Scholarly article on topic 'A review of the role of emerging environmental contaminants in the development of breast cancer in women'

A review of the role of emerging environmental contaminants in the development of breast cancer in women Academic research paper on "Agricultural biotechnology"

CC BY-NC-ND
0
0
Share paper
Academic journal
Emerging Contaminants
OECD Field of science
Keywords
{"Breast cancer" / "Emerging contaminants" / "Endocrine disrupting chemicals" / Toxicology / Epidemiology / BPA / PBDEs}

Abstract of research paper on Agricultural biotechnology, author of scientific article — Shabana Siddique, Cariton Kubwabo, Shelley A. Harris

Abstract The incidence of breast cancer is on a rise worldwide; it is a disease having a complex etiology. Besides genetics, environmental and other lifestyle factors play a role in the development of the disease. There has been a keen interest in studying associations between breast cancer and exposures to emerging environmental chemicals, which mimic estrogens or influence estrogen levels and signaling in the human body. The common consequence of an endocrine disrupting chemical exposure is that it may have an impact on breast cancer etiology by stimulating formation as well as progression of breast cancer. Exposures to selected emerging environmental contaminants such as alkylphenols (APs), bisphenol A (BPA), parabens, perfluoroalkyl substances (PFASs), phthalates, polybrominated diphenyl ethers (PBDEs), synthetic musks and triclosan, and their probable role in breast cancer development are reviewed. Studies evaluated include the experimental in vitro and in vivo studies as well as human population based studies. In vitro and in vivo evidences indicate that a number of emerging environmental contaminants may play a role in the initiation and/or progression of breast cancer. Although exposures have been assessed in some human populations, breast and other cancer risks associated with these exposures are largely unknown. Efforts should be focussed on the evaluation of these environmental exposures in human populations and their interactions with each other and other genetic and lifestyle risk factors.

Academic research paper on topic "A review of the role of emerging environmental contaminants in the development of breast cancer in women"

Emerging Contaminants xxx (2016) 1—16

ADVANCING RESEARCH EVOLVING SCIENCE

Contents lists available at ScienceDirect

Emerging Contaminants

journal homepage: http://www.keaipublishing.com/en/journals/ emerging-contaminants/

Emerging Contaminants

A review of the role of emerging environmental contaminants in the development of breast cancer in women

Shabana Siddique a' *, Cariton Kubwabo b, Shelley A. Harris a'c'd

a Prevention and Cancer Control, Cancer Care Ontario, 620 University Avenue, Toronto, ON, Canada b Environmental and Radiation Health Sciences Directorate, Health Canada, 50 Colombine Drive, Ottawa, ON, Canada c Occupational Cancer Research Centre, Cancer Care Ontario, Toronto, ON, Canada d Dalla Lana School of Public Health, University of Toronto, Toronto, ON, Canada

ARTICLE INFO

ABSTRACT

Article history: Received 12 August 2016 Received in revised form 7 December 2016 Accepted 12 December 2016 Available online xxx

Keywords:

Breast cancer

Emerging contaminants

Endocrine disrupting chemicals

Toxicology

Epidemiology

The incidence of breast cancer is on a rise worldwide; it is a disease having a complex etiology. Besides genetics, environmental and other lifestyle factors play a role in the development of the disease. There has been a keen interest in studying associations between breast cancer and exposures to emerging environmental chemicals, which mimic estrogens or influence estrogen levels and signaling in the human body. The common consequence of an endocrine disrupting chemical exposure is that it may have an impact on breast cancer etiology by stimulating formation as well as progression of breast cancer. Exposures to selected emerging environmental contaminants such as alkylphenols (APs), bisphenol A (BPA), parabens, perfluoroalkyl substances (PFASs), phthalates, polybrominated diphenyl ethers (PBDEs), synthetic musks and triclosan, and their probable role in breast cancer development are reviewed. Studies evaluated include the experimental in vitro and in vivo studies as well as human population based studies. In vitro and in vivo evidences indicate that a number of emerging environmental contaminants may play a role in the initiation and/or progression of breast cancer. Although exposures have been assessed in some human populations, breast and other cancer risks associated with these exposures are largely unknown. Efforts should be focussed on the evaluation of these environmental exposures in human populations and their interactions with each other and other genetic and lifestyle risk factors.

