Environment International 139 (2020) 105731
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Environment International
journal homepage: www.elsevier.com/locate/envint
Severe dioxin-like compound (DLC) contamination in e-waste recycling T
areas: An under-recognized threat to local health
Qingyuan Daia, Xijin Xub, Brenda Eskenazic, Kwadwo Ansong Asanted, Aimin Chene, Julius Fobilf,
Åke Bergmang,h,i, Lesley Brennanj, Peter D. Slyk, Innocent Chidi Nnoroml, Antonio Pascalem,
Qihua Wanga, Eddy Y. Zenga, Zhijun Zengb, Philip J. Landrigann, Marie-Noel Bruné Drisseo,
Xia Huoa,⁎
aGuangdong Key Laboratory of Environmental Pollution and Health, School of Environment, Jinan University, China
b Laboratory of Environmental Medicine and Developmental Toxicology, Shantou University Medical College, China
c School of Public Health, University of California, Berkeley, USA
d CSIR Water Research Institute, Accra, Ghana
e Department of Biostatistics, Epidemiology and Informatics, University of Pennsylvania, USA
f School of Public Health, University of Ghana, Ghana
g Department of Environmental Science, Stockholm University, Sweden
hDepartment of Science and Technology, Örebro University, Sweden
i College of Environmental Science and Engineering, Tongji University, China
jDepartment of Obstetrics and Gynaecology, University of Alberta, Canada
k Child Health Research Centre, University of Queensland, Australia
l Department of Pure and Industrial Chemistry, Abia State University, Nigeria
mDepartment of Toxicology, University of the Republic, Uruguay
nDepartment of Biology, Boston College, USA
o Department of Environment, Climate Change and Health, World Health Organization, Geneva, Switzerland
A R T I C L E I N F O A B S T R A C T
Handlng Editor: Adrian Covaci Electrical and electronic waste (e-waste) burning and recycling activities have become one of the main emission sources
Keywords: of dioxin-like compounds (DLCs). Workers involved in e-waste recycling operations and residents living near e-waste
Dioxin-like compound recycling sites (EWRS) are exposed to high levels of DLCs. Epidemiological and experimental in vivo studies have
Toxicity reported a range of interconnected responses in multiple systems with DLC exposure. However, due to the composi-
E-waste recycling site tional complexity of DLCs and difficulties in assessing mixture effects of the complex mixture of e-waste-related con-
Human exposure taminants, there are few studies concerning human health outcomes related to DLC exposure at informal EWRS. In this
Health risk paper, we have reviewed the environmental levels and body burdens of DLCs at EWRS and compared them with the
levels reported to be associated with observable adverse effects to assess the health risks of DLC exposure at EWRS. In
general, DLC concentrations at EWRS of many countries have been decreasing in recent years due to stricter regulations
on e-waste recycling activities, but the contamination status is still severe. Comparison with available data from in-
dustrial sites and well-known highly DLC contaminated areas shows that high levels of DLCs derived from crude e-
waste recycling processes lead to elevated body burdens. The DLC levels in human blood and breast milk at EWRS are
higher than those reported in some epidemiological studies that are related to various health impacts. The estimated
total daily intakes of DLCs for people in EWRS far exceed the WHO recommended total daily intake limit. It can be
inferred that people living in EWRS with high DLC contamination have higher health risks. Therefore, more well-
designed epidemiological studies are urgently needed to focus on the health effects of DLC pollution in EWRS.
Continuous monitoring of the temporal trends of DLC levels in EWRS after actions is of highest importance.
Abbreviations: DLC, dioxin-like compound; e-waste, electrical and electronic waste; EWRS, e-waste recycling site; PCDD/F, polychlorinated dibenzo-p-dioxin and
dibenzofuran; DL-PCB, dioxin-like polychlorinated biphenyl; PBDD/F, polybrominated dibenzo-p-dioxin and dibenzofuran; PXDD/F, mixed halogenated dibenzo-p-
dioxin and dibenzofuran.; TCDD, 2,3,7,8-tetrachlorodibenzo-p-dioxin; AhR, aryl-hydrocarbon receptor; TEQ, toxic equivalent; TEF, toxic equivalency factor; dw, dry
weight; PeCDF, pentachlorodibenzofuran; bw, body weight; SWHS, Seveso Women’s Health Study; TSH, thyroid stimulating hormone; T4, thyroxine; T3, triio-
dothyronine; BMI, body mass index
⁎ Corresponding author at: Laboratory of Environmental Medicine and Developmental Toxicology, School of Environment, Jinan University, Guangzhou 511443,
China.
E-mail address: xhuo@jnu.edu.cn (X. Huo).
https://doi.org/10.1016/j.envint.2020.105731
Received 22 November 2019; Received in revised form 7 April 2020; Accepted 7 April 2020
Available online 18 April 2020
0160-4120/ © 2020 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND license 
(http://creativecommons.org/licenses/BY-NC-ND/4.0/).
Q. Dai, et al. Environment International 139 (2020) 105731
1. Introduction 2016; van den Berg et al., 2006). 2,3,7,8-Substituted polybrominated
and mixed halogenated dibenzo-p-dioxins and dibenzofurans (PBDD/Fs
With the rapid development of the economy and technology, the and PXDD/Fs) have chemical structures similar to PCDD/Fs and have
replacement cycle of electronic products is getting shorter, which has been recommended to be included in the WHO toxic equivalency factor
resulted in electrical and electronic waste (e-waste) growing at an (TEF) scheme (van den Berg et al., 2013; WHO, 1998a). TEFs are as-
alarming rate, threatening the environment and human health signed based on the relative potency of each dioxin congener to induce
(Heacock et al., 2016; Awasthi et al., 2018; Kumar et al., 2017; Bakhiyi AhR activation compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin
et al., 2018; Parajuly et al., 2019; Pascale et al., 2018; Xu et al., 2015). (TCDD), the most toxic of these DLCs, for which TEF value is fixed to
The Global E-waste Monitor 2017 shows that the total amount of the 1.0 (van den Berg et al., 2006). Total toxicity of such a complex mixture
world’s e-waste generation in 2016 reached 44.7 million tons, on mediated by the aggregate activation of AhR receptors is expressed as
average 6.1 kg per inhabitant (Balde et al., 2017). It is reported that TCDD toxic equivalent (TEQ) which is calculated by summing the
80% of e-waste generated globally has been treated informally (Balde multiplication products of congener concentrations with congener-
et al., 2017), especially in developing countries of Asia and Africa specific TEFs (van den Berg et al., 1998, 2006). Many studies reported
(Breivik et al., 2014), including the cities of Guiyu, Qingyuan, and that PBDD/Fs contribute larger dioxin TEQs than PCDD/Fs and DL-
Taizhou, China (Chan et al., 2007; Leung et al., 2006; Wen et al., 2009; PCBs in EWRS (Xiao et al., 2016; Tue et al., 2016, 2019; Ma et al.,
Zhang et al., 2017b); Accra, Ghana (Tue et al., 2019; Wittsiepe et al., 2009). However, they are less studied because of the complexity of
2015); Bengaluru and Delhi, India (Chakraborty et al., 2018; Karri analytical procedures and paucity of analytical reference standards
et al., 2008); Lagos, Nigeria (Iwegbue et al., 2019); and Trang Minh and (Zhang et al., 2016a).
Bui Dau, Vietnam (Kincaid, 2019; Tue et al., 2010) (Fig. 1). China once Dioxin-like compounds are released as unintentional byproducts in
received the largest share of e-waste from around the world (Zhang low-tech e-waste recycling operations including manual disassembly,
et al., 2012a), and has surpassed the United States to become the shredding/comminution, roasting circuit boards, acid-stripping metals,
world’s largest producer of e-waste (Zeng et al., 2016). After the ban on and open burning of e-waste containing chlorinated polymers and/or
the import of foreign solid waste effective 1st January 2018 (Fu et al., brominated additives such as polyvinyl chlorides and brominated flame
2018), predominant sources of e-waste in China have changed to do- retardants (Wong et al., 2007; Li et al., 2007; Leung et at., 2006b; Xiao
mestic generation. This move leaves a potential for the South Asian and et al., 2016; Duan et al., 2011; Zennegg et al., 2014). PBDD/Fs on the
African regions, especially India, to receive and recycle e-waste (Asante other hand can be generated from chemical reaction, photochemical
et al., 2019; Garg and Adhana, 2019). degradation and thermolysis of plastics containing brominated flame
Primitive crude recycling processes in informal e-waste recycling retardants (Weber and Kuch, 2003; Ebert and Bahadir, 2003; Kajiwara
sites (EWRS) (unregulated, unregistered, and low-technology units) et al., 2008; Kannan et al., 2012). DL-PCBs are also found in dielectric
result in the environmental release of dioxin-like compounds (DLCs) (Ni fluids, lubricants and coolants in generators, capacitors and transfor-
et al., 2010; Liu et al., 2008; Vaccari et al., 2019; Leung, 2019). The mers (WHO/IPCS, 1992). Furthermore, the production of chlorinated
term “DLCs” is a general name for structurally and chemically related chemicals such as pesticides, herbicides, insecticides, and processes like
planar aromatic hydrocarbons with the capability of binding to the aryl- metal smelting, paper production, petroleum refining, and incomplete
hydrocarbon receptor (AhR) and dioxin-like toxicity (van den Berg combustion of municipal, medical and industrial waste can unin-
et al., 2006, 2013). The most toxic DLCs include 17 congeners of tentionally produce PCDD/Fs and DL-PCBs (Kulkarni et al., 2008;
2,3,7,8-polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/ Ssebugere et al., 2019; Zhu et al., 2008; Zheng et al., 2008; Trinh et al.,
Fs) and 12 dioxin-like polychlorinated biphenyls (DL-PCBs) (WHO, 2018). PCDD/Fs and DL-PCBs persist widely in the environment and
Fig. 1. E-waste production, flows, and recycling sites with the highest soil PCDD/F and PBDD/F concentrations reported (Chen et al., 2011; Balde et al., 2017;
Eurostat Statistics Explained, 2019; Pascale et al., 2018).
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Q. Dai, et al. Environment International 139 (2020) 105731
bioaccumulate through the food chain because they are lipophilic and damage (Wang et al., 2018; He et al., 2015; Grant et al., 2013; Wen
resistant to biological and chemical degradation (Sadler and Connell, et al., 2008).
