air PCBs levels. It is obvious that the total PCBs concentrations in industrial area were much higher than the indoor air PCBs concentrations in urban site. In Lzmir the summer concentration of the industrial area (3136 ± 824 pg m⁻³) was 10 times higher than the concentration in the urban site (314 ± 129 pg m⁻³), and the winter concentration of the industrial area (1371±642 pg m⁻³) was much higher than the concentration in the urban site (847 ± 610 pg m⁻³) too. Opposite to the situation in the urban site, in the industrial area of Lzmir the PCBs concentration in summer was 1.6 times higher than the winter concentration, which may be because in summer the high temperature lead to more PCBs entering into indoor air from the products. Though the samples from Gui Yu were collected from an residential area of four villages, the indoor air PCBs concentrations were much higher than those samples from the urban area because the villages were beside an electronic waste recycling site. Consisted with the congener pattern of urban air PCBs, the dominant congeners were tri- and tetrachlorinated homologues, and PCB 28 was the most abundant congener. The major congeners in gas-phase PCBs were light PCBs such as PCB-28, -37, -49, and -77; however more heavy PCBs such as PCB-156, -170, -189, -183, and -194 were found in the particulate phase. Surprisingly unlike other reports about urban PCBs concentrations, the particulate-phase PCBs accounted for a larger part than gas-phase PCBs. In Pohang, Guang-Zhu Jin etc. collected samples from 44 sites (9 production, 8 in use, 14 storage, 10 recycling, 1 disposal, and 2 boundary) related to PCB-containing products or wastes including indoor air and ambient air. The concentrations of outdoor air PCBs (mean = 1670 pg m⁻³) were generally lower than indoor air concentrations (mean = 7420 pg m⁻³), which could indicate that the PCB-containing products in those sites were the pollution source. The PCBs levels in air samples were highest at the disposal facility (7336–104,048 pg m⁻³), followed by the production facility (330–25,057 pg m⁻³), the recycling facility (160–17,710 pg m⁻³), the storage site (106–2527 pg m⁻³), and the in-use site (37–2273 pg m⁻³).

In conclusion, the total PCBs concentrations in industrial areas were much higher than the indoor air PCBs concentrations in urban sites, especially in electronic waste disposal sites. Similar congener patterns were found in urban samples and industrial samples: The major congeners in gas-phase PCBs were light PCBs, such as PCB-28 and -37. And heavy PCBs, such as PCB-156 and -170, were found in the particulate phase. This may because light PCBs are more volatile components than heavy PCBs. In urban samples of Milwaukee, WI, the particulate-phase PCBs concentrations are much lower than gas-phase PCBs concentrations. Also, in industrial samples of Lzmir, the particulate-phase PCBs concentrations are even higher than gas-phase PCBs concentrations.

Exposure assessment

The estimation of human nondietary PCBs exposure includes two parts: air inhalation and ingestion of dust. means the daily intake from ingestion of dust and represents the daily intake from air inhalation. The human exposure to PCBs is generally calculated by eqs 1 and 2:

(1)

(2)

In eq 1, is the dust ingestion rate; generally it is set to medium (20 mg/d) and high (50 mg/d) ingestion, respectively (Jones-Otazo et al., 2005). The term is the medium concentration of PCBs in dust of place i (such as offices, living room, bedroom, and so on). In eq 2 means the inhalation rate and is usually set to 16 m³/d (EPA, 2008); represents the concentration in air of microenvironment i, and is the fraction of time spent in place i. Compared with the exposure pathway of dust ingestion, the contribution of inhalation was higher, especially for di-tetra CBs (Tue et al., 2013).

Wang et al. assessed human daily intake of PCBs via indoor and outdoor dust collected from Hong Kong and Guangzhou (Wang et al., 2013). The total PCBs levels in indoor (51.9–264 ng g⁻¹) and outdoor (4.02–228 ng g⁻¹) dust in Guangzhou were found higher than levels in indoor (17.4–137 ng g⁻¹) and outdoor (7.75–114 ng g⁻¹) dust of Hong Kong. The daily intake of PCBs via dust ingestion was in the ranges 0.02–8.95 and 0.37–17.8 ng d⁻¹ in Guangzhou and 0.01–4.95 ng d⁻¹ and 0.16–9.83 ng d⁻¹ in Hong Kong for adults and children, respectively. Wang et al. (2013) used the data from the report of Chen et al. to estimate inhalation exposure for Guangzhou residents. The PCBs levels detected in atmosphere samples in Guangzhou ranged from 0.17 to 2.72 ng m⁻³ (Chen et al., 2006). The estimated inhalation exposure for residents was 3.4–54.4 ng d⁻¹ for adults and 1.7–27.2 ng d⁻¹ for children, respectively. Compared with inhalation, dust ingestion contributed 0.49–10.6% of nondietary PCBs exposure for adults and 12.9–35% for children, indicating the dominant contribution from inhalation especially for adults.

