Fs and dl-PCBs in atmosphere of East Asia

Atmospheric Environment 180 (2018) 23–36

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Atmospheric Environment journal homepage: www.elsevier.com/locate/atmosenv

Review article

Review on occurrence and behavior of PCDD/Fs and dl-PCBs in atmosphere of East Asia

T

Minh Man Trinh, Moo Been Chang∗ Graduate Institute of Environmental Engineering, National Central University, Chungli 320, Taiwan

A R T I C L E I N F O

A B S T R A C T

Keywords: PCDD/Fs Dioxin-like PCBs Occurrence Ambient air East Asia Spatial/temporal trends Contents

This paper reviews the data from studies mainly published after 2000 to provide the current understanding of the physicochemical properties, atmospheric occurrence, gas/particle partitioning, fate and temporal trends in atmospheric matrix of PCDD/Fs and dl-PCBs of East Asia. Ambient PCDD/Fs and dl-PCBs concentrations in East Asia are found to be tens to hundreds times higher than that measured in Europe and North America. After strict regulations on PCDD/Fs and dl-PCBs emissions are enacted, the concentrations of these compounds decrease dramatically in Eastern Asian countries. In general, most of PCDD/Fs distribute in particle phase while dl-PCBs majorly exist in gas phase. Three main factors including physicochemical properties of the compounds, properties of particle and atmospheric condition affect the gas/particle partitioning of PCDD/Fs and dl-PCBs. The accuracy of absorption and adsorption models on predicting gas/particle partitioning of PCDD/Fs and dl-PCBs is evaluated. Gas-phase compounds are mostly removed from the atmosphere via reactions with OH radicals while those in particle phase are majorly removed by wet/dry deposition processes. The effects of removing processes and long-range transport on gas/particle partitioning are also discussed.

1. Introduction Global living standard has been increasingly enhanced after World War II due to the advancement of technology in general and chemical engineering in particular. However, a higher living standard also means higher demand in chemical products as well as higher chemical waste discharging. Among them, persistent organic pollutants (POPs) have caused great public concerns due to their unique physicochemical characteristics, including persistence in environment, ability to transport in a long distance, high toxicity to living creatures and accumulation in organic substances as defined by the Stockholm Convention. These properties make POPs ubiquitous in environment and they have been found in various environmental compartments not only in the area close to emission sources but also in remote areas. Among 12 initial POPs regulated by Stockholm Convention, polychlorinated dibenzo-p-dioxins (PCDDs) (Fig. 1a), polychlorinated dibenzo-furans (PCDFs) (Fig. 1b) and polychlorinated biphenyls (PCBs) (Fig. 2) are listed under Annex C (unintentional production) and have received great interests from researchers all over the world during the past 30 years because of their high toxicity and bioaccumulation compared to other POPs (Alcock and Jones, 1996; Jones and Sewart, 1997; Lohmann and Jones, 1998; Tuppurainen et al., 1998; McKay, 2002; Pereira, 2004; Tanabe and Minh, 2010; Dopico and Gomez,

∗

2015). Physicochemical characteristic of PCDD/Fs and dioxin-like PCBs (dl-PCBs) are summarized in Table 1 and Table 2, respectively. Generally, all PCDD/Fs are of low vapor pressures (Pv ranges from 2.57 × 10−5 to 6.61 × 10−8 Pa at 25 °C), low water solubilities (S ranges from 1.03 × 10−4 to 2.43 × 10−2 mg L−1 at 25 °C) and high octanol-water coefficients (Kow = 107.06 to 108.48) which suggest that they are more likely to distribute in suspended solids rather than in water (Lohmann and Jones, 1998). The octanol-air coefficient (Koa) ranges from 1010.05 for TCDD/F to nearly 1013 for OCDD/F and this is the key factor determining their gas/particle partitioning in ambient air (Lohmann and Jones, 1998; Harner et al., 2000). Similar to PCDD/Fs, all PCBs are of low water solubilities (2.7 × 10−3 to 6.3 × 10−5 μg L−1 at 25 °C), high octanol-water coefficients (Kow = 106.47 to 107.15) and high octanol-air coefficients (Koa = 109.62 to 1011.15). The physical and chemical properties of PCDD/F and PCBs make the degradation of those compounds a slow process in the environment. PCDD/Fs are very persistent and highly bioaccumulative in environment. The half livies of PCDD/Fs range from a few weeks in rodents to several years in human (Rose et al., 1976). Moreover, some studies indicate that the half life of PCDD/Fs is dose-dependent, and also the persistence increases with the increase of body fat (Emond et al., 2005). Therefore, some congeners take shorter time as 6 months to decompose while others may need even 2 years to do so (Schecter et al., 2006). According to the European

Corresponding author. E-mail address: [email protected] (M.B. Chang).

https://doi.org/10.1016/j.atmosenv.2018.02.037 Received 11 October 2017; Received in revised form 13 February 2018; Accepted 19 February 2018 Available online 23 February 2018 1352-2310/ © 2018 Elsevier Ltd. All rights reserved.

Atmospheric Environment 180 (2018) 23–36

M.M. Trinh, M.B. Chang

Fig. 1. Chemical structure of PCDDs (left) and PCDFs (right).

Table 2 Typical physicochemical characteristics of dl-PCBs (Mackay et al., 2006).

Fig. 2. Chemical structure of PCBs.

Table 1 Typical physicochemical characteristics of PCDD/Fs (Mackay et al., 2006). Congeners

2378-TCDD 12378-PeCDD 123478-HxCDD 123678-HxCDD 123789-HxCDD 1234678HpCDD OCDD 2378-TCDF 12378-PeCDF 23478-PeCDF 123478-HxCDF 123678-HxCDF 234678-HxCDF 123789-HxCDF 1234678HpCDF 1234789HpCDF OCDF a b

Pv (Pa at 25 °C)a

log Kowa

2.57 × 10−5 4.17 × 10−6 8.91 × 10−7 8.51 × 10−7 8.51 × 10−7 2.04 × 10−7

log Koa at 25 °Cb

S (mg L−1 at 25 °C)a

7.06 7.55 7.94 7.94 7.92 8.27

10.0 10.6 11.1 12.2 12.3 11.4

0.0158 4.00 × 10−3 1.15 × 10−3 1.10 × 10−3 1.10 × 10−3 3.15 × 10−4

6.61 × 10−8 8.91 × 10−5 1.70 × 10−5 1.15 × 10−5 3.24 × 10−6 3.09 × 10−6 2.45 × 10−6 2.09 × 10−6 9.33 × 10−7

8.48 6.13 6.47 6.56 6.92 6.93 7.01 7.07 7.37

13.0 10.0 11.4 11.5 11.9 12.0 12.1 12.2 12.2

1.03 × 10−4 0.0243 0.0679 0.00481 1.64 × 10−3 1.56 × 10−3 1.24 × 10−3 1.06 × 10−3 4.92 × 10−4

5.75 × 10−7

7.60

12.3

2.83 × 10−4

2.88 × 10−7

8.03

12.8

1.22 × 10−4

Homologue groups

Pva (Pa at 25 °C)

Log Kowb

Log Koa at 25 °Cc

Sd (mg L−1, 25 °C)

3,3′,4,4′-TeCB (77) 3,4,4′,5-TeCB (81) 2,3,3′,4,4′-PeCB (105) 2,3,4,4′,5-PeCB (114) 2,3′,4,4′,5-PeCB (118) 2′,3,4,4′,5-PeCB (123) 3,3′,4,4′,5-PeCB (126) 2,3,3′,4,4′,5-HxCB (156) 2,3,3′,4,4′,5′-HxCB (157) 2,3′,4,4′,5,5′-HxCB (167) 3,3′,4,4′,5,5′-HxCB (169) 2,3,3′,4,4′,5,5′-HpCB (189)

0.002 3 × 10−3 8.85 × 10−4 1.24 × 10−3 1.23 × 10−3 1.3 × 10−3 4.84 × 10−4 2.1 × 10−4

6.50 6.53 6.61 6.47 6.49 6.50 6.56 6.75

9.70 9.88 10.2 9.62 9.86 9.83 10.6 10.4

2.70 × 10−3 3.13 × 10−3 1.66 × 10−3 2.63 × 10−3 2.07 × 10−3 8.99 × 10−4 1.33 × 10−3 1.10 × 10−3

2.02 × 10−4

6.73

10.6

2.96 × 10−4

2.8 × 10−4

6.82

10.6

1.07 × 10−4

6.88 × 10−4

7.01

11.3

1.30 × 10−4

4.77 × 10−5

7.15

11.2

6.30 × 10−5

a b c d

Pv from Foreman and Bidleman (1985). Log Kow from Yeh and Hong (2002). Log Koa from Chen et al. (2002). S from Huang and Hong (2002).