2012). The half-life of PCDD/Fs and DL-PCBs in the body is estimated to Due to the ubiquity of DLCs and their high toxicity, humans po-
be 7–11 years (WHO, 2016). PBDD/Fs are more lipophilic and more tentially have chronic exposure through the environment throughout
sensitive to UV degradation, and appear to be less persistent in the their entire life which may pose a serious threat to public health (Berry
environment (WHO, 1998a). However, bioaccessibility of PBDD/Fs and et al., 1993). In addition, accidental and/or occupational exposure may
PXDD/Fs and their persistence in the human body are still un- occur (Hens et al., 2016). Even a small daily exposure can accumulate
determined (Piskorska-Pliszczynska and Maszewski, 2014). to yield detectable amounts over time. It can be inferred that people
Many epidemiological studies have reported health effects of PCDD/ living in EWRS with heavy DLC contamination are at higher health
F and DL-PCB exposure, such as skin diseases, reproductive and de- risks. There have been several reviews on DLC levels in EWRS (Tue
velopmental abnormalities, nervous system disorders, immune defi- et al., 2013; Chan and Wong, 2013) and many studies on the health
ciency, cancer promotion, and endocrine disruption (Kogevinas, 2001; effects of DLC exposure. However, partly due to the huge toxicological
Marinkovic et al., 2010; Schecter, 2013; World Health Organization data gap for the possible mixture effects of a large number of e-waste-
WHO, 2010). Although there are limited data on toxicity of PBDD/Fs to related contaminants and limited techniques and finances in developing
human health, comparable biological and toxic effects as their chlori- countries, current data on DLCs in EWRS provide very limited in-
nated analogues (PCDD/Fs) have been found in mammalian and fish formation regarding body burdens and health impacts on e-waste
models (van den Berg et al., 2013; WHO, 1998a; Mennear and Lee, workers and local residents (Wang et al., 2019; Zhang et al., 2010; Xu
1994; Birnbaum et al., 2003). The follow-up studies include occupa- et al., 2014), especially the vulnerable populations of pregnant women
tional or accidental high PCDD/F and DL-PCB exposure cohorts in Se- and developing children. Therefore, it is critical to assess the DLC ex-
veso, Italy (Bertazzi et al., 1998; Consonni et al., 2008; Eskenazi et al., posure and health risk of people in EWRS. In this context, we identified
2018; Slama et al., 2019; Warner et al., 2013b, 2019), Japan (Kondo and reviewed the published literature available online up to November
et al., 2018; Nagayama et al., 2001; Tsukimori et al., 2008), Germany 2019, by using PubMed, Web of Science, Google Scholar, and Medline
(BASF Study) (Ott et al., 1993, 1994, Zober et al., 1994, 1997), the to search the combined keywords “dioxins” and “health effects,” “e-
Netherlands (McBride et al., 2009, 2018; Mannetje et al., 2018; waste recycling site” for health effects of DLCs and their exposure levels
Mannetje et al., 2005), the Ranch Hand Vietnam Veterans (The Air at EWRS. We compare the DLC body burdens at EWRS with the levels
Force Health Study) (Buffler et al., 2011; Knafl, 2018; Wolfe et al., reported to be associated with observable adverse effects to help better
1990, 1995), and the US (National Institute for Occupational Safety and understand the contamination status and human health risks of DLCs at
Health) (Calvert et al., 1999; Ruder and Yiin, 2011; Sweeney et al., EWRS.
1997). Background exposed general population cohorts have also been
studied in Norway (Caspersen et al., 2016a, 2016b), Germany (the
Duisburg birth cohort study) (Neugebauer et al., 2015; Wilhelm et al., 2. Health hazards of DLCs
2008; Nowack et al., 2015; Winneke et al., 2014), Japan (the Hokkaido
Study on Environment and Children's health) (Kishi et al., 2010, 2013; Dioxin-like compounds interrupt multiple systems and organ func-
Miyashita et al., 2018a; Nakajima et al., 2006, 2017), and Taiwan, tions of the body (nervous, endocrine, immune, and reproductive sys-
China (Su et al., 2010, 2012; Wang et al., 2005). Chloracne and cancer tems), mediated through an interaction with the AhR (National
are the only TCDD-induced effects established with certainty (Warner Research Council, 2006; Schecter, 2013; Lundqvist et al., 2008; Mandal,
et al., 2013a; Mannetje et al., 2018). The possible mechanisms of DLC- 2005). It induces inappropriate modulation of gene expression and an
induced health hazards may include excessive oxidative stress and inflammatory response which represent the initial steps in a succession
oxidatively generated damage to DNA and lipids (Zhang et al., 2019). of biochemical reactions, cellular and tissue changes that lead to the
People working or living in EWRS show evidence of greater DNA observed toxicity and health effects (Fig. 2) (Lindén et al., 2010;
Tuomisto, 2019; Mandal, 2005). Epidemiological studies indicate that
Fig. 2. Toxicity mechanism and health effects of PCDD/Fs and DL-PCBs (Lindén et al., 2010; Tuomisto, 2019).
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Q. Dai, et al. Environment International 139 (2020) 105731
4
Table 1
Epidemiological studies examining the effects of PCDD/F and DL-PCB exposure.
Country Population Exposure level Health effect Reference
Neurotoxicity (children)
Japan 42 mother-neonates pairs Breast milk (mean) 0.5a; 8.6b ↓ Newborn head circumference, r = 0.424, p < 0.01 (Nishijo et al., 2007b)
Japan 134 mother-infant pairs Maternal blood (mean) 11.9b; 18.8c 6 m, PCDD/Fs ↓ Mental development, β = − 0.22, p = 0.014; DLCs ↔ Mental and (Nakajima et al., 2006)
motor development
Japan 191 (6 m) and 122 (8 m) mother- Maternal blood (mean) 9.8b; 15.1c 6 m, DLCs ↓ Motor developmental, p < 0.05; 18 m, Six DLC isomers ↑ Mental (Nakajima et al., 2017)
infant pairs developmental in female, p < 0.05
Da Nang, Vietnam 216 mother-infant pairs Breast milk (mean) 12.5b 4 m ↓ Cognitive scores, p = 0.009; 1 y ↓ Social–emotional scores, p = 0.049; 3 y ↓ (Pham et al., 2015; Tai et al., 2013,
Expressive communication scores in boys, p = 0.03; 5 y ↓ Total test and balance 2016; Tran et al., 2016)
scores in boys, p < 0.05
Norway 44,092 (3 y) and 1024 (3.5 y) Maternal dietary intake (median) 0.6c1 3 y ↑ Language delay in girls, OR = 2.9 (1.1–7.1); 3.5 y ↔ IQ scores; Executive (Caspersen et al., 2016a, 2016b)
mother–child pairs functions; Attention Deficit/ Hyperactivity Disorder
Germany 100 (9–10 y) and 117 (8.5 y) Maternal blood (median) 12.9b; 13.2b 9–10 y ↓ Social Responsiveness Scale scores, β = − 6.66, p < 0.05; 8.5 y ↑ Omission (Nowack et al., 2015; Neugebauer
mother–child pairs errors in the subtest Divided Attention, exp (β) = 1.47 (1.08–2.00), p < 0.05 et al., 2015)
Seveso, Italy 161 children of 120 mothers Maternal serum (median) 1976 74.6a; 7–17 y ↔ Reverse learning; Memory; Attention/Impulsivity; Non-verbal intelligence (Ames et al., 2019)
Pregnancy 4.5a
Neurotoxicity (adults)
Seveso, Italy 154 women for physical Serum (median) 1976 45.2a; 60.1a ↓ Grip strength, inverted U-shaped association; ↔ Walking speed, upper body (Ames et al., 2018a)
functioning and 459 for working mobility, manual dexterity, working memory
memory
Goto, Japan 140 Yusho patients Blood 2,3,4,7,8-PeCDF ≥ 72.27 pg/g lipid ↑ Pittsburgh Sleep Quality Index global score ≥ 8, OR = 4.84 (Kondo et al., 2018)
(1.10–21.25)
Reproductive toxicity (males)
Japan 183 mother–child pairs Maternal blood (median) 14.5c ↓ Testosterone/estradiol ratio, sex hormone-binding globulin, and inhibin B levels in (Miyashita et al., 2018a)
male; ↑ Adrenal androgen/glucocorticoid ratios, follicle-stimulating hormone, and
dehydroepiandrosterone levels in male
Seveso, Italy 39 exposed mother-son pairs vs 58 Maternal serum (median) 1976 51.7a vs 10.0a; ↓ Sperm concentration, progressive motility, total motile count, inhibin B, p < 0.05; (Mocarelli et al., 2011)
reference pairs Conception 26a vs 10.0a ↑ Follicle-stimulating hormone
Seveso, Italy 71 male (1–9 y); 44 (10–17 y); 20 Serum (median) 1976 210.0a; 164.0a; 123.0a; 1–9 y in 1976 ↓ Sperm concentration, motile sperm, and estradiol; ↑ FSH; 10–17 y in (Mocarelli et al., 2008)
(18–26 y) 1976 ↑ Sperm concentration, motile sperm, and FSH; ↓ Estradiol; ↔ 17–26 y in 1976
Chapaevsk, Russia 473 boys (8–9 y); 315 (17–18 y) Serum (median) 21.1c ↓ Pubertal onset [testicular volume = 11.6 months (3.8–19.4) and (Burns et al., 2016)
genitalia = 10.1 months (1.4–18.8)]; ↓ Sexual maturity [testicular
volume = 11.6 months (5.7–17.6); genitalia = 9.7 months (3.1–16.2)]
Bien Hoa, Vietnam 42 male residents Blood (mean) 37.8c ↑ Prolactin levels; ↓ Testosterone levels; ↔ Follicle-stimulating hormone luteinizing (Van Luong et al., 2018)
hormone, or progesterone levels
Reproductive toxicity (females)
Bien Hoa, Vietnam 162 mother-infant pairs Breast milk (mean) 8.8b ↓ Testosterone in girls, β = − 0.23, p < 0.05; ↔ Estradiol (Boda et al., 2018)
Seveso, Italy 310 women; 616; 278; 601 Serum (median) 1976 67.5a; 43.7a; 50.0a; ↑ Menstrual cycle, 0.93 days in women who were premenarchal in 1976; ↑ Early (Eskenazi et al., 2002a, 2002b,
54.9a menopause, 20.1–100 ppt; ↑ Time to pregnancy, 25%; ↑ Infertility, OR = 1.9 2005, 2010)