Copyright © 2016, KeAi Communications Co., Ltd. Production and hosting by Elsevier B.V. on behalf of KeAi Communications Co., Ltd. This is an open access article under the CC BY-NC-ND license (http://

creativecommons.org/licenses/by-nc-nd/4.0/).

1. Introduction

Breast cancer is a disease with a complex etiology and is the most common invasive malignancy among women worldwide. Incidence rates of breast cancer are approximately 90—130 per 10,000 women in developed countries and 10—60 per 10,000 women in developing countries [1]. In the United States, breast cancer accounts for 25% of cancers among women, and in 2011, approximately 230,480 new cases of invasive breast cancer and 57,650 cases of in situ carcinoma were diagnosed [2]. According to the Canadian Breast Cancer Foundation (2013), breast cancer continues to be the most common cancer diagnosed in Canadian women over the age of 20, representing 1 in 4 cancer diagnoses [3].

* Corresponding author. E-mail address: shim23@gmail.com (S. Siddique). Peer review under responsibility of KeAi Communications Co., Ltd.

It is the second leading cause of cancer deaths in Canadian women, after lung cancer. Worldwide, breast cancer incidence is on the rise in both developed and developing countries, perhaps secondary to dietary and reproductive changes. The incidence of breast cancer is higher in the developed/industrialised nations as compared with developing countries, and may in part reflect differences in population level exposures to environmental contaminants.

The factors involved in the etiology of breast cancer are genetic, reproductive, lifestyle related and environmental. The most common and frequently studied inherited mutations, classified as important susceptibility genes, are BRCA1 and BRCA2, which primarily control homologous recombination and repair of somatic damage to DNA and regulate transcription of genes in the DNA damage response pathway [4]. Inheritance of a mutated form of either of the BRCA genes or of other more moderate susceptibility variants in genes such as PALB2, CHEK2 and ATM [5], is associated with a significantly increased risk for development of breast cancer. However, it is estimated that approximately 20—25% of a woman's

http://dx.doi.org/10.1016/j.emcon.2016.12.003

2405-6650/Copyright © 2016, KeAi Communications Co., Ltd. Production and hosting by Elsevier B.V. on behalf of KeAi Communications Co., Ltd. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).

risk for developing breast cancer may be attributed to inherited genetic factors [5].

Epidemiological and clinical evidence confirms that cumulative and sustained exposure to estrogen is a well-established risk factor for breast cancer in women [6,7]. Reproductive factors directly associated with increased estrogen levels, such as early menarche, late menopause, nulliparity or late age at first pregnancy, and lack of breast feeding after childbirth (suppresses estrogens), are consistently associated with increased breast cancer risk in observational epidemiologic studies. Late menarche and early menopause reflect reduced estrogen exposures and are associated with a reduced breast cancer risk [8,9]. Other factors, directly and/or indirectly, can affect estrogen levels in the body or act to modify breast cancer risks. These include dietary factors (mediated via obesity or early menarche), alcohol intake, exercise, exogenous hormone consumption (hormone replacement therapies, HRTs), and oophorectomy [10]. These factors, along with a advancing age, family history of breast cancer, and a history of benign breast disease likely account for 25% of breast cancer risk disease [11].

Environmental (including exogenous estrogen exposures) and other lifestyle factors may account for the remainder of breast cancer risk and therefore could play a major role in the development of the disease [5]. Mammary gland development is a complex process extending from gestation through multiple life stages (like puberty and pregnancy). Normal breast development consists of a series of well-orchestrated events that are finely regulated by a balance of hormones, growth factors, and stromal factors. Chemical exposures during susceptible windows of development may alter the mammary gland in ways that increase risk for development of disease later in life.