bioaccumulation via food chains. The review conducted by Lohmann and Jones (1998) on the levels, behavior and processes affecting PCDD/ Fs in air and the deposition in Europe, North America and Japan indicated that PCDD/Fs concentrations in ambient air varied from 0.5 to 20 pg m−3. However, the atmospheric occurrence of PCDD/Fs has changed significantly during the last three decades and the global emission of PCDD/Fs has shifted to less developed countries in East Asia. In addition, it is confirmed that PCDD/Fs are more likely to distribute in particle phase, especially, higher chlorinated homologues and lower ambient temperatures give a higher particle bound fraction of total PCDD/Fs concentration (Lohmann and Jones, 1998). For loss process, PCDD/Fs in gas phase are mainly removed by OH· radicals in ambient air (Atkinson, 1987; Kwok et al., 1994) while those in particle phase are mostly removed via wet and dry depositions (Lohmann and Jones, 1998). To our knowledge, the occurrence and gas/particle partitioning of dl-PCBs have been conducted in some areas of East Asia (Cui et al., 2017; Devi et al., 2014; Ding et al., 2013; Liu et al., 2015; Tian et al., 2016; Zhang et al., 2017; Zhu et al., 2017; Li et al., 2010; Hu et al., 2014; Nie et al., 2014; Yang et al., 2017; Zhan et al., 2017; Li et al., 2011a, 2015), it is important to summarize and discuss the information of atmospheric dl-PCBs levels as well as their behavior and loss process in the ambient air of this region. Moreover, the trend of long-term atmospheric concentration, the characteristics of gas-particle partitioning and wet/dry deposition processes and the evidence of longrange transport of PCDD/Fs and dl-PCBs have become the focus of many studies in recent years. Relevant studies found that PCDD/Fs and dl-PCBs levels in ambient air of developed countries have been decreasing overtime especially after strict regulations on the emissions

Pv, pKow, S from Wang and Wong (2002, 2003). Log Koa from Harner et al. (2000)

Environmental Agency, exposure to PCDD/Fs may lead to heart problems, cause cancer and damage to unborn children even with the exposures to low concentration (Agency, 2016). The major pathway for human intake is from food consumption, especially via fatty dietary (Llobet et al., 2008). A review conducted by Schecter et al. (2006) consolidates critical negative effects of PCDD/Fs to human health such as cancer, immune deficiency, endocrine disruption, skin rashes and liver damage. Exposure to dl-PCBs with high doses reveals similar effects as PCDD/Fs on organisms (Pereira, 2004). The levels of PCDD/Fs and dl-PCBs in food, animal, fish and human are declining in developed countries while rising in less developed ones (Fries, 1995; Domingo and Bocio, 2007; Williams and Cseh, 2007; Srogi, 2008; Consonni et al., 2012; Malisch and Kotz, 2014). Ambient air is an important transport pathway of PCDD/Fs and dl-PCBs from the emission sources to other environmental matrixes prior to the 24

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M.M. Trinh, M.B. Chang

significant amount of PCDD/Fs is released by mobile sources. Among all kind of vehicles, motorcycles are regarded with the highest emission factor of 94.2 pg I-TEQ km−1 compared to those of unleaded gasoline vehicles (1.5–2.6 pg I-TEQ km−1) and diesel-fueled vehicles (2.4 pg ITEQ km−1) in Taiwan (Chuang et al., 2010a, 2010b). Although there is no consolidated study characterizing the emission sources of dLl-PCBs in the ambient air, it is believed that dl-PCBs emission mechanisms are similar to those of other PCBs. Among 209 congeners of PCBs, 130 congeners were produced massively and used widely as a technical mixture during 1930s-1980s. Highly chlorinated congeners were applied as cooling/isolation fluid in transformers, hydraulic system, gas turbine and vacuum pump; or fired retardants in electric devices and as plasticizers in adhesives, textiles, surface coatings and sealants. Low chlorinated congeners, on the other hand, were used as basis for resinous products and inks (Pereira, 2004). In developed countries, production and consumption of PCBs has been terminated but serious concerns remain on the spreading of these pollutants in the environment due to their re-volatilization from sites where they have been stored or disposed of, and incineration of PCBs-containing materials. In addition, formation of PCBs during thermal processes has been proven to have similar formation mechanisms as PCDD/Fs by many works (Abad et al., 2006; Weber et al., 2001; Brubaker and Hites, 1998). A recent study conducted by Davies and Delistraty (2016) evaluated the source of PCBs in atmospheric environment in Washington State, USA and found that the largest source is attributed to the evaporation from lamp ballasts (400–1500 kg/year), followed by unintentional industrial process (900 kg/year), sealants and capacitors (approximately 400 kg/year).

from different kinds of incinerators are adopted (Alcock and Jones, 1996; Kulkarni et al., 2008). Based on a young, high-skilled population and an exploding economy, east Asian countries are becoming the largest production factory for the world. As a consequence, developing countries in Asia are considered as the largest POPs (especially PCDD/Fs and dl-PCBs) producers due to massive industrialization and loose regulation in this region (Hogarh et al., 2012). Therefore, it is important to understand the occurrence of atmospheric PCDD/Fs and dl-PCBs over time for developing countries in East Asia. In this region, to our knowledge, studies related to these pollutants in ambient air have already been reported in Japan, South Korea, China, Taiwan, Vietnam and Thailand. Hence, we mainly focus on the studies working on atmospheric PCDD/ Fs and dl-PCBs conducted after 2000 from these countries in this review. Furthermore, the removal processes in atmosphere and gas/particle partitioning modeling are discussed to understand the behavior of PCDD/Fs and dl-PCBs after being emitted from sources and before the accumulation in other matrixes. 2. Sources of PCDD/Fs and dl-PCBs PCDD/F and dl-PCB concentrations and congener distribution in the environment are strongly affected by the distance to nearby emission sources and the characteristics of those pollutants. Major PCDD/Fs emission sources include combustion activity (fossils fuel combustion, municipal, hazardous and medical waste incineration, open burning, vehicle), metal smelting, chemical producing and re-emission from water body, sediment and soil (McKay, 2002). In combustion processes, PCDD/Fs can be formed via three main mechanisms. Firstly, PCDD/Fs already exist in feeding material and survive through combustion process. Secondly, precursors including polyhalogenated phenols, orthohalogenated phenols and chlorinated aromatic compounds are thermally destroyed and rearranged to form PCDD/Fs. Thirdly, PCDD/Fs are synthesized via de novo process after combustion. De novo process is defined as a pathway by which PCDD/Fs are formed via heterogeneous reaction on fly ash involving carbon, hydrogen, oxygen and chorine with participation of metal catalyst. Various kinds of waste incinerators, especially municipal waste incinerators (MSWIs) are identified as major PCDD/F contributors since incineration has become the mainstream technology for waste treatment in many countries (Tuppurainen et al., 1998; McKay, 2002; Buekens and Huang, 1998). In China, the average emission factors of PCDD/Fs and dl-PCBs from MSWIs to the atmosphere are 1.728 μg TEQ ton−1 MSW (Ni et al., 2009) and 4.52 ng-TEQ ton−1 MSW (Li et al., 2016), respectively, and the amount of PCDD/Fs emitted from MSWIs to the atmosphere was 19.64 g TEQ year−1 in 2006 (Ni et al., 2009). In addition to MSWIs, metal smelting processes including iron ore sintering, electric arc furnace and secondary nonferrous (zinc, lead, copper and aluminum) smelting processes are also important sources of PCDD/Fs emission in China (Liu et al., 2013a). The national implementation plan of China for the Stockholm Convention on persistent organic pollutants indicate that total PCDD/Fs emissions from all sources of China reached 10.2 kg-TEQ in 2004, and among them 45.6% was contributed by ferrous and nonferrous metal production (China-NIP, 2007). Quaβ et al. (2004) indicate that iron and steel industry releases more PCDD/Fs than other nonferrous metal metallurgy. Recent study indicated that the PCDD/Fs emission factors of steel, zinc, copper, aluminum and lead industries are 0.8 ng TEQ kg−1 (Zou et al., 2012), 50.6 ng I-TEQ kg−1, 14.8 ng I-TEQ kg−1, 2.65 ng I-TEQ kg−1 and 0.612 ng I-TEQ kg−1, respectively (Liu et al., 2013a). Emission factors of dl-PCBs from steel production in China are 12.4 ng TEQ ton−1 for iron ore sintering and 73.1 ng TEQ ton−1 for electric arc furnace (Liu et al., 2013b). On the other hand, the emission factors from zinc, copper, aluminum and lead production are 1464 ng TEQ ton−1 (Ba et al., 2009a), 98.1 ng TEQ ton−1 (Ba et al., 2009b), 194 ng TEQ ton−1 (Ba et al., 2009a) and 3.9 ng TEQ ton−1 (Ba et al., 2009b), respectively. In addition to stationary emission sources, a