(1.14–3.22); ↑ Endometriosis, ≥ 100 ppt
Taiwan, China 56 mother–child pairs Placenta (median) 14.8c 8 y ↓ Estradiol, p = 0.003; ↓ Fundi and uteri lengths in girls, p = 0.016 (Su et al., 2012)
Developmental toxicity
Norway 50,651 mother–child pairs Maternal dietary intake (median) 0.6c1 ↓ Birth weight, 62 g (−73, −50); ↓ Birth length, 0.26 cm (−0.31, −0.20); ↓ Head (Papadopoulou et al., 2013)
circumference, 0.10 cm (−0.14, −0.06); ↓ Male proportion, 0.483 (0.46, 0.51); ↑
Spontaneous stillbirth, RR = 2.16 (1.58–2.97)
Japan 10,959 residents PCBs + PeCDFs 1968–1977 ↑ Spontaneous stillbirth rates ratio, 2.16 (1.58–2.97); ↓ Sex ratio (male (Yorifuji et al., 2013)
proportion), 0.483 (0.457–0.508); 1978–1987 ↑ Spontaneous stillbirth rates ratio, 1.8
(1.25–2.60)
Japan 214 Yusho women Blood at delivery PCBs + PeCDFs 1968–1977 ↑ Induced abortion and preterm delivery, p = 0.03; 1968–1977 ↑ (Tsukimori et al., 2008)
Spontaneous abortion, p = 0.11; pregnancy loss, p = 0.08; ↔ 1978–2003
Seveso, Italy 510 women (1996–1998) and 611 Serum (median) in 1976 and estimated at 1996–1998 ↔ Spontaneous abortions, birth weight, births small for gestational age; ↓ (Eskenazi et al., 2003; Wesselink
(2008–2009) pregnancy 1996–1998 46.6a; 2008–2009 Gestational age, 1.0–1.3 day; ↑ Preterm delivery, 20–50%; 2008–2009 ↔ Spontaneous et al., 2014)
55.0a; 9.9a abortion, fetal growth, or gestational length; ↓ Birthweight, β = − 22.8 (−80.1,
−34.6)
(continued on next page)
Q. Dai, et al. Environment International 139 (2020) 105731
5
Table 1 (continued)
Country Population Exposure level Health effect Reference
Seveso, Italy 239 men and 296 women Serum (median) 1976 Fathers 96.5a, mothers ↓ Sex ratio in fathers < 19 years in 1976 and TCDD > 15 ppt, 0.38 (0.30–0.47); ↔ (Mocarelli et al., 2000)
62.8a Sex ratio in fathers ≥ 19 years in 1976 and TCDD > 15 ppt
Seveso, Italy 611 children of 402 mothers Maternal serum (median) 1976 63.2a 2–39 y ↓ BMI in daughters; ↔ In sons (Warner et al., 2019)
Belgian, Norway, Slovak 367 infant for growth (0–24 m); Cord blood or breast milk (mean) 18.2c2 0–24 m ↑ Growth, β = 0.07 (−0.01, 0.14); 7 y ↑ BMI in girls but not in boys, β = 0.49 (Iszatt et al., 2016)
251 for BMI (7 y) (0.07, 0.91); ↑ Overweight in girls, 54% (−6%, 151%)
Taizhou 50 exposed mother–child pairs vs Breast milk (mean) 12.9b vs 4.3b 6 m ↔ Height; 3 y ↑ Height in girls, p = 0.001; ↓ Height in boys, p = 0.046 (Wang et al., 2019)
50 reference pairs
Endocrine toxicity
12 countries (1939–1992) 21,863 workers Blood TCDD and highly-chlorinated dioxins ↑ Diabetes, RR = 2.25 (0.53–9.50) (Vena et al., 1998)
New Zealand 245 workers Serum (mean) 15.2a ↑ Diabetes, workers worked in TCDD exposed jobs, OR = 4.0 (1.0–15.4); ('t Mannetje et al., 2018)
TCDD ≥ 10 pg/g, OR = 3.1 (0.9–10.7); ↓ FT4 < 12.8 pmol/l, OR = 4.5 (1.4–14.4)
Seveso, Italy 981 women Serum (median) 1976 55.9a ↔ Diabetes, HR= 0.76 (0.45–1.28); ↑ Metabolic syndrome in women≤ 12 y in 1976, (Warner et al., 2013b)
OR = 2.03 (1.25–3.29), p = 0.01
Nurses' Health Study II 756 diabetes and 766 control Blood PCB-105, 118, 156, 157, and 167 OR = 1.78 (1.14–2.76), p = 0.006 (Zong et al., 2018)
Seveso, Italy 51 mother–child pairs Extrapolated pregnancy (median) 18.9a; 44.8c ↑ b-TSH, p < 0.01 (Baccarelli et al., 2008)
Seveso, Italy (2008–2009) 909 women Serum (median) 1976 60.2a; 1996 7.0a, 25.6c 1996 ↓ TT4, β = − 0.27 (−0.49, −0.05); ↔ TSH, FT3, FT4; 2008 ↓ TT4, β = − 0.11 (Chevrier et al., 2014)
(−0.32, 0.12); ↔ TSH, FT3, FT4 in 2008
Taiwan, China 118 (0 m) and 92 (2 y) mother- Placenta (median) 16.3c Newborns, non-ortho PCBs ↓ FT4 × TSH, r = − 0.2, p < 0.05; 2 y, PCDD/Fs ↑ FT4 (Su et al., 2010; Wang et al., 2005)
newborn pairs × TSH; ↔ 5 y
Immunotoxicity
Japan 239 mother-infant pairs (0 m); 364 Maternal blood (median) during pregnancy Newborns ↓ IgE in boys, −0.87 (−1.68, −0.06); 18 m, PCDFs ↑ Otitis media, (Kishi et al., 2013; Miyashita et al.,
(18 m); 327 (3.5 y); 264 (7 y) 14.0c; 13.9c; 14.2c; 15.0c OR = 2.5 (1.1–5.9); 3.5 y ↓ Frequency of wheezing in boys, OR = 0.03 (0.00–0.94); 7 2011, 2018b)
y ↑ Frequency of wheezing, OR = 7.81 (1.42–42.9)
Netherlands 207 mother-infant pairs Maternal and cord plasma, human milk PCBs Prenatal PCB ↑ T-cell lymphocyte population; ↓ Antibody levels to mumps and (Weisglas-Kuperus et al., 2000,
and PCDD/Fs measles, and shortness of breath with wheeze, chickenpox; Postnatal PCB ↑ Middle- 2004)
ear infections, chicken pox; ↓ Allergic reactions; Postnatal DLCs ↑ Coughing, chest
congestion, and phlegm
Seveso, Italy 676 children of 438 mothers Maternal serum (median) 1976 64.7a; 1976 TCDD ↓ Eczema; ↔ Asthma, hay fever; Pregnancy TCDD↔ Eczema, asthma, and (Ye et al., 2018)
Pregnancy 11.2a hay fever
Japan 2264 participants Blood (median) 16.0c Significant inverse dose–response ↓ Atopic dermatitis, OR = 0.26 (0.08–0.70) (Nakamoto et al., 2013)
Carcinogenic teratogenicity
Seveso, Italy 278,108 residents; 981 women Serum (median) 1976 447.0a vs 94.0a; 71.8a ↑ Lymphatic-hematopoietic cancer; ↑ Breast cancer, HR = 2.1 (1.0–4.6) (Consonni et al., 2008; Warner
vs 51.1a et al., 2002)
French 429 breast cancer cases vs 716 Air (mean) 0.14b1 vs 0.12b1 ↔ Breast cancer (Danjou et al., 2019)
controls
TCDD, apg g−1; PCDD/Fs, bpg WHO-TEQ g−1 lipid, b1μg-TEQ m−2; DLCs, cpg WHO-TEQ g−1 lipid, c1pg TEQ kg−1 bw day−1, c2pg CALUX TEQ g−1 lipid; CALUX, chemical-activated luciferase expression; TT4, total
thyroxine; FT4, free thyroxine; HR, hazard ratio; RR, rate ratio; OR, odd ratio; (95% confidence interval).
Q. Dai, et al. Environment International 139 (2020) 105731
PCDD/F and DL-PCB exposure is associated with numerous adverse and a more advantageous parental and home environment
health outcomes (Table 1). The magnitude of effects varies depending (Vreugdenhil et al., 2002; Patandin et al., 1999). Although DLCs can be
on sex, sources and routes of exposure, and timing of cumulative ex- transferred to infants through breastfeeding, the advantages of breast-
posure at different life stages (Sexton and Hattis, 2007). There is almost feeding far outweigh the potential adverse effects (van den Berg et al.,
no data on the health effects of PBDD/Fs on people possibly due to 2017). Infant DLC exposure should be lowered by reducing maternal
limited evidence for general population exposure. However, high exposure rather than discouraging breastfeeding.
PBDD/F concentrations have been found in environmental matrices Adults accidently exposed to a highly TCDD-contaminated en-
(Xiao et al., 2016; Ma et al., 2009) and human tissues in EWRS (Bruce- vironment also show effects on their peripheral and central nervous
Vanderpuije et al., 2019; Tue et al., 2014), as well as high PCDD/F and systems (Urban et al., 2007; Pelcl et al., 2018), especially the biological
DL-PCB levels. Thus, investigations on DLC exposure and the combined clock and sleep/wake systems. Frequent symptoms of insomnia and
human health effects in EWRS is urgent. lower subjective sleep quality were associated with higher blood con-
centrations of 2,3,4,7,8-pentachlorodibenzofuran (PeCDF) in Yusho
2.1. Neurotoxicity patients (Japan), following ingestion of rice bran oil contaminated with
PCBs. Blood 2,3,4,7,8-PeCDF concentrations were high among patients
In experimental animal studies, exposure to TCDD during the cri- with depression or schizophrenia (Kondo et al., 2018). Some studies
tical period of brain development has been shown to disrupt expression also suggest that workers exposed to any TCDD-contained phenoxyacid
of key developmentally regulated genes (Hill et al., 2003), alter neu- herbicide or chlorophenol have higher rates of psychological disorders,
rogenesis (Williamson et al., 2005), and induce premature senescence including fatigue, headache, and suicide (Pelclova et al., 2009, 2018;
and apoptosis in neuronal cells (Parng et al., 2007; Wan et al., 2014). Vena et al., 1998). However, the Seveso Women’s Health Study (SWHS)
AhR-mediated dioxin neurotoxicity produces sex-specific disruptions in on women who were in their 30s at the time of the explosion does not
behavioral and cognitive functions of learning, memory, emotion and indicate an adverse effect on neuropsychological and physical func-
motor development in mice (Haijima et al., 2010; Nishijo et al., 2007a; tioning except for grip strength twenty and thirty years after exposure,
Seo et al., 1999). The toxins that induce neurotoxicity in zebrafish and possibly since the exposure time was not a critical period of brain de-
rodents may cause similar neurotoxicity in humans (Wan et al., 2014; velopment and the cohort is not yet old enough for age-related sus-
Parng et al., 2007). ceptibility to dioxin neurotoxicity to manifest (Ames et al., 2018a). The
Epidemiological studies indicate that perinatal PCDD/F and DL-PCB heterogeneity in exposure time and methods, and neurological assess-
exposure can interfere with the brain development of fetuses and in- ment measures, limits drawing definite conclusions on health outcomes
fants and adversely affect a variety of neuropsychological functions of neuropsychological function associated with PCDD/F and DL-PCB
which may last into childhood, including cognition, memory, learning, exposure in children and adults.