A number of epidemiological studies have examined breast cancer risks associated with the legacy environmental contaminants, including the organochlorine pesticides [12], dioxins [13], and polychlorinated biphenyls [14]. The biochemical mechanisms of toxicity for these chemicals have been well studied in laboratory animals, and to a lesser extent in humans, and some health effects have been characterized in occupationally exposed individuals. Much less attention has been paid to the adverse health outcomes as a result of exposure to other more contemporary environmental contaminants, and especially for exposures to these contaminants during susceptible windows of development. Environmental pollutants exhibit a high degree of chemical variability, however, a number of environmental compounds with potentially adverse health effects have common structural motifs, such as containing one or more aromatic rings [15]. As a result, unrelated chemicals may have similar molecular functional pathways and influence human disease in similar ways. For example, diverse chemicals may act as carcinogens by causing DNA damage, or may mimic the biological function of the estrogen hormones. Estrogens play an important role in the etiology of breast cancer through two distinct pathways, first, the by-products of estrogen metabolism cause DNA damage by forming DNA adducts and oxidized bases, thereby, leading to mutations in oncogenes and tumor suppressor genes that control normal cell growth and proliferation. Second, estrogens also alter expression of genes which stimulate growth and proliferation of ductal epithelial cells in the breast [16]. Thus, lifetime exposure to estrogen is a well-established risk factor for breast cancer. A survey conducted by Choi and group [17] revealed that 79% of the endocrine disrupting compound (EDCs) studied were shown to produce carcinogenicity (when positive in at least one animal species). They also demonstrated that EDCs showing estrogen-modulating effects were closely related to carcinogenicity or mutagenicity with a high degree of sensitivity. Very recently a review on mechanisms of action of estrogenic EDCs and their role in development of different diseases, including cancer, has been

published [18]. One common consequence of the endocrine disrupting chemical exposure is that, it may have a profound impact on breast cancer etiology by stimulating formation as well as progression of breast cancer. The chemicals are present in such varied concentrations, that functionality can be reached in many different ways. Thus, any functionality of chemical concentrations in a tissue would need to take account of the amount of all the chemicals (even their isomers), the relative receptor binding affinity (RBA) values of each and the sum of the estrogenic stimulus generated by all the chemicals in combination. Moreover, using animal (rodent) data to assess human health risk must be approached with caution because the exposure profile and often the body elimination is very different. In animal studies, the exposure is often with single compounds in higher concentrations but over relatively short times, whereas in human the exposure is lifelong with low doses and complex mixtures. The contribution of low-level exposures to chemical mixtures that occur in utero and throughout our lifetime is an important factor to consider when evaluating chemical car-cinogenicity. It is likely that low-dose carcinogenic effects occur due to exposures to chemical mixtures, at relevant environmental concentrations, but these effects are difficult to assess in human populations [19].

The objective of this work was to review the laboratory and human evidence, including the presence in biological matrices, on emerging environmental contaminants and their potential to be risk factors for breast cancer.

2. Materials and methods

A PubMed and Scopus search of literature was carried out covering studies conducted over the past two decades (1990—2015), using key words 'breast cancer', 'emerging contaminants', and 'environmental contaminants'. The full text articles were obtained either online or directly from corresponding authors and information on exposure to the contaminants, estimated risks of breast cancer and geographic location was abstracted from the articles.

2.1. Contaminants of emerging concern

Environmental chemicals of emerging concern are substances that have often been present in the environment for quite some time (often decades) but whose presence and significance are only now being elucidated. Data for such substances are often scarce and analytical methods are at the research and development stage or have not yet been internationally harmonised. Thus, it is difficult to interpret and compare the results of exposure and biological monitoring studies, and when available, epidemiologic data, and this poses challenges for regulatory bodies during their decision-making related to the risk assessment and risk management of these compounds. NORMAN (Network of reference laboratories, research centres and related organisations for monitoring of emerging environmental substances) has identified a list of emerging substances and emerging pollutants [20]. These substances were selected by the NORMAN experts, based on citations in the scientific literature. Hence, 'emerging substances' can be defined as substances that have been detected in the environment, but are currently not included in routine monitoring programmes and whose fate, behaviour and (eco)toxicological effects are not well understood. According to the U.S. Environmental Protection Agency (USEPA), an emerging contaminant is defined as 'a chemical or material that is characterized by a perceived, potential or real threat to human health or the environment or lack of published health standards' [21]. A contaminant may also be 'emerging' because of the discovery of a new source or a new pathway of

exposure to humans, or a new detection method or treatment technology has been developed [22]. The U.S. Geological Survey (USGS) defines emerging contaminants as "any synthetic or naturally occurring chemical or any microorganism that is not commonly monitored in the environment but has the potential to enter the environment and cause known or suspected adverse ecological and/or human health effects" [22]. Selected emerging contaminants and their probable role in breast cancer, are discussed in this review.