3. Occurrence and behavior of atmospheric PCDD/Fs and dl-PCBs in East Asia 3.1. Occurrence of atmospheric PCDD/Fs and dl-PCBs in East Asia Tables 3 and 4 summarize the typical PCDD/F and dl-PCBs concentrations in ambient air in East Asia collected by high volume active samplers (HVAS). In general, the total concentration of these pollutants in industrial areas ranged from 885 fg m−3 (49.8 fg WHO-TEQ m−3) in an industrial site in Shanghai to 164 pg m−3 (3030 fg I-TEQ m−3) near an MSWI in Hangzhou, China. For urban areas, both the lowest ambient air concentration are measured at Hongkong (8 fg I-TEQ m−3) and highest concentrations are measured at Guangzhou, China, varying from 3.08 to 16.1 pg m−3 (139–1170 fg I-TEQ m−3). In rural areas, PCDD/F concentrations ranging from 1.01 fg I-TEQ m−3 to 630 fg ITEQ m−3 were measured in southern Taiwan. On the other hand, the lowest concentration was measured in the remote area of Lanyu island, Taiwan, with the value of 1.24 fg I-TEQ m−3. Ambient concentration of PCDD/Fs in East Asia is found to be higher than those reported for Europe and North America by Lohmann and Jones (1998), which ranged from 20 to 50 fg TEQ m−3 in rural areas and from 100 to 400 fg TEQ m−3 in urban/industrial areas. In addition, the congeners profiles in urban and rural sites are consistent with the higher contribution of highly chlorinated congeners such as OCDD/F and 1,2,3,4,6,7,8HpCDD/F (together contributing over 70%). For TEQ concentration, 2,3,4,7,8-PeCDF and 1,2,3,7,8-PeCDD are the dominant congeners in all samples regardless the sampling site, together, they contribute over 50% of the total TEQ. This trend complies with those measured in Europe and North America as reported by Lohmann and Jones (1998). In term of dl-PCBs, the highest concentration (144 fg TEQ m−3) was recorded in an industrial area in Tianjin, China and the lowest value was 2.23–4.49 fg TEQ m−3, being sampled in the urban site of Taipei city, Taiwan. Dominant congeners for dl-PCBs are PCB-118, PCB-105 and PCB-77 which contribute around 50%, 20% and 15%, respectively, to the total dl-PCBs concentration. The most dominant contributors to the total TEQ of dl-PCBs are PCB-126 (92%) and PCB 169 (around 5%). The spatial trends of PCDD/Fs concentration measured in east Asian 25

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M.M. Trinh, M.B. Chang

Table 3 Levels of PCDD/Fs in ambient air of industrial, urban, rural, and remote sites of East Asia (samples collected by HVAS).

Industrial site

Country

Location

Mass Concentration

TEQ concentration

PCDD/PCDF ratio

Reference

China

Vicinity of an MSWI in Hangzhou Vicinity of the abandoned PCP plant, Tianjin Vicinity of an MSWI before and after demolishment

3.96–164 pg m−3 29.5 pg m−3

59 to 3030 fg I-TEQ m−3 104 fg I-TEQ m−3

0.4 (0.2–0.7) 2.57

(Xu et al., 2009) (Li et al., 2012)

Before demolishment: 5.67 pg m−3 After demolishment: 3.63 pg m−3

Before demolishment: 514 fg I-TEQ m−3 After demolishment: 345 fg I-TEQ m−3

(Zhang et al., 2014)

Industrial site in Shanghai

Summer: 0.88 pg m−3 Winter: 2.30 pg m−3

Vicinity of an MSWI in central Taiwan Vicinity of an MSWI in northern Taiwan Seoul City of Nagoya

3.50 and 5.70 pg m−3 N/A

Summer: 49.8 fg WHO-TEQ m−3 Winter: 128 fg WHO-TEQ m−3 208 and 210 fg I-TEQ m−3 91 to 271 fg I-TEQ m−3

Before demolishment: 0.61 (0.35–0.94) After demolishment: 0.48 (0.40–0.54) N/A

N/A N/A

(Chao et al., 2004) (Chang et al., 2004)

617 fg I-TEQ m−3 N/A

N/A N/A

18 to 644 fg I-TEQ m−3 148.4 to 497.1 fg I-TEQ m−3 186 to 1170 fg I-TEQ m−3 240 fg I-TEQ m−3 18 to 430 fg I-TEQ m−3 32.2 to 52.5 fg I-TEQ Spring: 229 to 393 fg I-TEQ Nm−3 Fall: 362 to 720 fg I-TEQ Nm−3 1.01 to 27.4 fg I-TEQ m−3 Spring: 270 fg I-TEQ m−3, Summer: 160 fg I-TEQ m−3, Fall: 240 fg I-TEQ m−3, Winter: 630 fg I-TEQ m−3 23.4 to 146 fg I-TEQ m−3 11.1 to 59.5 fg I-TEQ m−3 139 fg I-TEQ m−3 7.64 to 12.1 fg I-TEQ m−3 1.52 to 10.8 fg I-TEQ m−3 1.24 to 10.8 fg I-TEQ m−3 0.71 to 3.41 fg I-TEQ m−3

0.37 0.44 0.98 0.62 N/A 0.66 N/A

(Shin et al., 2011) (Kadowaki and Naitoh, 2005) (Li et al., 2008a) (Li et al., 2008b) (Deng et al., 2011) (Li et al., 2014a) (Sin et al., 2002) (Chi et al., 2008) (Mi et al., 2012)

Taiwan

Urban site

South Korea Japan China

Taiwan

Rural site

Remote site

Beijing Shanghai Guangzhou Hongkong Taipei Tainan

Taiwan

Southern Taiwan Rice field in southern Taiwan

Vietnam

Danang Sonla Hochiminh city Chiangmai Dongsha island Pengjiayu and Lanyu island Lulin mountain in central Taiwan

Thailand Taiwan

N/A 13.5–32.3 pg m−3 (mean: 18.7 pg m−3) 0.275–10.78 pg m−3 0.699–8.03 pg m−3 3.08–16.1 pg m−3 5.30 pg m−3 N/A N/A N/A

N/A Spring: 0.342 pg m−3, Summer: 0.221 pg m−3, Fall: 0.675 pg m−3, Winter: 0.741 pg m−3 N/A N/A N/A N/A N/A N/A N/A

(0.21–0.91) (0.15–1.17) (0.23–3.29) (0.22–1.23)

(Die et al., 2015)

N/A N/A

(Thuan et al., 2013) (Shih et al., 2006)

N/A N/A N/A N/A N/A N/A N/A

(Thuan et al., 2013) (Chi et al., 2016) (Ngo et al., 2017) (Chi et al., 2013a) (Thuan et al., 2013) (Chi et al., 2013b) (Chi et al., 2010)

*N/A: Not available.

Although HVAS provides accurate data, it is considered as costly for both capital and operating expenses. Despite relatively high uncertainties of sampling rate, passive sampler (PS) which is cheap and simple to operate has become a suitable monitoring tool for developing states (Klanova and Harner, 2013). The data of PCDD/Fs and dl-PCBs collected by passive sampler are represent in Table 5. Most studies applied the sampling rate of PS equal to 3.5 m3 per day (see Table 5) to

cities are presented in Fig. 3. In general, significantly higher levels of PCDD/Fs are observed in urban and industrial sites in China compared to those measured in South Korea and Taiwan. The lowest PCDD/Fs concentration clearly occur in rural site of South East Asia with a higher level in Vietnam compared to those in Thailand. Higher dl-PCBs concentrations are also observed in China compared to South Korea, Taiwan and Japan as presented in Fig. 5.

Table 4 Levels of dl- PCBs in ambient air of industrial, urban, and rural sites of East Asia (samples collected by HVAS).

Industrial site

Urban site

Country

Location

Mass Concentration

WHO-TEQ concentration

Reference

South Korea China

Gyeonggi-do Industrial site in Shanghai

Gyeonggi-do Yokohama Taipei city Hongkong

3.52–4.14 pg m−3 N/A N/A 2.96 pg m−3

31.1 to 49.2 fg TEQ m−3 Summer: 7.12 fg TEQ m−3 Winter: 16.5 fg TEQ m−3 5 fg TEQ m−3 (Excluding PCB-126 and-169) 0.04 to 2.3 fg TEQ m−3 in Linfen 0.26 to 0.4 fg TEQ m−3 in Datong 8 to 15 fg TEQ m−3 2 to 14 fg TEQ m−3 2.23 to 4.49 fg TEQ m−3 70–110 fg TEQ m−3

(Kim et al., 2011) (Die et al., 2015)

Iron and steel industrial bases in Liaoning, China Industrial base in Shanxi, China

6.66–7.73 pg m−3 Summer: 2.96 pg m−3 Winter: 4.76 pg m−3 4.14 pg m−3 N/A

Tianjin Beijing

N/A 1.68 pg m−3

144 fg TEQ m−3 8.78 fg TEQ m−3

(Li et al., 2015) (Zhu et al., 2017)

South Korea Japan Taiwan China

*N/A: Not available.