attention, language, motor function, executive function, and behavior
(Neugebauer et al., 2015; Nishijo et al., 2007b; Pham et al., 2015; Tai 2.2. Reproductive toxicity
et al., 2013, 2016). However, in some studies only weak associations
were found in subgroups (Ames et al., 2019; Caspersen et al., 2016a). The reproductive toxicity of PCDD/Fs and DL/PCBs in humans is
Sex-specific differences in the effect of prenatal PCDD/F and DL-PCB mainly seen as influencing the reproductive hormones (Egeland et al.,
exposure (15.1 pg WHO-TEQ g−1 lipid) on mental and motor devel- 1994), sexual maturation (Burns et al., 2016; Den Hond et al., 2002),
opment were found in some Japanese infants aged 6–18 months. The and fertility (Eskenazi et al., 2010; Hsu et al., 2003). Except for de-
effect may be stronger in male infants and be attenuated or disappear at creased serum testosterone levels (Sweeney et al., 1997; Gupta et al.,
later stages of infancy (Nakajima et al., 2006, 2017; Kishi et al., 2013). 2006; Sun et al., 2020), studies on TCDD’s effects on luteinizing, es-
Some studies demonstrated that poorer cognitive and motor abilities in tradiol and follicle-stimulating hormones of highly exposed cohorts
42-month-old Dutch children were negatively associated with prenatal have yielded inconsistent results. Significant positive associations with
but not postnatal PCDD/F and DL-PCB exposure (Patandin et al., 1999), luteinizing hormone and follicle-stimulating hormone were found in
which indicates a more dominant role of in utero exposure than lacta- chemical plant workers (Egeland et al., 1994), and decreased estradiol
tional exposure (van den Berg et al., 2017). In Norwegian 3–4 year old and increased follicle-stimulating hormone in Seveso males who were
children, higher odds for language delay was significantly associated infants and adolescents at exposure, however, not in adults (Mocarelli
with high maternal dietary PCDD/F and DL-PCB exposure, but only in et al., 2008), nor in residents of dioxin hotspots in the Vietnam war
girls (Caspersen et al., 2016b). The sex-specific associations may occur (Van Luong et al., 2018). The age and dose of exposure may account for
because girls are more vulnerable to endocrine active effects of DLCs in the inconsistent results. TCDD exposure in infancy was negatively re-
terms of language development (Caspersen et al., 2016b). In general, lated to sperm concentration and motility, but an opposite association
boys tend to develop language skills later than girls and catch up before was found with exposure during puberty, which may reflect the dif-
school age (Rice et al., 2008; Koenigsknecht and Friedman, 1976). No ferences in the hormonal regulation of Sertoli cells with age (Mocarelli
associations were found with Attention-Deficit/Hyperactivity Disorder, et al., 2008). Pilsner et al. (2017) found that DLCs may induce epige-
verbal/non-verbal intelligence, or executive functions (Caspersen et al., nomic reprogramming in male germ cells of adults during spermato-
2016b). The methodological limitation of dietary PCDD/F and DL-PCB genesis, causing an increase in the incidence of birth defects in their
exposure measurement from food frequency questionnaires should be offspring. TCDD exposure may affect ovarian function of exposed
taken into consideration. High prenatal PCDD/F and DL-PCB exposure women, which is more sensitive during maturation such as when ex-
is also found to increase autistic traits of 8–10 year-old children (Guo posure occurs before menarche, as evidenced by early onset of meno-
et al., 2018; Nishijo et al., 2014; Nowack et al., 2015), particularly in pause (Eskenazi et al., 2005), prolonged menstrual cycle (Eskenazi
boys (Tran et al., 2016). However, in older Seveso children aged et al., 2002b), endometriosis (Cano-Sancho et al., 2019; Eskenazi et al.,
7–17 years who were born after the industrial accident in 1976, which 2002a; Koninckx et al., 1994; Martínez-Zamora et al., 2015; Mayani
resulted in one of the highest levels of residential TCDD contamination et al., 1997), and infertility (Eskenazi et al., 2010). This may be because
known, no association between in utero TCDD exposure and their TCDD compromises the hypothalamic–pituitary–gonadal regulatory
neuropsychological function was found, except for subtle deficits axis, making women less sensitive to estrogen, as well as decreasing
among sensitive subgroups of boys and children with shorter lactation ovarian weight and the number of follicles (Schmidt, 2017), although
histories (Ames et al., 2019). The evidence supports the view that this was not confirmed in the Seveso cohort (Warner et al., 2007).
neurodevelopmental decrements associated with in utero PCDD/F and Prenatal exposure to relatively low PCDD/F and DL-PCB doses,
DL-PCB exposure may be compensated in later infancy by breastfeeding below those that induce overt maternal toxicity, may affect steroid
6
Q. Dai, et al. Environment International 139 (2020) 105731
hormone levels (e.g. lower estradiol and testosterone) (Boda et al., growth may account.
2018; Miyashita et al., 2018a), behavioral sexual dimorphism (Winneke
et al., 2014) and pubertal onset (e.g. shorter penile length and delayed 2.4. Endocrine toxicity
initiation of breast development) of offspring (Leijs et al., 2008; Guo
et al., 2004), the consequence of which differ between boys and girls, Dioxin-like compounds are known to induce hormone and growth
presumably by interacting with the hypothalamic–pituitary–gonadal dysregulation and therefore act as environmental endocrine disruptors.
axis (Bergman et al., 2013). It suppresses the male secretion of inhibin B When incorporated into the body, they may interfere with the normal
and follicle-stimulating hormone which influences the testicular func- functioning of the endocrine system, which is responsible for regulating
tion, including decreased spermatogenic capacity and sperm quality hormone balance, growth, reproduction, and behavior (Colborn et al.,
(Mocarelli et al., 2011). Lower estradiol levels in 8-year-old Taiwan 1993; Bergman et al., 2013; Damstra et al., 2002). Dioxins can alter the
children and shorter fundi and uteri lengths in girls were also found to metabolic levels of thyroid hormones and insulin, and lower insulin
be associated with in utero background PCDD/F and DL-PCB exposure levels or insulin receptors, resulting in glucose metabolic disorders, and
(14.8 pg WHO-TEQ g−1 lipid) (Su et al., 2012). However, no re- ultimately diabetes (Calvert et al., 1999; Remillard and Bunce, 2002;
productive developmental effects in boys were found, probably because Henriksen et al., 1997; Kern et al., 2004; Lee et al., 2014) and its un-
they were too young at the time of examination to have any signs of derlying conditions such as modified glucose metabolism, insulin re-
reproductive maturation. The evidence indicates that in utero DLC ex- sistance, impaired insulin secretion and metabolic syndrome (Jaacks
posure may impact the reproductive development of offspring. and Staimez, 2015; Kokichi, 2018; Uemura et al., 2009; Warner et al.,
2013b, 2019). The US Department of Veterans Affairs added type 2
2.3. Developmental toxicity diabetes to the list of presumptive diseases associated with exposure to
Agent Orange (Department of Veterans Affairs, 2001).
Prenatal and perinatal period, infancy, childhood, and puberty are Epidemiological and experimental in vivo studies have revealed that
considered to be the most sensitive exposure windows in terms of DLC exposure may induce diabetes through insulin resistance and the
growth and development (Beszterda and Franski, 2018; Rice and destruction of beta cell function (Alonso-Magdalena et al., 2011; Chang
Barone, 2000), during which exposure to PCDD/Fs and DL-PCBs is as- et al., 2010, 2011; Cranmer et al., 2000; Kim et al., 2019; Lee et al.,
sociated with poor birth outcomes (Revich et al., 2001), decreased male 2017; Lim et al., 2010). The most informative human data on this as-
to female birth ratio (Mocarelli et al., 1996, 2000; Pesatori et al., 2003), sociation come from studies of high PCDD/F and DL-PCB exposed co-
and child growth retardation (Guo et al., 2004; Rogan et al., 1988). horts, involving Vietnam War veterans (Henriksen et al., 1997; Kang
Studies on Vietnam veterans and residents exposed to dioxin-containing et al., 2006; Kim et al., 2003; Michalek and Pavuk, 2008), occupa-
Agent Orange reported a higher incidence of poor birth outcomes, such tionally exposed workers in Czech Republic (Pelcl et al., 2018), New
as miscarriage, preterm birth, stillbirth risk (Le and Johansson, 2001), Zealand ('t Mannetje et al., 2018), the United States (Calvert et al.,
congenital birth defects (Kang et al., 2000; Knafl, 2018; Ngo et al., 1999), and the International Agency for Research on Cancer multi-
2006), and developmental enamel defects (Ngoc et al., 2019). Among centric mortality study comprising 36 cohorts from 12 countries (Vena
Japanese severely affected by PCB- and PCDF-contaminated rice oil, the et al., 1998), and residents of contaminated communities in the United
incidences of spontaneous abortion and stillbirth, pregnancy loss, and States (Klein et al., 2011), Taiwan (Chen et al., 2006; Huang et al.,
preterm delivery increased, and the ratio of male to female births de- 2015; Wang et al., 2008) and Seveso (Bertazzi et al., 2001; Consonni
creased, but the associations diminished over time (Yorifuji et al., 2013; et al., 2008). Chronic low-level exposures to DLCs of the general po-
Tsukimori et al., 2008). Higher maternal blood PCDD/F and DL-PCB pulation in Japan (Nakamoto et al., 2013; Uemura et al., 2008), Bel-
levels in Yusho women were associated with lower birth weight, gium (Fierens et al., 2003), and the National Health and Nutrition Ex-
especially in male infants (Tsukimori et al., 2012). The mechanism may amination Survey from the US (Dzierlenga et al., 2019; Everett et al.,
be the genetic variation across the maternal AhR gene that increases the 2007; Lee et al., 2006a, 2006b), the Nurses' Health Study II (Zong et al.,
susceptibility to low birth weight in newborns prenatally exposed to 2018), and a pregnancy cohort in Crete, Greece (Vafeiadi et al., 2017),
DLCs (Ames et al., 2018b). Male infants are more susceptible than fe- are also linked with an increased risk of diabetes (Kokichi, 2018). A
males to growth restriction induced by prenatal DLC exposure systematic review and meta-analysis concluded that there is sufficient
(Tsukimori et al., 2012). The SWHS reported that initial maternal TCDD evidence of a significant association between serum PCDD/F and PCB
levels in 1976 were non-significantly associated with low birthweight concentrations and the risk of type 2 diabetes (Song et al., 2016).
and increased incidents of preterm delivery. However, no association However, Goodman et al. (2015) found that a positive dose-response
between maternal TCDD levels at pregnancy and spontaneous abortions relationship is present only in cross-sectional studies of populations
or gestational age was found (Eskenazi et al., 2003; Wesselink et al., with current TCDD levels less than 10 pg g−1 lipid. The cause-effect
2014). The findings suggest that, in epidemiologic studies of pregnancy relationships and detailed molecular mechanisms between DLC body
outcomes associated with dioxin exposure, the highest dose may be burdens and diabetes remain to be elucidated.
more relevant than pregnancy dose. Prenatal and perinatal PCDD/F and Thyroid hormones play a crucial role in human growth and devel-
DL-PCB exposure is adversely associated with fetal growth and posi- opment, and in the maintenance of normal physiological status. Any
tively associated with infant growth (Papadopoulou et al., 2013; Iszatt deficiency or increase of thyroid hormones during critical periods of
et al., 2016). The association exhibit gender-specific differences in 3- brain development may result in irreversible impairment, morpholo-
year-old children, with a negative effect in boys but a positive effect in gical and cytoarchitecture abnormalities, disorganization, mal-
girls (Wang et al., 2019; Tai et al., 2016). Studies on European birth development, and physical retardation (Ahmed et al., 2008). DLCs in-
cohorts found that perinatal PCDD/F and DL-PCB exposure was asso- teract strongly with transthyretins and thyroid hormone receptors,
ciated with increased body mass index (BMI) and risk of overweight in which accelerate thyroid hormone clearance (McKinney et al., 1985;
7-year-old girls (Iszatt et al., 2016). However, the Seveso Second Gen- Ishihara et al., 2003). In utero and lactational exposure to PCDD/Fs and
eration Study found an inverse association between initial maternal PCBs may have different effects on thyroid hormone levels [thyroid
TCDD levels in 1976 and BMI of their 2–39 year-old daughters (Warner stimulating hormone (TSH), thyroxine (T4), triiodothyronine (T3)] and
et al., 2019). The wide age range of the cohort likely increased the thyroid function in neonates and children (Koopman-Esseboom et al.,
variability of the results. These studies on populations worldwide have 1994; Wang et al., 2005; Baccarelli et al., 2008), and may diminish or
described the inconsistent effects of DLC exposure on birth and growth disappear over time (Su et al., 2010; Pluim et al., 1993). Once again the
outcomes at various ages, for which limited sample size, various ex- thyroid hormone metabolism may be influenced by current background
posure assessments, and lack of a standardized measurement of child exposure to DL-PCBs during adolescence (Leijs et al., 2012). Systematic
7
Q. Dai, et al. Environment International 139 (2020) 105731
reviews of epidemiological studies on the association between perinatal 2.6. Carcinogenicity
TCDD exposure and thyroid function of children showed inconsistent
results, despite some evidence of sub-clinical hypothyroidism found to The International Agency for Research on Cancer classified TCDD,
be induced by dioxin exposure within three months of birth (Giacomini 2,3,4,7,8-PeCDF, PCBs and DL-PCBs as multi-site carcinogens (group 1)
et al., 2006). Epidemiological studies on highly exposed cohorts of (IARC, 1997, 2012, 2016). Overall, human and animal studies indicate
Vietnam Air Force Veterans during 1982–1992 found that serum TCDD that the carcinogenicity of TCDD is for all cancers combined rather than
levels were positively correlated with TSH but not thyroid disease for any specific site (IARC, 1997; Warner et al., 2011; Xu et al., 2016).