2.2. Alkylphenols

Alkylphenols (APs) are used as raw materials, in industry, for detergents and surfactants and as antioxidants in manufacturing plastic and rubber polymers. APs are also found in personal care products, especially hair products, and as an active component in many spermicides. The group of alkylphenols include 4-nonylphenol (NP), 4-tert-octylphenol (OP) and 4-tert-butylphenol (BP). For decades, NP and OP have also been extensively used as raw materials for the production of alkylphenol ethoxylates (APEs), especially nonylphenol ethoxylates (NPEs) and octylphenol ethoxylates (OPEs). APEs are the second largest group of nonionic surfactants in commercial use [23] They are widely used in household and industrial detergents, latex paints, pesticides, lubricating oils, emulsifiers, plastics, and paper and the textile industry. Annual global production of APEs was estimated in 2002 to be approximately 650,000 tons [24]. NPEs account for about 90% of APEs and OPEs make up most of the remaining 10% [25].

Short chain APEs and carboxylic acid derivatives (APECs) have been detected in drinking water [26]. Estimates indicate that approximately 60% of APEs end up in the aquatic environment, most entering via sewage treatment plants [25]. They are then degraded aerobically with sequential cleavage of ethyl moieties to shorter-chain and more persistent nonylphenol and octylphenol [27]. These AP metabolites are hydrophobic and thus tend to accumulate in sewage sludge and river sediments. They persist in air, soil and aquatic environments in concentrations that can affect the health of animals and also probably humans [28] [29]. Moreover, APs are modified by chlorination in manufacturing and wastewater treatment plants which complicates the study of their biological effects [30]. Levels reported in the literature from human biological matrices are presented in Table 1. NP is a lipophilic compound and bioaccumulates in aquatic organisms. It has been identified as the most critical congener because of its high resistance to biodegradation, toxicity and strong estrogenic effects [31].

It has been reported that NP stays biologically active for a longer period of time in the body than an endogenous estrogen. NP has

been shown to increase MCF-7 breast cancer cell growth in vitro as well as increases development and proliferation of mammary glands in vivo. Research has suggested that the increased production of estriol (E3), which is an estrogenically active 16-hydroxy product, may be involved in increased susceptibility to breast cancer. Exposure to 4-NP can increase the production of E3, by activating the pregnane-X receptor and inducing P-450 enzymes, thereby, it is suggested that any increase in E3 concentrations by NP could have greater effects on mammary gland proliferation and the development of cancer than NP alone [32]. In 2008, an in vitro study (100 nM, 1 mM, and 10 mM) found an association between exposure to nonylphenol and breast cancer, where nonylphenol was shown to promote the proliferation of breast cancer cells, possibly because of its agonistic activity on ERa in estrogen-dependent and estrogen-independent breast cancer cells [33]. It has been reported that these compounds bind to the nuclear estrogen receptor-a (ER-a) and that the receptor affinity increases with increasing chain length of the alkyl groups [34]. It has also been demonstrated that 4-tertiary branched structures with moderate (C4—C5) and long (C8—C9) alkyl chain length result in higher estrogenic potency than corresponding 4-secondary and 4-normal structures [35]. However, there is inconsistency in results with regards to the level of estro-genic potency and exposure to alkylphenols [36]. It has been recently reported that OP upregulates oncogene expression by inducing abnormal activation of ER-mediated signaling on binding to the ER, thereby inducing cancer cell growth [37]. In 2014, increased expression of proteases on exposure to OP at a concentration of 10~6 M was shown by a study group both in vitro (human MCF-7 breast cancer cells) and in vivo (xenograft mouse model). Investigators concluded that exposure to OP increases cancer proliferation and metastases by altering the expression of proteases through ER-mediated signaling pathways [38]. Additionally, alkyl-phenols are suspected to alter germ cell epigenetic reprogramming during fetal and perinatal development, thus triggering long-term disruption of gene expression which in turn could be a risk factor for hormone-dependent cancers such as breast cancer [39].