26

(Li et al., 2010) (Nie et al., 2014)

(Kim et al., 2011) (Kim and Masunaga, 2005) (Chi et al., 2008) (Choi et al., 2008a)

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Fig. 3. Levels of PCDD/Fs (fg TEQ m−3) in ambient air of different locations across East Asia. Data are from (Xu et al., 2009), (Li et al., 2012), (Hu et al., 2014), (Die et al., 2015), (Chao et al., 2004), (Chang et al., 2004), (Shin et al., 2011), (Li et al., 2008a), (Li et al., 2008b), (Deng et al., 2011), (Li et al., 2014a), (Sin et al., 2002), (Chi et al., 2008), (Thuan et al., 2013), (Chi et al., 2013b), (Ngo et al., 2017) and (Chi et al., 2010).

Table 5 Levels of PCDD/Fs and dl-PCBs in ambient air of industrial, urban, rural, and remote sites of East Asia (samples collected by PS).

Industrial site

Country

Location

Sampling rate

Mass Concentration

TEQ concentration

Reference

China

A steel industrial complex in northeast China

3.50 m3 day−1

3.50 m3 day−1

Vicinity of secondary copper and aluminum metallurgical facilities A steel complex in Gyeongsangbuk-do Gyeonggi Province

3.50 m3 day−1

South Korea

Gyeonggi Province

N/A

China

Shanghai

N/A

PCDD/Fs: Average of 74.0 fg TEQ m−3 dl-PCBs: Average of 23.0 fg TEQ m−3 PCDD/Fs: 54.3 fg TEQ m−3 in summer, 51.2 fg TEQ m-in winter dl-PCBs: 9.65 fg TEQ m−3 in summer, 6.95 fg TEQ m−3 in winter PCDD/Fs + dl-PCBs: 4450 fg TEQ m−3 N/A PCDD/Fs: 625 fg TEQ day−1 dl-PCBs: 62 fg TEQ day−1 PCDD/Fs: 209 fg TEQ day−1 dl-PCBs: 15 fg TEQ day−1 PCDD/Fs: 63.4 fg TEQ m−3 in summer, 83.4 fg TEQ m-in winter dl-PCBs: 9.46 fg TEQ m−3 in summer, 4.57 fg TEQ m-in winter PCDD/Fs: 122 fg TEQ day−1 dl-PCBs: 8 fg TEQ day−1

(Li et al., 2011b)

Vicinity of an MSWI

PCDD/Fs: 0.74 pg m−3 in summer, 1.97 pg m−3 in winter dl-PCBs: Average of 4.56 pg m−3 PCDD/Fs: 0.85 pg m−3 in summer, 1.46 pg m−3 in winter dl-PCBs: 2.26 pg m−3 in summer, 1.42 pg m−3 in winter PCDD/Fs: 41.8 pg m−3 dl-PCBs: 1560 pg m−3 dl-PCB: 26.2 pg m−3 PCDD/Fs: 8.37 pg day−1 dl-PCBs: 32.9 pg day−1 PCDD/Fs: 2.81 pg day−1 dl-PCBs: 17.8 pg day−1 PCDD/Fs: 0.63 pg m−3 in summer, 0.91 pg m−3 in winter dl-PCBs: 4.16 pg m−3 in summer, 2.56 pg m−3 in winter PCDD/Fs: 1.43 pg day−1 dl-PCBs: 18.8 pg day−1

South Korea

Urban site

Rural

South Korea

Gyeonggi Province

3

3.00 m day N/A

N/A

*N/A: Not available.

27

−1

(Gao et al., 2014)

(Hu et al., 2014) (Choi et al., 2008b) (Yoonki et al., 2014) (Yoonki et al., 2014) (Tian et al., 2015)

(Tian et al., 2016)

(Yoonki et al., 2014)

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et al., 2008b) and PCDDs were more likely to distribute in particle phase than PCDFs due to lower vapor pressures. In addition to the physicochemical properties of PCDD/Fs and dlPCBs compounds, properties of particles itself can affect gas/particle portioning. Whitby (1967) classified the aerosol into three size groups, i.e., nuclei mode (or ultrafine) with d < 0.1 μm; accumulation mode (or mid-sized) 0.1 < d < 2.0 μm; and coarse size modes: d > 2.0 μm and Bidleman (1988) found that the concentrations of SVOCs increased as the particle size decreased. Kaupp et al. (1994) reported that approximately 90% of the PCDD/Fs were found on particles with d < 1.35 μm while Zhang et al. (2015) reported that the highest PCDD/Fs concentrations were found in the d < 1.0 μm particles and more than 80% of the PCDD/Fs were found in the d < 2.5 μm particles. Kaupp and McLachlan (1999) indicate that coarse particles contribute most to dry deposition of PCDD/Fs, in contrast, the wet deposition of the PCDD/Fs is dominated by fine particles. Particle size also affects the congeners distribution of PCDD/Fs, i.e., 92% of the Hepta and OCDD/Fs are associated with particles with d < 2.9 μm (Kaupp and McLachlan, 2000) and the fraction of less chlorinated PCDD/F in the homologue groups increases with increasing particle size (Kurokawa et al., 1998; Oh et al., 2002) especially when the sampling site is located far away from the emission source (Chao et al., 2003). Accumulation of higher chlorinated PCDD/Fs in fine particle is confirmed by Zhang et al. (2016) as ambient air sampling in suburban Beijing is conducted. Mono-to tri-CDD/Fs mainly exist in the gas phase, while tetra-to octaCDD/Fs mainly distribute in the particulate phase and the majority of mono-to tri-CDD/Fs DD/Fs (70%) is reported to present in particles with d > 1.0 μm while the opposite trend is observed for tetra-to octaCDD/Fs for which 22% are found in d > 1.0 μm particles. Lee et al. (1996) investigated the particle size distributions and found the highest peak of total PCB mass was localized between 5.6 and 10.0 μm which belong to the coarse particle mode. And the second highest peak was found in the particle size ranging from 0.31 to 0.52 μm which belongs to the accumulation mode. Degrendele et al. (2014) studied the distribution of dl-PCBs in different particle sizes and found that 8.86% of the dl-PCBs existed in coarse mode in rural site during autumn, winter and spring time but its distribution increased to 53.8% in summer. The reason could be the difference in particle size distribution between summer and other seasons, i.e., the higher coarse fractions observed in summer were attributed to additional vehicle-related sources, in particular, particle resuspension from road surfaces. The fractions of PCDD/Fs and dl-PCBs in fine particles from emission sources reveals a similar trend to those collected in ambient air. Han et al. (2017) collected the samples from the flue gas of two MSWIs (> 600 tons of solid waste per day) and found that distributions of PCDD/Fs and dl-PCBs in particulate matter with diameter < 10 μm accounted for more than 81% and 79% of the total particulate matter. Moreover, more than 54%

calculate the ambient concentration of these pollutants. However, the uptake rate of each PCDD/Fs and dl-PCBs congeners is not similar due to the differences in their concentration and diffusion coefficient. Heo and Lee (2014) conducted a sampling campaign using both HVAS and PS in Gyeonggi-do, South Korea to find out the accurate uptake rates of each congener. As a result, without considering the effect of seasonal variation, the uptake rate of PCDD/Fs range from 1.2 to 1.6 m3 day−1, and dl-PCBs range from 1.9 to 2.1 m3 day−1, depending on the sampling sites. Greater uncertainty is expected as PS is operated in coastal/ mountain sites or in windy condition (Klanova and Harner, 2013), and a more reliable sampling protocol is needed for these circumstances. 3.2. Gas/particle partitioning of PCDD/Fs and dl-PCBs in atmosphere PCDD/Fs and dl-PCBs fate and behavior in environment are influenced by their gas/particle partitioning. For instance, degradation mechanisms including photodegradation and OH radical reactions in gas phase are more effective than that in particle phase (Brubaker and Hites, 1997) while particle-phase PCDD/Fs and dl-PCBs are affected by long range atmospheric transport and deposition mechanisms more than gas phase (Lohmann and Jones, 1998; Welsch-Pausch and McLachlan, 1998). Saral et al. (2015) suggest that the factors affecting gas/particle partitioning of PCDD/Fs and dl-PCBs can be divided into 3 groups: (1) physicochemical properties of the compounds such as molecular weights, vapor pressures, octanol-air coefficients, subcooled liquid vapor pressures and Henry's constants, (2) properties of particles including size distribution, particles concentration, organic and elemental carbon contents and (3) atmospheric conditions such as ambient temperature, rainfall and relative humidity. PCDD/Fs in atmosphere are mainly distributed in particle phase while dl-PCBs are mainly distributed in gas phase (Lohmann and Jones, 1998; Yeo et al., 2003), and this aspect has been extensively studied in Asian cities as presented in Fig. 4. Chi et al. (2008) conducted a sampling campaign to evaluate the gas/particle partitioning of PCDD/Fs in the rural northern coast of Taiwan and urban site of Taipei city, and the results indicate that the average PCDD/Fs in particle phase accounted for 65% and 77%, respectively. Most studies indicate that highly chlorinated congeners with low vapor pressures (Pv) are more likely to distribute in particle phase and the fraction of PCDD/Fs in particle phase increases with the increase of congeners’ molecular weight. The gas/particle partitioning of PCDD/Fs in the vicinity of two MSWIs situated in southern Taiwan were investigated seasonally (Huang et al., 2011a, 2011b), and the results indicate that the fraction of particles-phase TeCDD/Fs is 0.01 and increases to 0.9960 for OCDD/F. In addition, a monitoring campaign carried out in Beijing, China indicated the average contribution of particle phase to the total PCDD/Fs concentration reached 94.4% (Li