(Pavuk et al., 2003). The SWHS found that initial TCDD exposure before Excess risks have been observed in TCDD-exposed subjects for all can-
girls’ menarche may permanently alter total T4 (but not free T4, free T3, cers (Manuwald et al., 2012), such as lymphomas (Viel et al., 2008),
or TSH levels), which tends to be more influential than later TCDD body multiple myeloma (Consonni et al., 2008), soft-tissue sarcoma (Pesatori
burden (Chevrier et al., 2014). However, no associations with T3, T4, et al., 2003), lung (Steenland et al., 1999) and liver (Kogevinas et al.,
and TSH were found in Yusho patients (Nagayama et al., 2001). Be- 1997) cancers. Mortality and morbidity findings of the Seveso popula-
tween-studies discrepancies in thyroid levels may not only reflect tion exposed to TCDD showed increased risk from lympho-hemopoietic
complex time-specific actions of mixed DLC exposure, but also differ- neoplasms, digestive system, and respiratory system cancers, and also
ences in laboratory methodology, age and sample size (Giacomini et al., thyroid gland cancer (Pesatori et al., 2003). An increased prevalence of
2006; Khoa et al., 2015). The toxicological mechanism of the effect of reproductive cancers (breast, cervical, testicular, endometrial cancer)
dioxin exposure on the thyroid gland needs to be further explored. has been found in high TCDD exposed cohorts (Revich et al., 2001;
Warner et al., 2002; Kogevinas, 2001). However, in a recent French
case-control study, no increased risk of breast cancer was found for
2.5. Immunotoxicity higher airborne PCDD/F levels but an increased risk in estrogen re-
ceptor-positive breast cancer with low PCDD/F exposure was observed
The immune system is one of the most sensitive targets for dioxins. (Danjou et al., 2019). The findings might be consistent with a non-
The immune response of mammals exposed to DLCs during their ma- monotonic dose-response effect of dioxins on breast cancer. The cancer-
turation stage is more severe and persistent than that caused by adult induced effects found in high DLC-exposed cohorts are still inconsistent,
exposure (Nowak et al., 2019; Dietert, 2009). A broad range of im- for which specific study limitations (e.g. sample sizes and adjustments
munotoxicity of DLCs has been observed in animal studies (Gehrs et al., for confounders) and the differences in study population and metho-
1997; O'Driscoll et al., 2019; Vorderstrasse et al., 2003), including dology (e.g. exposure assessment) may partially explain (Xu et al.,
thymic atrophy, immunosuppression of both humoral and cell-medi- 2016). The impact of temporal variation in exposure should be taken
ated immune responses (Fine et al., 1989; Kerkvliet, 2002; Vos et al., into consideration (Rodgers et al., 2018).
1973; Camacho et al., 2004). In humans, prenatal background exposure Since DLCs are not directly genotoxic, their carcinogenicity is likely
to DLCs may cause postnatal immune dysfunctions and increased sus- mediated via activation of the AhR leading to tumor promotion. The
ceptibility to infectious and allergic diseases in infancy which persist tumor promoting action of TCDD is reflected in facilitating the growth
into childhood (Miyashita et al., 2011; Stølevik et al., 2013; Guo et al., rate and clonal expansion of initiated cells (Baccarelli et al., 2006;
2004). Common infectious diseases acquired in early life may affect the Knerr and Schrenk, 2006), inhibiting the apoptosis of tumor precursor
maturation of the immune system and may thereby prevent the de- cells, and increasing oxidative stress (Schwarz and Appel, 2005). The
velopment of allergic diseases (Weisglas-Kuperus et al., 2004). Breast- dose-response relationship and the mechanisms of action between DLC
feeding for 4 months or more can counteract the negative effect of exposure and cancer needs to be clarified, of which inter-individual
perinatal exposure and prevent children from developing allergic and variability and susceptibility in response are key determinants (Guyton
infectious diseases (Pluim et al., 1994; Ten et al., 2003; Ip et al., 2007; et al., 2009). Careful consideration should be taken when formulating
van den Berg et al., 2017). The findings of the Seveso Second Genera- the maximum tolerated dose of long-term environmental DLC exposure
tion Health Study (Ye et al., 2018), the Hokkaido Study on Environment and the cancer potency threshold for TCDD.
and Children's Health (Kishi et al., 2013; Miyashita et al., 2011, 2018b),
and a Dutch birth cohort study (Weisglas-Kuperus et al., 1995, 2000, 3. Environmental levels and human exposure in EWRS
2004) support the conclusion that prenatal TCDD exposure may in-
crease the development of infectious diseases and decrease allergic E-waste recycling is a significant source of DLC contamination in
conditions. China and some other developing countries (Lei et al., 2020; Prithiviraj
Adult exposure to dioxins also adversely affects the immune system et al., 2019). High levels of DLCs were detected in environmental (air,
function, altering T-lymphocyte subpopulations and immunoglobulins particulate, soil, dust, and sediment) and biological (hair, blood, breast
levels (Kim et al., 2018), and increasing susceptibility to infectious milk, placenta, and meconium) samples in well-known EWRS (Chan
diseases (Zober et al., 1994). Increased TCDD levels of Seveso in- and Wong, 2013). Concentrations of PCDD/Fs in air and soils near e-
habitants were found to be inversely associated with only plasma IgG waste recycling operations were greater than those reported around
but not with IgM, IgA, C3, or C4 levels, which indicates that TCDD may municipal solid waste incinerators and chemical industries (Ma et al.,
play a role in suppressing antibody production (Baccarelli et al., 2002, 2008). The PCDD/F and DL-PCB concentrations in the blood and breast
2004). A large sample-size cross-sectional study on background dioxin milk samples from EWRS were higher or comparable to the levels re-
exposure of the general Japanese residents found that low-level DLC ported to be associated with observable adverse effects in epidemiolo-
exposure (16.0 pg WHO-TEQ g−1 lipid) is associated with reduced risk gical studies. It suggests that the health risk of DLC exposure on local
of atopic dermatitis (Nakamoto et al., 2013). However, no immune workers and residents in EWRS should be assessed in a timely manner.
response (immunoglobulins, autoantibodies, and lymphocyte subsets)
was found in 83 adult Yusho patients with relatively high blood PCDD/ 3.1. Atmosphere
F and DL-PCB levels, possibly due to the rather small-scale sample size
(Nagayama et al., 2001). Therefore, in order to obtain more conclusive Dioxin-like compounds can be generated when incineration tem-
results concerning the effects of DLCs on immune response systems, peratures are not high enough and will then be released into the at-
further uniform large-scale investigations with standard exposure as- mosphere (Shibamoto et al., 2007; Zhang et al., 2017a; Zhang et al.,
sessment methods are needed. 2016a). The atmospheric particulates act as an important vector for the
transport of DLCs in the environment from the emission sources to other
environmental matrices and may be the largest exposure route for e-
8
Q. Dai, et al. Environment International 139 (2020) 105731
waste recycling workers (Ma et al., 2008; Qin et al., 2019). However, 3.2. Soil
possibly due to the methodological complexity of air sampling and
analysis, data on DLC concentrations in the air of EWRS is scarce, and Levels of DLCs in soils of EWRS and nearby agricultural lands have
virtually absent in developing countries except China (Table 2). Air been examined in many studies (Table 3). As expected, relatively high
samples taken near Guiyu, the “World Electronic Waste Terminal,” in PCDD/F levels were found in the soils from an open e-waste burning
2005 contained the highest DLC level, with PCDD/F concentrations of site [2.1 × 104 pg g−1 dry weight (dw)] and an acid leaching site
64.9–2765.0 (mean 891.4) pg m−3 and 2,3,7,8-PBDD/F concentrations (3.9 × 104 pg g−1 dw) in Guiyu (Leung et al., 2007), which were
of 8.12–461.0 (mean 118.0) pg m−3 (Li et al., 2007). These are 12–18 comparable to those reported for EWRS in Qingyuan (2.1 × 104 pg g−1
times higher than in Chendian, a non-EWRS only 9 km away from dw) (Hu et al., 2013). The soil TEQ levels of PCDD/Fs in Guiyu were
Guiyu, and 37–133 times higher than in Guangzhou, the nearest major comparable or even higher than that of the former storage of Agent
city. Since the atmospheric PCDD/F level in Chendian is still much Orange in A-So airbase collected in 1996 (nearly 26 years after last
higher than that in other urban areas, the data indicate that the severe exposure), which is hotspot during the Vietnam War (112.6 pg TEQ g−1
dioxin pollution in Guiyu may affect the nearby Chendian. Fortunately, dw) (Dwernychuk et al., 2002). Open burning and dismantling of e-
due to the enhanced law enforcement and environmental regulations on waste and acid leaching activities are identified as the major sources of
centralized e-waste dismantling with advanced technology by the au- high PCDD/F concentrations (Leung et al., 2007; Tue et al., 2019). The
thorities of Guangdong province and local governments in 2010, air mean soil PCDD/F concentration of household e-waste facilities in
quality in Guiyu has improved greatly. The mean atmospheric TEQ Fengjiang, Taizhou (4.3 × 103 pg g−1 dw) (Ma et al., 2008), was one
concentrations of PCDD/Fs in 2013 was seven times lower than that in order of magnitude greater than that of a nearby agricultural field
2005 (Xiao et al., 2012; Zhang et al., 2017b). Similarly, in 2013, at- (5.0 × 102 pg g−1 dw) (Shen et al., 2009), and a chemical industrial
mospheric PCDD/F concentrations in Qingyuan (Guangdong, China), complex in Wujing, Shanghai (2.9 × 102 pg g−1 dw) (Ma et al., 2008).
famous for its e-waste recycling industry, significantly decreased by 31 Soils from informal backyard EWRS in Bangalore (India) where crude
times compared with the values measured in 2010, and by 49 times processes, like heating, acid treatment, and open burning, were em-
than in 2006, possibly due to implementation in recent years of laws ployed, also contained high levels of PCDD/Fs, DL-PCBs, and PBDD/Fs,
banning open burning of waste wires, acid washing, and other un- with mean concentrations of 3.6 × 104, 1.3 × 104, and 1.1 × 105 pg
controlled e-waste recycling activities (Ren et al., 2015; Xiao et al., g−1 dw, respectively (Karri et al., 2008). These concentrations were
2014; Zhang et al., 2017b). Consistent results were also observed in much higher than those in soils of an Indian e-waste facility where
Taizhou (Zhejiang, China), another large e-waste recycling region with recycling of e-waste did not involve any thermal processes, with mean
a nearly 40-year history of informal e-waste disposal and recycling. The PCDD/F, DL-PCB, and PBDD/F concentrations of 6.4 × 102, 6.5 × 103,
average atmospheric PCDD/F concentration (0.4 pg TEQ m−3) in 2010 and 2.9 × 104 pg g−1 dw, respectively (Karri et al., 2008). However, in
was 62% lower than that in 2005 (1.1 pg TEQ m−3) (Zhang et al., India, 95% of the e-waste is informally treated and crude processed
2012b). This may have benefited when the local government started to (Raghupathy et al., 2010). With the awareness about e-waste among
standardize the dismantling processes in 2005. Small-sized informal government and general public, in 2016, the Indian Government re-
workshops were urged to close and replaced with formal e-waste dis- leased a more stringent and powerful legislative management frame-
mantling centers. Sophisticated recycling technologies were en- work, where e-waste is collected by municipalities, authorized retailers
couraged to minimize open burning activities (Fu et al., 2012). These or commercial pick-up services and then sent to different centers for
data indicate that the strengthened regulations and centralized dis- further processing. The whole process is controlled by the Central
mantling measures in EWRS showed a significant reduction in the re- Pollution Control Board to ensure the collection and disposal of e-waste
lease of PCDD/Fs. However, the atmospheric PCDD/F concentrations in a sustainable manner according to the principle of Producer Re-
are still much higher than reference areas and municipal solid waste sponsibility. Such rules are expected to control the release of hazardous
incinerators and heavily industrialized areas. substances in the e-waste disposal stream (Government of India, 2016;
Table 2
Mean concentrations of atmospheric 2,3,7,8-substituted PCDD/Fs and PBDD/Fs at typical EWRS and some other locations.