Harrad et al. assessed human exposure to PCBs via dust ingestion, inhalation, and diet under mean and high scenarios of Canadian, New Zealand, UK, and U.S. adults and toddlers (Harrad et al., 2009). The average indoor levels of PCB and exposure scenario in Canada are nearly equal to those in the United States. The average indoor levels of PCB in New Zealand are also nearly equal to those in the United Kingdom. The dietary exposure plays a more important role in the United Kingdom than in New Zealand, indicating that the PCBs levels in foods of the United Kingdom are relatively high. In Canada, the daily intake of PCBs via air inhalation accounted for 53.9% (average = 138 ng d⁻¹) of overall PCBs exposure, while the intake via dust ingestion just accounted for 2.3% (average = 5.8 ng d⁻¹) for adults, and 23.0% (average = 26 ng d⁻¹) from air inhalation and 13.3% (average = 15 ng d⁻¹) from dust ingestion for toddlers. In the United Kingdom, the daily intake of PCBs via air inhalation accounted for 30.5% (average = 150 ng d⁻¹) of overall PCB exposure while the intake via dust ingestion just accounted 0.5% (average = 2.3 ng d⁻¹) for adults, and 12.3% (average = 28 ng d⁻¹) from air inhalation and 2.5% (average = 5.6 ng d⁻¹) from dust ingestion for toddlers. In New Zealand, the daily intake of PCBs via air inhalation accounted for 62.2% (average = 150 ng d⁻¹) of overall PCB exposure while the intake via dust ingestion just accounted 0.5% (average = 1.3 ng d⁻¹) for adults and 34% (average = 28 ng d⁻¹) from air inhalation and 4% (average = 3.3 ng d⁻¹) from dust ingestion for toddlers. Regardless of mean dust intake scenario and high dust intake scenario, dietary exposure is the dominant pathway especially for kids; inhalation follows, then dust ingestion. These results are in accord with the estimates of Lehmann et al. that the extent of inhalation exposure to PCBs in some indoor settings may be at least as large as typical dietary exposure (Lehmann et al., 2015). Inhalation of indoor air was estimated to account for 60.8%, 50.5%, and 34.6% of total exposure, whereas diet accounted for 28.9%, 42.7%, and 62.8% of total exposure for children ages 2–3 years and 6–12 years and adults, respectively. Also, two toxicological studies of PCB inhalation exposure were cited by Lehmann et al., and they reported notable health effects, identifying inhalation exposure to PCBs as a potential health hazard.

Conclusions

Studies from various countries showed that indoor air PCBs concentrations of industrial sites were tens of thousands of times higher than those in urban sites, so the pollution in industrial sites was significantly serious, especially in e-waste storage and recycling facilities. Therefore, the health of workers in industrial area needs to be paid special attention. Actually, the data on PCBs concentrations in the various countries are still not enough, especially in Africa.

It was found that the gas-phase PCBs accounted for most proportion of the total atmosphere PCBs in indoor air both in urban and industrial areas. Also, studies showed that the dominant congeners were lower molecular weight PCBs congeners, such as tetra-CBs and penta-CBs.

The estimation of human exposure to PCBs included dietary exposure and nondietary exposure. Dietary exposure is the dominant pathway and usually accounts for 25–65% of the total PCB exposure. For nondietary exposure, air inhalation accounts for a larger proportion than dust ingestion, sometimes even reaching 60–70%, and dust ingestion usually accounts for less than 20%.

Studies over the world weren’t conducted under exactly the same conditions. Thus, some results that came from active sampling and passive sampling might be different. Also, different instrument analytical methods would also influence the results. Thus, uniform sampling and analytical methods could contribute to a more reliable results and comparability.

Funding

The authors are grateful for the financial support provided by the National Natural Science Foundation of China (21207116) and the Scientific Research Foundation of Zhejiang Agriculture and Forestry University (2010FR090).

Additional information

Funding

The authors are grateful for the financial support provided by the National Natural Science Foundation of China (21207116) and the Scientific Research Foundation of Zhejiang Agriculture and Forestry University (2010FR090).

Notes on contributors

Qizhou Dai

Qizhou Dai is an associate professor of Zhejiang University and Technology.

Xia Min

Min Xia is a master of Zhejiang University and Technology.

Mili Weng

Mili Weng is an assistant professor of Zhejiang Agriculture and Forestry University and the program director for the National Natural Science Foundation of China (21207116) and the Scientific Research Foundation of Zhejiang Agriculture and Forestry University (2010FR090).