Fig. 4. Gas/particle distribution of PCDD/Fs in different locations of East Asia. Data are from (Chi et al., 2008), (Qin et al., 2012), (Li et al., 2008b), (Li et al., 2014b).

28

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Fig. 5. Levels of dl-PCBs in ambient air of different locations of East Asia. Data are from (Kim et al., 2011), (Kim and Masunaga, 2005), (Chi et al., 2008), (Choi et al., 2008a), (Li et al., 2010), (Zhu et al., 2017), and (Li et al., 2015).

where Kp is the particle partitioning coefficient, Cp and Cg are the particulate and gaseous PCDD/Fs or dl-PCBs concentrations (ng m−3), respectively, and PM is the particulate matter concentration (μg m−3). P0L represents the subcooled liquid vapor pressure of each PCDD/Fs and dl-PCBs congener at a specific temperature. The mr value (the slope) is supposed to indicate whether adsorption or absorption process dominates the gas/particle partitioning of each congeners while intercept br is affected by the particle properties. At equilibrium, mr should be close to −1 or 1 if either adsorption or absorption is dominant (Pankow, 1994). The parameter φ is the fraction of particle-phase SVOCs, c is a constant depending on the molecular weight and the heat of condensation (Pa cm), θ is the particle surface area (cm2 cm−3), and P0L is the subcooled liquid vapor pressure (Pa). An empirical value of c = 0.172 (Pa m) was assumed and would not vary among different compounds while θ = 1.1 × 10−3 cm2 cm−3 for urban, 1.5 × 10−3 cm2 cm−3 for rural, and 4.2 × 10−3 cm2 cm−3 for clean background air were adopted by previous studies (Falconer and Harner, 2000; Cotham and Bidleman, 1992; Pankow et al., 1993). In addition to Junge–Pankow model, Harner and Bidleman (1998) proposed an absorption model that only requires the values of Koa (octanol-air coefficient) and fom (fraction of organic matter on the particle) to predict the gas/particle partitioning:

of PCDD/Fs and 49% of dl-PCBs distributed in the particulate matter with diameter the < 2.5 μm. In terms of atmospheric condition, the fact that higher fraction of PCDD/Fs being distributed in particle phase than those in gas phase during winter time indicates significant effect of temperature on gas/ particle partitioning. By comparing the gas/particle partitioning between urban site in northern and southern Taiwan, Chi et al. (2013b) reported that PCDD/F in the particle-phase in northern Taiwan was higher than that in southern Taiwan since the higher ambient temperature in the southern areas compare to those in higher latitude areas. As the results, the lower fraction of PCDD/F congeners being adsorbed onto particles in southern Taiwan. In China, a study conducted in Dalian via a high volume sampler found that the PCDD/Fs appeared to be mainly present in the particle phase, with an average contribution of 97.1% to the total concentration during winter time (Qin et al., 2012). The lower fractions of particle phase were found in summer, spring and autumn, being 38.7%, 69.3% and 78.3%, respectively, suggesting that ambient temperature has a great impact on gas/ particle partitioning for PCDD/Fs. In terms of dl-PCBs, previous studies found that almost all dl-PCB congeners were predominant in gas phase. Chi et al. (2008) report that dl-PCBs in gas phase contributes 90% to the total dl-PCBs concentration in northern Taiwan and Yeo et al. (2003) found the similar contribution of gas phase in South Korea. Similar to PCDD/Fs, the gas/particle partitioning of dl-PCBs is also correlated with molecular weight. The monitoring campaign conducted in Yokohama, Japan indicated the contribution of particle-phase dl-PCBs increased with the increasing number of substituted chlorines which are of lower vapor pressure, and with the decrease of temperature (Kim and Masunaga, 2005). Two mechanisms including adsorption onto particle surface and absorption into the organic matter on particle surface have been applied to predict the gas/particle partitioning of PCDD/Fs and dl-PCBs. The Junge-Pankow adsorption and the Harner-Bidleman absorption models have been applied to explain the fate of these compounds. For adsorption process, Junge (1977) and Pankow (1987) developed a linear Langmuir isotherm for predicting gas–particle partitioning of atmospheric semi-volatile organic compounds (SVOCs) as:

log Kp = log Koa + log fom - 11.91 Fraction of SVOCs on the particulate can be derived from the equation: φ = Kp (PM) / (1 + Kp (PM) These models were applied in some studies conducted in Asia to predict the particle-phase fraction of PCDD/Fs and dl-PCBs. In Nagoya, Japan, a PCDD/Fs monitoring campaign was conducted and the data on gas phase and particle phase were applied to evaluate the Junge–Pankow adsorption and the Harner–Bidleman absorption models (Kadowaki and Naitoh, 2005). As a result, good agreement was obtained between the measured and partition coefficient (log Kp) values predicted by the Harner–Bidleman absorption models for both PCDDs and PCDFs, indicating that the Harner–Bidleman absorption models is more suitable to predict the gas/particle partitioning of PCDD/Fs than the Junge–Pankow adsorption model. Lee et al. (2007) applied both partitioning models to analyze the gas/particle partitioning during Asian dust storm event (ADS) in South Korea. The P0L and Koa, respectively, are used for the description of the equilibrium conditions. The slopes of plots (mr) are close to either −1 or 1 in adsorption and

Kp = (Cp/PM)/Cg logKp = mr logP0L + br Fraction of SVOCs on the particulate can be derived from the equation: φ = cθ/ (P0L + cθ) 29

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2,3,7,8-TCDD and 2,3,7,8-TCDF with OH radicals in the presence of O2 and NO/H2O. The results show that the rate constant of 2,3,7,8-TCDD reacting with OH radicals ranges from 1.26 × 10−21 to 2.83 × 10−20 (cm3 molecule− 1 s− 1) at 25 °C while that of TCDF ranges from 2.17 × 10− 24 to 5.72 × 10−23 (cm3 molecule− 1 s− 1) at 25 °C. The rate constants of gas-phase PCBs reacting with OH radicals have been widely studied (Brubaker and Hites, 1998; Atkinson, 1996; Kwok et al., 1995). The research conducted by Atkinson (1996) measured the second-order rate constant (kOH) for PCBs reacting with the OH˙ radical and the value ranges from 5.0 × 10−12 cm3 molecule−1 s−1 for 3chlorobiphenyl to 0.4 × 10 −12 cm3 molecule−1 s−1 for 2,2′,3,5’,6pentachlorobiphenyl. For a steady-state OH radical concentration of 3 × 106 molecules cm−3, these rate constants correspond to the halflives of 0.5–7 days for PCBs containing five or fewer chlorines. Totten et al. (2002) conducted a field study to examine the reaction of PCBs with OH radicals and calculate the rate constants for such reactions. The result obtained is higher than that reported by Atkinson (1996), but generally, it well agrees with the trend of lower reaction rate constant for higher molecular weight congeners PCBs with the OH radicals. For example, the rate constant decreases from 11 × 10−12 cm3 molecule−1 s−1 for dichlorobiphenyl to 2.8 × 10−12 cm3 molecule−1 s−1 for pentachlorobiphenyl compared to 2.0 × 10−12 cm3 molecule−1 s−1 to 0.6 × 10−12 cm3 molecule−1 s−1 for the same congener measured by Atkinson (1996). Another field study has been conducted by Mandalakis et al. (2003) in mid-latitude area of eastern Mediterranean to find out the rate constant with the diurnal variation of OH radicals concentration. The rate constant in this study decreases from 1.4 to 1.9 × 10−12 cm3 molecule−1 s−1 for dichlorobiphenyl to 0.6–1.0 × 10−12 cm3 molecule−1 s−1 for pentachlorobiphenyl. Moreover, higher levels of OH radicals are expected in low latitude areas, hence, shorter lifetimes of PCBs are expected at low latitudes than polar regions. Tropospheric lifetimes of PCDD/Fs and PCBs have been estimated as presented in Table 6. In general, tropospheric lifetimes of PCDDs range from 1 to 230 days which are significantly lower than those of PCDFs (2.9–580 days) and PCBs (2–90 days).