City/Country Site PCDD/Fs (pg m−3) PBDD/Fs (pg m−3) Type Time Reference
Guiyu, China E-waste open-burning sites 6.5 (0.7a) Gas + TSP 2004 (Wong et al., 2007)
Guiyu, China EWRS, summer/winter 127.3/253.0 (8.8b/16.0b) 21.9/118.0 (4.5b/26.9b) Gas + TSP 2005 (Li et al., 2007)
Guiyu, China Circuit boards recycling sites 317 (14.5a) 481 (91.3a) TSP 2007 (Ren et al., 2014)
Guiyu, China Circuit board recycling sites 17.6 (0.8b) Gas + TSP 2013 (Zhang et al., 2017b)
Guiyu, China Plastic recycling workshops 31.1 (1.7b) Gas + TSP 2013 (Zhang et al., 2017b)
Guiyu, China Remote small villages 1.1 (0.02b) Gas + TSP 2013 (Zhang et al., 2017b)
Chendian, China Underwear industry 22.9 (1.2a, 1.4b) 3.6 (0.9b) Gas + TSP 2005 (Li et al., 2007)
Guangzhou, China Urban 6.7 (0.3a, 0.4b) 1.0 (0.2b) Gas + TSP 2005 (Li et al., 2007)
Taizhou, China EWRS 14.3 (1.1b) Gas + TSP 2005 (Li et al., 2008)
Taizhou, China EWRS, summer/winter 84.5/212.2 (4.9a/12.7a) 5.4/17.6 (0.8a/2.4a) Gas + TSP 2006–2007 (Zhou, 2011)
Taizhou, China EWRS, summer/winter 32.8/32.9 (0.5a/0.4a) 58.9/73.7 (0.2a/0.2a) Gas + TSP 2010 (Zhang et al., 2012b)
Hangzhou, China Solid waste incinerator 31.9 (0.5b) Gas + TSP 2007–2008 (Xu et al., 2009)
Qingyuan, China Courtyard open burning sites 237.0 (13.6a) Gas + TSP 2006 (Ren et al., 2015)
Qingyuan, China Courtyard open burning sites 159.4 (8.0a/8.5b) Gas + TSP 2009–2010 (Xiao et al., 2014)
Qingyuan, China A green recycling center 34 (0.6b) Gas + TSP 2013 (Zhang et al., 2017b)
Qingyuan, China Neighborhoods near EWRS 5.0 (0.3b) Gas + TSP 2013 (Zhang et al., 2017b)
Qingyuan, China Remote small villages 2.9 (0.1b) Gas + TSP 2013 (Zhang et al., 2017b)
Chennai, India EWRS (1.65c) Gas (Chakraborty et al., 2016a)
Korea Solid waste incinerators (0.2b) Gas + TSP (Kim et al., 2008)
Duisburg, Germany Heavily industrialized areas (0.3b, 0.1b) Gas + TSP TSP 1987, 1993 (Hiester et al.,1997)
TEQs are given in parentheses.
a pg WHO-TEQ m−3.
b pg I-TEQ m−3.
c pg CALUX-TEQ m−3; TSP, total suspended particulate.
9
Q. Dai, et al. Environment International 139 (2020) 105731
10
Table 3
Concentrations of soil/dust/sediment DLCs at typical EWRS and some other locations.
Country City Site Data type (sampling year) PCDD/Fs DL-PCBs PBDD/Fs Reference
Soil (pg g−1 dw)
China Guiyu Open-burning sites Range (2004) 3.1 × 104–9.7 × 105 2.2 × 104–4.5 × 105 (7.0a–878.0a) (Wong et al., 2007; Yu et al.,
(627.0b–13,900.0b) 2008)
China Guiyu Combusted residue Mean (2004) 2.1 × 104 (129.0a) (Leung et al., 2007)
China Guiyu Acid leaching area Mean (2004) 3.9 × 104 (506.0a) (Leung et al., 2007)
China Guiyu Open burning and acid Mean (2009) 9.7b (Xu et al., 2017)
leaching sites
China Guiyu E-waste roasting sites Mean (2009) 8.4b (Xu et al., 2017)
China Guiyu Manual disassembly and Mean (2009) 5.4b (Xu et al., 2017)
shredding sites
China Taizhou Large EWRS Mean (2007) 1.0 × 103 (49.3a) 1.1 × 105 (799.0) (Ma et al., 2008, 2009)
China Taizhou Household EWRS Mean (2007) 1.5 × 103 (92.0a) (Ma et al., 2008)
China Wujing Chemical industrial complex Mean (2007) 1.6 × 102 (5.4a) 57.8 (0.5) (Ma et al., 2008, 2009)
China Taizhou Farmland near EWRS Mean (2006) 5.0 × 102 (12.2a) 2.1 × 104 (6.7a) (Shen et al., 2009)
China Qingyuan EWRS Mean (2009) 2.9 × 102 (68.4a) (Xiao et al., 2016)
China Qingyuan EWRS Mean (2009) 2.1 × 104 (43.0a) (Hu et al., 2013)
China Qingyuan Farmland near EWRS Mean (2006) (7.1a) (Ren et al., 2015)
China East Disassembly industrial park Mean 6.9 × 102 (7.2a) 6.0 × 103 (Liu and Liu, 2009)
China Yangtze River Delta Farmland in EWRS Mean (2004) 2.6 × 103 (21.0a) (Luo et al., 2005)
India Bangalore and Chennai Backyard recycling site Mean (2006) 3.6 × 104 1.3 × 104 1.1 × 105 (Karri et al., 2008)
India Bangalore and Chennai E-waste facility Mean (2006) 6.4 × 102 6.5 × 103 2.9 × 104 (Karri et al., 2008)
India New Delhi, Kolkata, Mumbai EWRS Mean (2014) 5.0 × 103 (31.0a) 4.6 × 104 (39.0a) (Chakraborty et al., 2018)
and Chennai
India New Delhi, Kolkata, Mumbai Open dump Mean (2014) 2.5 × 103 (16.0a) 1.8 × 103 (11.5a) (Chakraborty et al., 2018)
and Chennai
Ghana Agbogbloshie Open burning areas Median (2010) 2.8 × 105 4.2 × 104 9.3 × 105 (Tue et al., 2016)
Ghana Agbogbloshie Non-burning areas Median (2010) 3.8 × 103 1.9 × 103 7.1 × 103 (Tue et al., 2016)
Ghana Agbogbloshie Open burning areas Median (2013) 3.3 × 104 (360.0a) 1.8 × 105 (870.0a) (Tue et al., 2019)
Ghana Agbogbloshie Dismantling area Median (2013) 1.6 × 104 (270.0a) 4.5 × 105 (Tue et al., 2019)
(2200.0a)
Vietnam Bui Dau Open burning sites Median (2012) (77.0a) (4.8a) (14.0a) (Suzuki et al., 2016)
Vietnam Bui Dau E-waste-processing workshop Median (2012) (4.5a) (1.3a) (20.0a) (Suzuki et al., 2016)
Vietnam Bui Dau Farmland near EWRS Median (2012) (1.2a) (0.3a) (0.3a) (Suzuki et al., 2016)
Vietnam A-So Storage for herbicides during (1996 ~ 26 y) Mean (112.6b) 8.5 × 102 (91.1a) (Le et al., 2019; Dwernychuk
Vietnam War (2014 ~ 44 y) et al., 2002)
Dust (pg g−1 dw)
China Guiyu Workshop floor, EWRS 2004 1.3 × 103 (76.2a) (Leung et al., 2011)
China Guiyu Schoolyard near EWRS 2004 1.3 × 103 (103.0a) (Leung et al., 2011)
China Guiyu Street-lined with workshops 2004 6.5 × 102 (53.1a) (Leung et al., 2011)
China Guiyu Workshop, EWRS Mean (2013) (2.7 × 103) (Zhang et al., 2017b)
China Qingyuan Workshop, EWRS Range (2013) (49.0–446.0) (Zhang et al., 2017b)
China Qingyuan Workshop, EWRS Mean (2009) 5.7 × 104 (380.0a) 2.0 × 103 (400.0a) (Hu et al., 2013; Xiao et al.,
2016)
China Taizhou Floor, EWRS Mean (2007) 1.3 × 104 (1070.0a) 1.2 × 105 (Ma et al., 2008, 2009)
(1480.0a)
China Taizhou Floor, EWRS Mean (2006) 2.6 × 104 (724.1a) 26,200.0a (Wen et al., 2008)
India Chennai Metal recovery workshop Mean (2014) 5.8 × 104 (1000.0a) (Chakraborty et al., 2016b)
India Chennai Dismantling and shredding Mean (2014) 1.3 × 103 (20.0a) (Chakraborty et al., 2016b)
workshop
Vietnam Bui Dau Furniture and fan blades Median (2008) 1.4 × 103 1.0 × 103 4.9 × 104 (Tue et al., 2010)
surface, EWRS
Vietnam Trang Minh Furniture and fan blades Median (2008) 2.4 × 103 2.2 × 103 2.3 × 104 (Tue et al., 2010)
surface, EWRS
Sediment (pg g−1 dw)
(continued on next page)
Q. Dai, et al. Environment International 139 (2020) 105731
Chakraborty et al., 2018). In Agbogbloshie, Ghana, concentrations of
PCDD/Fs and PBDD/Fs found in soils from open burning (3.3 × 104
and 1.8 × 105 pg g−1 dw, respectively) and dismantling areas
(1.6 × 104 and 4.5 × 105 pg g−1 dw, respectively) in 2013 were lower
than in soil samples collected in 2010 (2.8 × 105 and 9.3 × 105 pg g−1
dw, respectively) and were comparable to the highest reported for in-
formal EWRS (Tue et al., 2019). These data suggest that low-tech e-
waste recycling facilities have been an important emission source for
DLCs, and have contaminated the nearby agricultural environment
(Shen et al., 2008).
3.3. Dust and sediment
Dust is a complex mixture of particulate matter and acts as a re-
pository of environmental pollutants which represents average levels of
contamination over long periods of time (Butte et al., 2002). Dust in-
gestion is an important pathway for human DLC exposure in EWRS (Tue
et al., 2014). However, investigations of DLC levels in dust and sedi-
ment of EWRS are so far scarce (Table 3). The PCDD/F concentrations
in dust from an e-waste recycling workshop in Guiyu (1.3 × 103 pg g−1
dw) were approximately 13 times greater than the concentrations found
in the dust from a reference site (96.1 pg g−1 dw) (Leung et al., 2011).
Zhang et al., 2017b detected even higher TEQ concentrations of PCDD/
Fs in indoor dust in Guiyu (1020.0–3637.0, mean 2662.0 pg TEQ g−1
dw). These were 59 times higher than the reference site (45.4 pg TEQ
g−1 dw) and 10 times higher than that from Qingyuan (49.0–446.0 pg
TEQ g−1 dw). The PCDD/F concentrations of dust samples collected
from Taizhou e-waste workshop floor were 17–28 times higher than
that of dust on the surface of furniture and fan blades from Vietnamese
EWRS (Ma et al., 2008; Tue et al., 2010). The average TEQ con-
centrations of PCDD/Fs and PBDD/Fs in Taizhou (1070.0 and
1480.0 pg WHO-TEQ g−1 dw, respectively) were three times higher
than that collected from large-scale e-waste facilities in Qingyuan
(380.0 and 400.0 pg WHO-TEQ g−1 dw, respectively) (Ma et al., 2008,
2009; Hu et al., 2013; Xiao et al., 2016). Dust from informal e-waste
recycling workshops engaged in metal recovery in Chennai, India,
showed a mean DL-PCB concentration of 5.8 × 104 pg g−1 dw, much
higher than that from shredding and dismantling sites (1.3 × 103 pg
g−1 dw) (Chakraborty et al., 2016b). This might have resulted from
recycling activities like burning electric wires to recover copper,
heating printed circuit boards to recover lead-tin solder and integrated
components, and using acid chemical strippers to recover gold and
other metals (Chakraborty et al., 2016b). These data show that high
PCDD/F values of dust samples derived from primitive crude e-waste
recycling activities may ultimately transfer to the indoor environments
in these EWRS.
Sediments have a large adsorption capacity which makes them
major reservoirs and sinks for chemicals. DLCs prefer adsorption to
sediments because of their hydrophobic properties (Evenset et al.,
2007). The sediment PCDD/F concentrations in Lianjiang river of Guiyu
ranged from 21.2 to 35,200.0 pg WHO-TEQ g−1 dw, which were no-
ticeably higher than that in locations downstream from Guiyu
(Luksemburg et al., 2002), and in Korle lagoon adjacent to Agbog-
bloshie e-waste disposal and burning areas (988.0 pg WHO-TEQ g−1
dw) (Brigden et al., 2008). For sediments collected from an EWRS of
Qingyuan, the mean TEQ concentrations of PCDD/Fs and PBDD/Fs
were 128.0 and 146.0 pg WHO-TEQ g−1 dw, respectively (Hu et al.,
2013), higher than that in sediments of an EWRS in Bui Dau (7.3 and
4.4 pg WHO-TEQ g−1 dw, respectively) (Suzuki et al., 2016).
3.4. Human body burdens
In recent years, DLC concentrations in human matrices, such as
human hair, breast milk and blood, have been extensively measured to
assess individual exposure (Table 4) and the health effects (Table 1).
Hair is one of the best mediums to reflect short-term contaminant
11
Table 3 (continued)
Country City Site Data type (sampling year) PCDD/Fs DL-PCBs PBDD/Fs Reference
China Guiyu Ash-dumped riverbank (3.5 × 104a) (Luksemburg et al., 2002)
China Guiyu Residential areas near the (2.7 × 103a) (Luksemburg et al., 2002)
river
China Qingyuan Pond in EWRS Mean (2009) 3.4 × 104 (128.0a) 5.3 × 102 (146.0a) (Hu et al., 2013; Xiao et al.,
2016)
China East Disassembly industrial park Mean 643.0 (8.9a) 1.3 × 104 (Liu and Liu, 2009)
Vietnam Bui Dau E-waste-processing area Median (2012) 1.2 × 103 (7.3a) (0.9a) 1.7 × 103 (4.4a) (Suzuki et al., 2016)
Ghana Agbogbloshie Lagoon near EWRS (988.0) (Brigden et al., 2008)
Vietnam A-So Hotspots in Vietnam War (1996 ~ 26 y) Mean (7.8b) 2.8 × 103 (22.0a − 385a) (Le et al., 2019; Dwernychuk
(2014 ~ 44 y) et al., 2002)
TEQs are given in parentheses; apg WHO-TEQ g−1 dw; bpg I-TEQ g−1 dw.
Q. Dai, et al. Environment International 139 (2020) 105731
exposure derived from direct atmospheric deposition (Covaci et al., Breast milk PCDD/F levels were reported to be associated with physical
2002; Nakao et al., 2005). Mean PCDD/F concentrations in the hair of growth of 3-year-old children (Wang et al., 2019). The placenta PCDD/
residents from EWRS of China, such as Guiyu, Taizhou and Fengjiang, F and DL-PCB concentrations in Taizhou (31.2 pg WHO-TEQ g−1 lipid)
have been measured at 21.0, 36.1 and 33.8 pg WHO-TEQ g−1 dw, re- were much higher than that in Taiwan (14.8 pg WHO-TEQ g−1 lipid)
spectively (Chan et al., 2007; Luksemburg et al., 2002; Ma et al., 2011). which have been reported to be associated with growth, hormone levels
These levels are markedly higher than hair samples of municipal solid and reproductive development (Su et al., 2012). The blood PCDD/F and
waste incineration workers (Liu et al., 2019b) and the general popu- DL-PCB levels of children in Taizhou (22.2 pg WHO-TEQ g−1 lipid)
lation from Shanghai (2.2 pg WHO-TEQ g−1 dw) and Hangzhou (5.6 pg were comparable to that of Russian boys (21.1 pg WHO-TEQ g−1 lipid)
WHO-TEQ g−1 dw) (Li et al., 2007). The average PCDD/F concentra- which were reported to be associated with delayed pubertal onset and
tion of maternal breast milk (21.0 pg WHO-TEQ g−1 lipid) in Taizhou is sexual maturity (Shen et al., 2010; Burns et al., 2016). Maternal blood
higher than that of Vietnam residents living near Da Nang Air Base PCDD/F and DL-PCB concentrations of primiparous Ghanaians living
(12.7 pg WHO-TEQ g−1 lipid) which has been proved to be adversely near EWRS (3.1 pg WHO-TEQ g−1 lipid) were lower than that of the
associated with the growth and neurodevelopment of children (Chan Sapporo cohort of the Hokkaido Study on Environment and Children’s
et al., 2007; Pham et al., 2015; Tai et al., 2016; Tran et al., 2016). Health in Japan (14.9 pg WHO-TEQ g−1 lipid) which were known to be
Compared with maternal breast milk samples collected in 2005 (21.0 pg related with lower birth weight, immune function and neurodevelop-
WHO-TEQ g−1 lipid), the PCDD/F concentrations in breast milk were ment of their infants (Bruce-Vanderpuije et al., 2019; Kishi et al., 2013).
notably lower in 2015 (12.9 pg WHO-TEQ g−1 lipid) (Wang et al., However, limited information was found on maternal blood DLC levels
2019), but were still much higher than the average level of the Chinese in other EWRS.
general population (4.9 pg WHO-TEQ g−1 lipid) (Zhang et al., 2016b). The blood PCDD/F and DL-PCB concentrations of these high (New
Table 4
Body burdens of PCDD/Fs and DL-PCBs at typical EWRS and some other locations.
Country City Study population Data type PCDD/Fs DL-PCBs Sampling year Reference
Hair (pg WHO-TEQ g−1 dw)
China Guiyu Residents living near EWRS Mean 21.0 (Luksemburg et al., 2002)
China Taizhou Mothers living near EWRS Mean (2005) 33.8 2005 (Chan et al., 2007)
China Taizhou E-waste workers Mean (2006) 42.4 41.5 2006 (Wen et al., 2008)
China Taizhou E-waste workers Mean (2007) 36.1 2007 (Ma et al., 2011)
China South Waste incineration workers Median 1.7 (Liu et al., 2019b)
China Hangzhou General mothers Mean 5.6 2005 (Chan et al., 2007)
China Shanghai General population Mean 2.2 2007 (Ma et al., 2011)
Cord blood (pg WHO-TEQ g−1 lipid)
China Guiyu Neonates born near EWRS Median 3.9 16.0 2007 (Liu et al., 2019a)
China Taizhou Neonates born near EWRS Median 18.0 2002 (Zhao et al., 2007)
China Taiwan General neonates Mean 5.4 1.3 2000–2001 (Wang et al., 2004)
Vietnam Bien Hoa Neonates born near airbase Mean 14.5 2012 (Boda et al., 2018)
Blood (pg WHO-TEQ g−1 lipid)
Ghana Agbogbloshie E-waste workers Median 6.2 2011 (Wittsiepe et al., 2015)
Ghana Agbogbloshie Mothers living near EWRS Mean 3.8 1.5 2017 (Bruce-Vanderpuije et al.,
2019)
China Taizhou Children living near EWRS Mean 10.3 11.9 2008 (Shen et al., 2010)
China Taizhou Children living near EWRS Mean 8.4 (Xu et al., 2014)
Japan Sapporo General mothers Mean 10 4.9 2002–2005 (Konishi et al., 2009)
China Taiwan General mothers Mean 9.1 4.5 2000–2001 (Wang et al., 2004)
Russia Chapaevsk Boys living near chemical industries Median DLCs 21.1 2003–2005 (Burns et al., 2016)
America Ranch Hand veterans Mean (40.8a) 1982 ~ 15 y (Michalek and Tripathi, 1999)
America Ranch Hand veterans Median PCDDs 18.7 (5.0a) 3.4 2002 ~ 33 y (Pavuk et al., 2014)
PCDFs
Vietnam Bien Hoa Men living near airbase Mean 34.0 (7.3a) 3.3 2014 (Van Luong et al., 2018)
Germany Duisburg Women living in industrialized area Median 15.3 10.8 2000–2003 (Wittsiepe et al., 2007)
New Zealand Herbicide production plant workers Mean (109.0a) 1987 ~ 0 y ('t Mannetje et al., 2016)
New Zealand Herbicide production plant workers Mean 30.0 (19.1a) 3.7 2007 ~ 20 y ('t Mannetje et al., 2016)
Italy Seveso Women experienced chemical Median (105.0a) 1976 ~ 0 y (Warner et al., 2013a)
explosion
Italy Seveso Women experienced chemical Median DLCs 26.2 (7.3a) 1996 ~ 20 y (Warner et al., 2013a)
explosion
Breast milk (pg WHO-TEQ g−1 lipid)
China Guiyu Mothers living near EWRS Mean 0.9 2006 (Xing et al., 2009)
China Taizhou Mothers living near EWRS Mean 21.0 2005 (Chan et al., 2007)
China Taizhou Mother living near EWRS Mean 12.9 2015 (Wang et al., 2019)
China Hangzhou General mothers Mean 9.4 2005 (Chan et al., 2007)
Vietnam Bui Dau Mothers living near EWRS Range PCDFs 13.0a–15.0a 6600.0a–7600.0a 2008 (Tue et al., 2014)
China Taiwan General mothers Mean 7.6 5.2 2000–2001 (Wang et al., 2004)
Vietnam Bien Hoa Mothers living near airbase Mean 11.3 2012 (Boda et al., 2018)
Spain Tarragon Mothers living near incinerators 7.6 2007 (Schuhmacher et al., 2009)
Germany Duisburg Mothers living in industrialized area Median 23.9 24.5 2000–2003 (Wittsiepe et al., 2007)
Placenta (pg WHO-TEQ g−1 lipid)
China Taizhou Mothers living near EWRS Mean 31.2 2005 (Chan et al., 2007)
China Hangzhou General mothers Mean 11.9 2005 (Chan et al., 2007)
China Taiwan General mothers Mean 10.3 2.9 2000–2001 (Wang et al., 2004)
Meconium (pg WHO-TEQ g−1 dw)
China Taizhou Neonates born near EWRS Median 0.8 2002 (Zhao et al., 2007)
a pg g−1 lipid; TCDD concentrations are given in parentheses.