absorption models when gas/particle partitioning reaches equilibrium conditions. Li et al. (2008b) found a linear correlation between the regressions of Kp versus P0L with the slope values < −1 when conducted a study in Beijing, China, suggesting that PCDD/Fs do not approach equilibrium between the particle and gas phases. Die et al. (2015) applied Junge–Pankow adsorption and Harner–Bidleman absorption models to analyze dl-PCBs and found that gas/ particle partitioning of dl-PCBs reached equilibrium more easily in winter than in summer. On the other hand, in contrast to those of PCDD/Fs, Junge-Pankow model fitted the dl-PCB data better than Harner–Bidleman absorption model, indicating that adsorption mechanism plays a major role in the partitioning process of dl-PCBs. Moreover, Kim et al. (2011) reported that the slopes obtained from regressing log Kp versus log P0L (log Kp versus log Koa) were smaller than −1 (or 1) and both models tended to overestimate the φ and Kp values of the dl-PCB congeners compared to the measured value in urban area of Gyeonggi-do, South Korea. 3.3. Removal processes in atmosphere Depending on the phase of PCDD/Fs and dl-PCBs in the atmosphere, the removing processes involved may include decomposition, wet/dry deposition and long-range transport. 3.3.1. Reaction with OH radical Atkinson (1996) proposed the tropospheric removal or transformation of gas-phase PCDD/Fs may include wet and dry deposition, photolysis, and reaction with OH, HO2, NO3 radicals and O3. While the reaction of PCDD/Fs with HO2, NO3 radicals and O3 are considered negligible or slow, some researchers found that the reaction with OH radicals in atmospheric matrix is the dominant removal process for most organic compounds including PCDD/Fs (Zhang et al., 2011, 2012; Lee et al., 2004a) and PCBs (Mandalakis et al., 2003; Anderson and Hites, 1996; Totten et al., 2002; Kwok et al., 1995). The rate constants of PCBs and PCDD/Fs in reacting with OH radicals have been reviewed by Atkinson (1996) and they ranged from 0.4 to 8.32 × 1012 cm−3 molecules −1 s−1 for pentachlorobiphenyl to biphenyl, respectively, and 14.8 × 1012 cm−3 molecules −1 s−1 for dibenzo-p-dioxin (DD) to 3.9 × 1012 cm−3 molecules −1 s−1 for dibenzofuran (DF). However, the rate constants for higher chlorinated dibenzo-p-dioxins and dibenzofurans are not available so far. Experiment on the behaviors of PCDD/Fs is difficult to conduct even in a control laboratory due to the difficulties of maintaining PCDD/Fs in gas phase, the sorptive loss processes and the difficulties in sampling/ analysis and their products (Atkinson, 1996). Therefore, most research on this topic applied theoretical method to calculate the reaction rate constant of PCDD/Fs with OH radicals. Lee et al. (2004a) applied the density functional theory (DFT) to calculate the reaction rate constant of dibenzo-p-dioxin (DD), 1,4,6,9-tetrachlorodibenzop-dioxin (1,4,6,9TCDD), 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD), and octachlorodibenzo-p-dioxin (OCDD) with OH radical to find out the favorable reaction pathways and reaction sites in the gas phase. Two pathways including “addition” and “substitution” are compared. The “addition” pathway means OH radicals replaces a hydrogen atom on benzene ring while “substitution” indicates that OH radicals replace a chlorine atom on benzene ring. The results indicate that under moderate atmospheric conditions, the “addition” pathway is dominant compared to “substitution” pathway. Moreover, the research also calculated the carbon charge distribution on PCDD while the oxygen in OH radicals carries a partial negative charge to react with more positive carbon sites. It can be applied to predict which site on the benzene ring will be attacked by OH radicals first when the reaction takes place. In environment, OH radicals not only degrade PCDD/Fs but also react with other compounds simultaneously, especially with O2, NO and H2O. Zhang et al. (2011, 2012) conduct a direct dynamic calculation to find out the favorable pathways among dozens of element reactions of

3.3.2. Deposition Deposition processes of SVOCs can take place in dry gas, dry particulate and wet form. To be more specific, gas-phase SVOCs from ambient air being absorbed onto air-water/soil surface is considered as dry gas deposition while particle-phase SVOCs come to contact with water/soil surface is called dry particulate deposition. Wet deposition occurs when both gas-phase and particle-phase PCDD/Fs and dl-PCBs in ambient air are transported to water/soil surface by precipitation (Strachan and Eisenreich, 1988). Because the studies related to PCDD/ Fs and dl-PCBs deposition in East Asia are very limited, we also collect the recent works from other regions and provide a comparison with the Table 6 Tropospheric lifetimes via reaction of PCDD/Fs and PCBs with the global 24-h averaged OH radical concentration of an 8 x l05 molecule cm−3 or b 9.7 x l05bmolecule cm−3. Number of Cl atom

0 1 2 3 4 5 6 7 8

Tropospheric lifetime (days) PCDDs

PCDFs

a

a

1.0 3.0a 2.0–2.4 2.5–3.3a 2.8–7.2a (11b) 4.0–8.5a (21b) 45b 89b 230b

N/A: Not available. a (Kwok et al. (1995). b Brubaker and Hites (1997).

30

3.7 2.9a 4.0–5.5 5.5–9.5a 7.7–18a (19b) 15-29a (10b) 83b 190b 580b

PCBs 2.0a 2.7–5.1a 3.4–7.2a 6.9–15a 8.5–40a 16–48a 29–90a N/A N/A

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Huang et al. (2011a) that a monthly deposition from July 2009 to June 2010 of wet deposition contributed around 20% to the total deposition flux. The differences in atmospheric temperature, meteorological condition and emission source between sampling sites could lead to the contrary conclusion in different studies. Wet PCDD/F deposition fluxes were calculated with gas and particle scavenging theories by Huang et al. (2011a) in a rural area in Taiwan, and the results indicate that the deposition flux in rainy seasons (18.52 pg m−2 day−1) was significantly higher than those in dry seasons (0.369 pg m−2 day−1), revealing a positive relationship between wet deposition flux and rainfall. Compared to those measured in Europe, Castro et al. (2012) has predicted the wet deposition of PCDD/ Fs in northern Italy as a function of precipitation rate and the contaminant concentration in rain water. The estimated fluxes varied from 2 to 4660 pg m−2 day−1 for dl-PCBs and from 0.2 to 3540 pg m−2 day−1 for PCDD/Fs. For those measured in New Jersey, USA, on average, 97% of the total atmospheric washout of PCBs was resulted from particle scavenging, and wet deposition fluxes of PCBs were of the same order of magnitude as dry-particle deposition fluxes in all landuse regimes (Van Ry et al., 2002). Unlike wet deposition which is usually collected by a stainless steel or cylindrical glass jar, dry deposition of SVOCs is collected by using flat plates, because it provides minimum airflow disruption and a good estimation of the low limit for dry deposition (Shih et al., 2006). Shih et al. (2006) conducted the research on dry deposition of PCDD/Fs in six rural sites in southern Taiwan. In addition, the dry deposition velocity was also calculated with a seasonal variation ranging from 0.39 in winter to 0.52 cm s−1 in summer and indicated that HxCDD and HpCDD congeners were of higher deposition velocity than other PCDD congeners while for PCDF congeners, higher chlorinated congeners have higher deposition velocities. An average dry deposition flux Ʃ106 PCBs of 4730 pg m−2 day−1 was measured by Lee et al. (1996) in a southern urban site of Taiwan and dry deposition flux was 1.5–5.0 times higher than the wet deposition. The mean dry deposition fluxes of dl-PCBs were 651 pg m−2 day−1 (2.51 pg WHO-TEQ m−2 day−1) and 574 pg m−2 day−1 (0.903 pg WHOe TEQ m−2 day−1) in fall and spring, respectively (Mi et al., 2012). In other parts of the world, Castro et al. (2010) evaluate dry deposition flux between atmosphere and open seawater in Mediterranean Sea. The results indicate that dry deposition fluxes of PCDD/F in the Mediterranean Sea ranged from 5 to 170 pg m−2 day−1. Ʃ106 PCB dry deposition flux in Chicago was measured as 4.5 pg m−2 day−1 by Holsen et al. (1991) while those collected by Bozlaker et al. (2008) in Turkey industrial site (Ʃ41 PCB) reached 409 × 103 pg m−2 day−1. Deposition fluxes of PCDD/Fs and PCBs measured from studies in Asia as well as other parts of the world are presented in Tables 7 and 8, respectively.

results reported in East Asia. In most studies, the bulk deposition which is a combination of wet and dry deposition is collected and analyzed. One of the first studies on atmospheric bulk deposition in Asia was conducted in Kanto region, Japan (Ogura et al., 2001). Annual average PCDD/Fs deposition ranged from 1232 to 3561 pg m−2 day−1 (15–46 pg TEQ m−2 day−1) with the higher value being measured in winter than in summer. To evaluate the effect of seasonality on deposition flux, Oka et al. (2006) measured the deposition flux in Kanazawa city in winter and summer and found the deposition in winter is 10 times higher than that in summer due to the lower atmospheric dilution and higher emission in winter. Annual average deposition flux was 360 pg m−2 day−1, suggesting that this area was less polluted if compared to other areas in Japan. Fang et al. (2012) investigated bulk deposition of PCBs around an iron/steel making plant in Pohang, South Korea and the fluxes of 74 × 103 and 2.1 × 103 pg m−2 day−1 for Ʃtri-decaPCBs and ƩdlPCBs, respectively, were found in this area. In Europe, Hovmand et al. (2007) measured the bulk deposition in background area in Denmark and found that mean winter flux was twice as that measured in summertime. Moreover, in wintertime atmospheric PCDD/Fs concentrations were correlated with sulfur dioxide plus particulate sulfate, indicating the long-range transport from industrial areas of Central Europe. On the other hand, no significant correlation was found for summer samples. To evaluate the difference between automated sampler and traditional cylindrical vessels in northern Taiwan, Chi et al. (2009) indicate that PCDD/F deposition flux collected using the traditional sampler (2.0 pg I-TEQ m−2 day−1 to 9.9 pg I-TEQ m−2 day−1) was lower than that sampled with the automated sampler (17.5 pg I-TEQ m−2 day−1 to 25.8 pg I-TEQ m−2 day−1). The authors suggested that photodegradation and evaporation could reduce part of the PCDD/Fs collected by the traditional cylindrical vessels. In Taiwan, Chi et al. (2009) evaluate the relative importance of wet and dry deposition fluxes, indicating that the dry deposition flux of PCDD/Fs (12.3 pg I-TEQ m−2 sunnyday−1 to 16.7 pg I-TEQ m−2 sunny day−1) was lower than wet deposition flux (39.4 pg I-TEQ m−2 rainyday−1 to 228 pg I-TEQ m−2 rainyday−1). This research therefore concluded that PCDD/F majorly removed in the atmosphere by wet deposition. Similar result was archived by Correa et al. (2006) via separate but simultaneous sampling campaign in Texas, USA, demonstrating that the wet deposition process far exceeds the dry deposition by a factor of 8. Furthermore, dry deposition was preferentially dominated by the Hp-OCDD/Fs while wet deposition included a mixture TeOCDD/Fs. However, Lin et al. (2010) found a different result when calculating deposition flux of PCDD/Fs entering the drinking water treatment plant where dry deposition contributes 88% of the total flux while that of wet deposition is only 12%. Similar trend was reported by Table 7 Deposition fluxes of PCDD/Fs. Country

Japan

Location

Site

Tokyo Yokohama Tsukuba Tanzawa Kanazawa Southern Taiwan

Urban Urban Suburban Rural Urban Rural

Northern Taiwan

Rural

USA

Houston

Industrial

Denmark France

Forest site Thau lagoon

Rural Rural

Taiwan

Deposition flux

Reference

pg m−2 day−1

pg TEQ m−2 day−1

Bulk: 3561 Bulk: 2694 Bulk: 3333 Bulk: 1232 Bulk: 360 Dry: 150 Bulk: 408 Wet: 7.07

Bulk: 46 Bulk: 30 Bulk: 23 Bulk: 15 Bulk: 7.7 Dry: 8.96 13.75 0.148 Dry: 14.5 Wet: 133.7 Dry: 0.69 Wet: 12 Bulk: 2.77 N/A

Dry: 351 Wet: 2873 N/A PCDD Bulk: 117

*N/A: Not available.

31

(Ogura et al., 2001)

(Oka et al., 2006) (Shih et al., 2006) (Huang et al., 2011b) (Chi et al., 2009) (Correa et al., 2006) (Hovmand et al., 2007) (Castro et al., 2011)

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Table 8 Deposition fluxes of PCBs. Country

Location

Site

Taiwan

Tainan

S.Korea

Pohang

Urban Urban/industrial/rural Industrial

USA

Chicago

Urban Resident

New Jersey

Urban Suburban Background

Turkey

Izmir

Industrial

France

Thau lagoon

Rural

Deposition flux

Reference

pg m−2 day−1

pg TEQ m−2 day−1

Ʃ106 PCBs (Dry): 4730 dl-PCBs (Dry): 574 to 651 dl-PCBs (Bulk): 2.1 × 103 Ʃ106 PCBs (Dry): 4.5 × 106 Ʃ50 PCBs (Dry): 190 × 103 Ʃ89 PCBs (Wet): 104 to 4 × 104 Ʃ89 PCBs (Wet): 0.9 × 103 to 3 × 103 Ʃ89 PCBs (Wet): 0.8 × 103 to 2 × 103 Ʃ41 PCBs (Dry): 409 × 103 Ʃ18 PCB Bulk: 715

N/A dl-PCBs (Dry): 0.903 to 2.51 N/A

(Lee et al., 1996) (Mi et al., 2012) (Fang et al., 2012)

N/A

(Holsen et al., 1991)

N/A

(Tasdemir et al., 2004)

N/A

(Van Ry et al., 2002)

N/A N/A N/A

(Bozlaker et al., 2008)

N/A

(Castro et al., 2011)

*N/A: Not available.

particles (TSPs) were monitored at a remote area of Lulin background station in central Taiwan by Chi et al. (2010). During the burning event, the PCDD/F concentration in ambient air at Lulin station (including both gas and particle phase) increased from 2.33 to 390 fg I-TEQ m−3 while the particle-phase PCDD/Fs in the TSP increased from 28.7 to 109 pg I-TEQ/g-TSP. The same trend was observed when significantly high levels of atmospheric PCDD/Fs were recorded at Lulin mountain in central Taiwan and in the source region of Northern Vietnam days during Asia biomass burning event (Chi et al., 2016). Moreover, to evaluate the effects of LRAT via northeast monsoon from the coastal regions China to Taiwan, atmospheric PCDD/F levels were monitored at two background stations and remote islands around Taiwan (Chi et al., 2013b, 2014). Significantly lower atmospheric PCDD/F concentrations and PCDD/F contents in TSPs were measured in Lulin station during the summer season when the air mass originated from the Pacific Ocean and South China Sea.

3.3.3. Evidence of long-range atmospheric transport Long-range atmospheric transport (LRAT) is defined as the presence of compound which is detected far away from its emission source. Discrepancy still exists regarding how LRAT affects the PCDD/Fs and dl-PCBs congener distribution in remote areas. The best evidence of LRAT of PCDD/Fs and dl-PCBs is found in polar regions where it is free of the effects from local emissions. Weathering process during LRAT affected the congener distribution of PCDD/Fs and dl-PCBs as the highly chlorinated compounds are more likely to accumulate in particle and be removed by deposition, as discussed earlier, stronger than lower chlorinated compounds. Oehme et al. (1996) reported that the PCDD/ Fs concentration in Arctic region ranged from 16 to 28 fg m−3. When the PCDD/Fs and PCBs congener profiles of Lancaster University (UK) sampling station were compared with those measured in Artic region, Lohmann and Jones (1998) found that lighter compounds including TeHxCDD/Fs and Tri-HxCBs in Lancaster University was only 10 times higher than that measured in Arctic region while those of heavy compounds obtained the difference of 50–100 times. However, there are some studies reported that high distribution of highly chlorinated congeners is due to the increase of particle phase which is transported from the higher polluted area during LRAT event while low chlorinated compounds were decomposed by abundant OH radicals in lower latitude regions. Chi et al. (2008) evaluate the influence of Asian dust storm (ADS) that originated in the deserts areas of Mongolia and northern China and eventually reached East Asia, including Taiwan. The results indicated that before the ADS episode, about 64% and 68% total concentration on PCDD/Fs distributed in particle phase in the northern coast of Taiwan and Taipei city, respectively. During the ADS episode 70% of total PCDD/Fs distributed in particle phase in northern coast while those in Taipei city reached 90%. The particle partitioning after ADS episode were reduced to 60% and 75% at the northern coast and Taipei city, respectively. dl-PCBs in gas phase account for 90% of the total dl-PCBs concentration in both north coast of Taiwan and Taipei city. The partitioning of gas-phase dl-PCBs did not change significantly during ADS episode because vapor pressures of dl-PCB congeners are much higher than those of PCDD/F congeners and would not be much affected by the increasing particulate matter during ADS episode. Moreover, Thuan et al. (2013) indicate that during the ADS episode, the atmospheric PCDD/F concentrations increased by 6.5 and 6.9 times in southern Taiwan and South China Sea, respectively. The other evidence of LRAT is related to the effect of Southeast Asia biomass burning event, and the concentrations of total suspended