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Q. Dai, et al. Environment International 139 (2020) 105731
Zealand, Seveso, Vietnam, Ranch Hand veterans) and background on the total PCDD/Fs and PBDD/Fs exposures via different pathways in
(Japan, Taiwan) exposed cohorts ranged from 13.6 to 37.8 pg WHO- EWRS (Chan and Wong, 2013).
TEQ g−1 lipid, which were reported to be associated with numerous Total daily intakes of PCDD/Fs through four exposure ways in EWRS
observable adverse health effects. The average blood PCDD/F and DL- of China are much higher than that in waste incinerators of
PCB levels of children in Taizhou were in the middle range. Although Wilrijk, Belgium (Chan and Wong, 2013; Nouwen et al., 2001). Soil
there is a lack of comprehensive data on the DLC body burden of people ingestion and dermal absorption of PCDD/Fs and PBDD/Fs in EWRS of
in EWRS, considering the high environmental levels in EWRS, we can China are comparable to that of Ghana (Hu et al., 2013; Ma et al., 2008,
infer that they are at high health risks, especially in the case of children 2009; Tue et al., 2019; Xiao et al., 2016). PCDD/F exposures through
and infants. Health effects related to DLC exposure of people in EWRS soil ingestion and dermal absorption in India and Vietnam are much
have to our knowledge been rarely studied. Meanwhile, information on lower than that in EWRS of China and Ghana, comparable to that in
human body burdens of PBDD/Fs and PXDD/Fs in EWRS is notably open dumping sites (Chakraborty et al., 2018; Suzuki et al., 2016; Minh
scarce, and is currently available only for small numbers of maternal et al., 2003). Although there is a lack of data on air and food DLC
serum samples of women living near Agbogbloshie EWRS (0.49 and concentrations in EWRS in other countries, according to the data from
0.50 pg WHO-TEQ g−1 lipid, respectively) (Bruce-Vanderpuije et al., EWRS in China, we can infer that the total daily intake of dioxins in
2019) and breast milk samples of women living in Vietnamese EWRS EWRS are comparable or much higher than the contaminated areas of
(Tue et al., 2014). The data suggest that e-waste recycling activities open dumping and waste incineration sites.
contribute to high DLC body burdens of people in EWRS and may put Food consumption is the main route for the general population,
them at high health risk. accounting for approximately 90% of the tolerable daily intake (WHO,
2016). However, this estimate is not appropriate for workers involved
in e-waste recycling and residents living near EWRS (Zhang et al.,
3.5. Exposure and health risk assessment 2017b). They are chronically highly-exposed via inhalation and dermal
absorption of DLC-contaminated vapors, fumes, fly ashes and dust
Human exposures to PCDD/Fs and PBDD/Fs in EWRS through air (Chakraborty et al., 2016b; Wu et al., 2016). Non-dietary exposure of
inhalation, soil/dust ingestion, dermal absorption, and dietary intake DLCs may contribute more to the total body burdens than dietary ex-
are either adopted or derived from the published data (Table 5), which posure, and may cause more severe effects on human health (Ma et al.,
are estimated based on the models, originated from Nouwen et al. 2008; Qin et al., 2019; Tue et al., 2010, 2014; Xing et al., 2009; Leung,
(2001) and simplified by Minh et al. (2003). Although direct summa- 2006). Dietary ingestion of food grown on DLC-contaminated cropland
tion and comparison of daily intakes from different exposure pathways next to EWRS can further increase their DLC exposure. The estimated
can in some cases be confounding as the differences in sampling, ana- non-dietary daily intake doses of PCDD/Fs for children and adults in
lyzing, and treating data of the studies, it provides some generalizations
Table 5
Comparision of estimated daily intakes of PCDD/Fs and PBDD/Fs from different exposure pathways at EWRS and some other polluted sites.
City Site Air inhalation Soil/dust ingestion Dermal exposure Dietary intake Total intake
Adults Children Adults Children Adults Children Adults Children Infants Adults Children Infants
PCDD/Fs (pg TEQ kg−1 bw day−1)
Guiyu, China EWRS 2.54a 4.50a 0.15b 1.68b 0.35b 0.30b 4.55d 6.67d,e 102.98f 7.59 13.15 105.14
Taizhou, China EWRS 0.68c 1.21c 0.22b 2.25b 0.14b 0.048b 4.55d 6.67d,e 102.98f 5.59 10.18 104.15
Qingyuan, China EWRS 1.82g 3.22g 0.081h 0.84h 0.065h 0.032h 4.55d 6.67d,e 102.98f 6.52 10.76 104.34
New Delhi, Kolkata, Mumbai, EWRS 0.0090i 0.10i 0.021i 0.018 i
Chennai, India
Agbogbloshie, Ghana EWRS 0.11j 1.20j 0.25j 0.21j
Bui Dau, Vietnam EWRS 0.023k 0.26k 0.053k 0.046k
Chennai, India Open dumping sites 0.015l 0.17l 0.036l 0.031l
Hanoi, Vietnam Open dumping sites 0.030l 0.34l 0.071l 0.061l
Wilrijk, Belgium Waste incinerators 0.0065m 0.001m 0.008m 0.068m 0.0073m 0.015m 0.71m 2.63m 0.73 2.72
PBDD/Fs (pg TEQ kg−1 bw day−1)
Guiyu, China EWRS 3.37a 5.98a
Taizhou, China EWRS 0.45n 0.79n 0.37o 4.06o 0.63o 0.11o 6.26p 7.71
Qingyuan, China EWRS 0.30q 0.71q 0.081q 0.94q 0.33q 0.28q 7.56p 8.27
Agbogbloshie, Ghana EWRS 0.25j 2.89j 0.60j 0.52j
Bui Dau, Vietnam EWRS 0.004k 0.047k 0.0097k 0.0084k
a Li et al. (2007).
b Ma et al. (2008).
c Wen et al. (2011).
d Song et al. (2011).
e Liu et al. (2010).
f Chan et al. (2007).
g Xiao et al. (2014).
h Hu et al. (2013).
i Chakraborty et al. (2018).
j Tue et al. (2019).
k Suzuki et al. (2016).
l Minh et al. (2003).
m Nouwen et al. (2001).
n Zhou (2011).
o Ma et al. (2009).
p Miyake et al. (2008).
q Xiao et al. (2016).
13
Q. Dai, et al. Environment International 139 (2020) 105731
three major EWRS of China account for 19%−54% (mean 40%) of the to be investigated and linked with atmospheric deposition or ambient
total daily intake, which far exceeds the WHO recommended total daily air monitoring data. Continuous monitoring in EWRS of DLC release,
intake limit [1–4 pg TEQ kg−1 body weight (bw) day−1] (WHO, 1998b; levels in different environmental media, food and the human body,
van Leeuwen et al., 2000). High exposure quantity and absorption rate potential trends over time, and related human health effects are ur-
lead to higher intake of PCDD/Fs by infants (103.0 pg WHO-TEQ kg−1 gently needed. PBDD/Fs and PXDD/Fs should be considered and in-
bw day−1) for their exclusive consumption of breast milk with rela- cluded in future biomonitoring studies on DLCs because of their sig-
tively high PCDD/F concentration in Taizhou (Li et al., 2007; Gies et al., nificant TEQ contribution. Further investigations on the potencies and
2007). That is higher than the daily intake of Taiwan infants (38.4 pg toxicokinetics of PBDD/F and PXDD/Fs are necessary to develop a
WHO-TEQ kg−1 bw day−1) (Wang et al., 2004) and comparable to that consensus for their TEF. E-waste groups worldwide should work to-
of Duisburg infants (131.0 pg WHO-TEQ kg−1 bw day−1) (Wittsiepe gether to develop unified environmental and biological monitoring
et al., 2007), both of which have been reported to adversely affect their indicators and harmonize survey instruments. Since expensive chemical
neurological and reproductive development (Su et al., 2012; Nowack DLC analytical techniques are out of reach for many resource-poor la-
et al., 2015). Infants and Children are more vulnerable to DLCs because boratories in developing countries, cheap and robust sensitive bioana-
they are at critical windows of growth and development with immature lytical techniques that are easy to set up should be developed and va-
body defenses and faster absorption (Chen et al., 2011). Their lower lidated. Moreover, large and high-quality epidemiological studies on
body weight and larger ingestion of contaminated dust and soil than human health effects of DLCs and their potential pathogenic mechanism
adults can further increase their toxicant body load. There is need for should be conducted. Future research should pay more attention to the
additional investigation and priority protection (Song and Li, 2014; unity of survey methods and the standardization of the exposure re-
Heacock et al., 2016; UN Human Rights, 2016; WHO, 2017). ference category to control heterogeneity.
4. Foresight from current knowledge 5. Conclusion
4.1. Data gaps This review presents the health effects of DLCs and the severe
contamination in EWRS. High DLC-contaminant concentrations derived
Since the Stockholm Convention came into force in 2004, many from informal e-waste recycling put the DLC body burden of workers
countries have carried out nationwide periodical monitoring of back- and residents in EWRS at a high level. Comparison with the levels re-
ground concentrations of DLCs (UNEP, 2016). However, there are ported to be associated with observable adverse effects indicate that
limited case studies on reducing environment levels and body burdens people in EWRS may suffer higher health risks, especially for infants
of DLCs after action was taken. Although there are many epidemiolo- and children. Given the current paucity of information on the human
gical studies on the health effects of PCDD/F and DL-PCB exposure in body burdens and health effects of DLCs in EWRS, these data and
the world, weak associations, inconsistent findings across studies, findings can serve as reference values for the population in EWRS who
complex contaminant mixture, and poor understanding of biological have had long-term exposure to high DLC contamination. More im-
mechanisms make it hard to establish causal relationships. There are portantly, intervention and prevention approaches, as well as con-
also few studies of the body burdens of DLCs and their related health tinuous monitoring of DLCs, should be taken in EWRS to promote the
outcomes among people in EWRS with such high DLC contamination. reuse of e-waste and minimize the adverse impacts on people.
Possible reasons for the scarce data may be the complexity and high
costs of sampling and analysis, and the difficulties in assessing com- Funding
pounded health risks posed by multi-contaminant exposure derived
from informal e-waste recycling. The potential complex interactions for This work was supported by research grant of the National Natural
synergistic effects between the chemical components of e-waste have Science Foundation of China, China (21876065).
not been well explored or understood. We can only compare the body
burdens of DLCs in EWRS with levels observed in epidemiological Declaration of Competing Interest
studies that are associated with various health outcomes to assess the
health risk of people living in EWRS. The authors declare that they have no known competing financial
interests or personal relationships that could have appeared to influ-
4.2. Future work ence the work reported in this paper.
In the future, intervention and prevention approaches should be Acknowledgments
adopted to address the DLC problems in EWRS, including enhancing
regulations, developing a new generation of electronics, assessing ex- We would like to express our deepest appreciation to Dr. Stanley Lin
posures and related health effects in informal e-recycling, improving and Dr. Nick Webber who devoted their time and efforts giving con-
formal recycling technologies, and developing effective remedial tech- structive comments and editing the English language.
niques to contain and eliminate legacy sources of exposure (Ceballos
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