4. Global monitoring plan and temporal trends in East Asia The Stockholm Convention took effect in 2004 to protect human health and the environment by reducing and eliminating the release of POPs. According to Article 16, comparable monitoring periodical data of the listed POPs (12 initial and 16 new POPs) are required to evaluate their occurrence and long range transport. The Global Monitoring Plan (GMP) is developed to fulfill this objective and provides a harmonized framework to identify the changes in concentrations of POPs over time, as well as information on their regional and global environmental transport (United Nations Environment Programme, 2016a). At this moment, the first (2000–2008) and second (2009–2015) phases of GMP in East Asia have been accomplished with the participation of 12 countries (Cambodia, Indonesia, China, Japan, South Korea, Laos, Malaysia, Mongolia, Philippines, Singapore, Thailand and Vietnam). However, only the data on PCDD/Fs and dl-PCBs in ambient air in Japan and China (including Hongkong) are available in these reports (United Nations Environment Programme, 2016b). In this review, we present more data reported in Taiwan, South Korea, Thailand and Vietnam to provide better knowledge related to the spatial and temporal trends of PCDD/Fs and dl-PCBs in East Asia. In Western Europe, dioxin concentrations in air are found to decline and the concentrations measured in urban areas are now close to those in rural areas, implying that most major and readily controllable sources have been reduced and current levels in both rural and urban areas remain close (United Nations Environment Programme, 2016c). 32

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Chi et al., 2013a) and are summarized in Fig. 6 to provide the time trend of PCDD/Fs and dl-PCBs in some Asian countries from 2003 to 2015.

In Asia, a temporal trend of atmospheric levels of PCDD/Fs and dl-PCBs was studied in South Korea from 1998 to 2008 (Shin et al., 2011). The results indicated that mean concentrations decreased from 425 to 28 fg I-TEQ m−3 from 1999 to 2008 (93% reduction) and from 48 to 1 fg WHO-TEQ m−3 from 2002 to 2008 (94% reduction) for PCDD/Fs and dl-PCBs, respectively. Since 1997, the Korean government has regulated PCDD/Fs emissions from MSWIs (capacity > 50 ton d−1, for existing incinerator: 0.5 ng I-TEQ m−3, for newly built incinerator: 0.1 ng I-TEQ m−3). In 2004 regulation has been setup for small-size incinerators which became the major point source of PCDD/Fs emission in Korea (capacity > 25 kg h−1, old 10 ng I-TEQ m−3, new 5 ng I-TEQ m−3), resulting in a significant decrease of PCDD/Fs emissions. In Taiwan, after enacting the regulation on PCDD/Fs emission from MSWIs in 2001 (capacity > 300 ton d−1: 0.1 ng I-TEQ m−3) (Taiwan-EPA, 2015), Lee et al. (2004b) found a significant reduction of PCDD/Fs level in ambient air of Taiwan from 300 fg I-TEQ m−3 in 1999 to nearly 70 fg I-TEQ m−3 in 2003. Taiwan EPA continued to conduct monitoring program from 2006 to 2015 and found that the average concentration did not show a significant decreasing trend (Taiwan-EPA, 2015). According to the United Nations Environmental Programme monitoring report, the PCDD/Fs level in ambient air in Hongkong showed a slightly decreasing trend from 85 fg I-TEQ m−3 in 1998 to 41 fg I-TEQ m−3 in 2013 (United Nations Environment Programme, 2016b) while the trend of PCBs in ambient air also decreased from 560 fg m−3 in 1998 to 320 fg m−3 in 2006. In Japan, PCDD/Fs and dl-PCB levels monitored throughout the nation were greatly reduced from 0.55 pg TEQ m−3 in 1997 to 0.029 pg TEQ m−3 in 2006 (United Nations Environment Programme, 2016b). The second report of GMP in East Asia indicates that there are 11 background, 3 rural, 3 urban sampling sites in mainland China that monitor PCDD/Fs and dl-PCBs in ambient air between 2008 and 2011. Although recent studies in China's cities still report relatively higher concentrations of PCDD/Fs and dl-PCBs compared to other East Asian countries, a significant decreasing trend was observed for the average atmospheric concentration of these pollutants in background sites from 2008 (PCDD/Fs: 19.6 fg WHO-TEQ m−3; dlPCBs: 1.06 fg WHO-TEQ m−3) to 2011 (PCDD/Fs: 13.2 fg WHO-TEQ m−3; dl-PCBs: 0.77 fg WHO-TEQ m−3) (United Nations Environment Programme, 2016b) which indicates that there are effective regulations to control PCDD/Fs and dl-PCBs in China. The second report also indicates that POPs sampling has been conducted in South Eastern Asian countries but no monitoring data related to PCDD/Fs and dl-PCBs in ambient air have been reported in these countries. The data of Thailand and Vietnam are collected from independent studies (Ngo et al., 2017;

5. Conclusions and suggestions for future work The atmospheric PCDD/Fs concentrations summarized in East Asia range from 49.8 fg I-TEQ m−3 to 3030 fg I-TEQ m−3 in industrial sites, 8 to 1170 I-TEQ m−3 in urban sites, and 1.01 to 630 I-TEQ m−3 in rural sites. Different PCDD/Fs and dl-PCBs levels in each country could be attributed to local emission sources, pollution control policy and the sampling location characteristics including geographic and meteorological features. Although there is no consensus regarding how the LRAT affects congener distribution of POPs, it has been proven that LRAT increases the pollutant concentration in ambient air far away from the sources. PCDD/Fs distribute majorly in particle phase while those of dl-PCBs exist mostly in gas phase due to the difference of vapor pressures. The Junge–Pankow adsorption and the Harner–Bidleman absorption models have been applied to predict the distribution of these compounds. Generally, the Junge–Pankow model predicts more precisely for particle-phase distribution of dl-PCBs while that of PCDD/Fs better fits the prediction of Harner–Bidleman model. However, there is still discrepancy between the modeled and measured values due to the difference of heat of condensation of congeners or particle surface properties in different studies. It also requires a longer time for PCDD/Fs and dl-PCBs to reach gas/particle equilibrium state in low-temperature environment in winter period compared to summer time, resulting in less accuracy of the model. Thus, both models need to be improved for better prediction. Gas-phase POPs are removed mostly by reacting with OH radicals in atmosphere while those in particle phase are mainly removed by wet and dry deposition processes. Although it is proven that PCDD/Fs distribute more in smaller particle such as PM10, PM2.5 or even ultrafine particles, most studies in this region focus on PCDD/Fs concentration in total suspended particle (TSP). Therefore, future research with the support of advanced sampling instruments should be conducted to better understand the behavior of POPs in general, PCDD/Fs and dlPCBs in particular in the atmosphere. The temporal and spatial levels of PCDD/Fs and dl-PCBs indicate that strict regulation enacted has effectively reduced the PCDD/Fs and dl-PCBs emissions in East Asian countries. More efforts are needed to maintain and improve the air quality in both developed and developing

PCDD/Fs concentration (fg I-TEQ m-3)

300

Taiwan Hongkong

250

Japan S.Korea

200

China (background sites) Vietnam

150

Thailand (background site)

100 50 0 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015

Fig. 6. Trend of PCDD/Fs in different countries in East Asia. (Data are from (Shin et al., 2011), (Lee et al., 2004b), (United Nations Environment Programme, 2016b), (Ngo et al., 2017), (Chi et al., 2013a) and (Taiwan-EPA, 2015)).

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countries. Significant relocation of industrial manufactory from China to South East Asia is taking place due to the increasing workforce wage in recent years. However, study focusing on level of PCDD/F in ambient air in South East Asia countries is still lacking. Most of the countries in Southeast Asia do not have enough funding for the technology development and transfer. In addition, the techniques and knowledge of specialists in the region cannot meet the standard of up-to-date research skills and data analyses on dioxin-like compounds, as well as the expensive sampling cost, insufficient monitoring sites, lack of programs on the emission control and lack of quality control attributed to the limited facilities for PCDD/Fs and dl-PCBs analysis and monitoring. PS technique is a promising and cost-effective sampling method for developing countries but further research is still needed to provide the accurate uptake rate of each congener in specific condition as mentioned earlier. More regional and international cooperation should be made to transfer knowledge and technologies to this region. National and international funding agencies such as the World Bank and Global Environment Fund (GEF) should provide more support to PCDD/Fs and dl-PCBs monitoring and analytical activities, and encourage more studies on these chemicals to fill the data and technical gap on the POPs measurement and control in the region.

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