2.3. Bisphenol A

Bisphenol A (2,2-bis-(4-hydroxyphenyl)propane; BPA) is a high production volume chemical [40]. It is widely used in polycarbonate plastic products (including toys, water pipes, drinking containers, eyeglass lenses, sports safety equipment, dental monomers, medical equipment and tubing and consumer electronics) and as epoxy resins used as liners in metal cans for foods and beverages and as coatings on metal lids for glass jars and bottles [41]. Under normal conditions of use, BPA has been shown

Table 1

Mean levels of alkylphenols (NP and OP) in different biological in different countries.

Matrix Nonylphenol (NP) Octylphenol (OP) Location Year of sampling Reference

Serum (ng/g) 27.5 ND Canada 2005 [184]

Maternal blood serum (ng/g) <0.5 <0.5 Amsterdam 2003 [185]

Cord blood serum (ng/g) 1.5 <0.5 Amsterdam 2003

Plasma (ng/g) 53.21 16.02 Taiwan NA [186]

Cord blood plasma (ng/mL) ND-15.7 ND Malaysia NA [187]

Fish Bile (ng/g) 29218.0 441.0 Europe 2007—2008 [188]

Fish Bile (ng/g) 4957.0 374.0 Europe 2007—2008 [188]

Urine (ng/mL) 6.25 6.52 Taiwan NA [186]

Urine (ng/mL) 42.06 3.40 Taiwan NA [186]

Breast milk (ng/mL) 32.0 0.12 Italy NA [189]

Breast milk (ng/mL) 0.3 ND Germany NA [24]

Breast milk (ng/mL) 1.05 ND Japan NA [190]

Breast milk (ng/mL) 4.47a 1.29a Taiwan 2000—2001 [191]

The data are expressed as arithmetic mean unless otherwise stated. a: Geometric mean; ND: Not detected; NA: Not available.

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

to leach from food and beverage containers and some dental sealants [42,43]. Numerous studies have determined the presence of BPA in human biological matrices, including serum, urine, am-niotic fluid, follicular fluid, placental tissue, and umbilical cord blood [44], some of which are presented in Table 2. In some studies, the levels of total BPA (free and conjugated) in human blood and other fluids are higher than the concentrations that are likely to stimulate a number of molecular endpoints in vitro [45] and appear to be within an order of magnitude of BPA levels in animal studies [46]. The principal source of human exposure to the chemical is through the diet with estimated dietary intake values ranging from 1 to 1.5 mg/kg bw/d (WHO and FAO) [47].

Biochemical assays have determined that BPA binds both to ERa and ERb, with approximately 10-fold higher affinity to ERb [48], though it is less potent than estradiol, it is known as an endocrine disrupting substance. Recent studies have revealed that BPA can stimulate cellular responses through a variety of molecular pathways, even at very low concentrations [49]. Endocrine disruption from environmental exposure to structural analogues of 17b-estradiol is hypothesized to cause of an array of diseases, including breast and prostate cancer, neurobehavioral disorders, heart ailments and obesity [50]. After reaching the blood, BPA is either reversibly bound to serum proteins (95%) or is unbound (5%). This unbound BPA traverses the capillaries and binds with protein receptors with varying affinity [51]. Estrogenic effects of BPA have been reported in both in vitro and in vivo studies [52]. At concentration range of 0.025—0.25 mg/kg body weight/day, it has been shown that fetal exposure to BPA results in morphological alterations in both the stroma and the epithelium of the developing mammary gland, which can lead to neoplasia later in adulthood. Numerous human studies have shown that BPA is readily and extensively absorbed by the oral route and is metabolized first in the gastrointestinal tract during absorption and then by the liver to biologically inactive glucuronide and sulfate metabolites [53]; this limits the systemic exposure to biologically active BPA to approximately 1% of the absorbed dose in adults [53] [54].

Both transient and permanent effects have been observed in various reproductive tissues of animal models when treated with doses lower (0.025, 0.25 or 25 mg/kg body weight/day) than the USEPA reference dose but within the range of human exposure [55]. BPA is reported to be rapidly eliminated with a terminal half-life of 2.8 h in rhesus monkeys following intravenous administration [56]. Recent studies in rats have demonstrated that the metabolite of BPA (BPA-glucuronide) can easily cross the placental barrier and is reactivated in the fetuses [57]. Exposure of the fetus to BPA could be more detrimental than adult exposure because of critical windows of fetal development that are susceptible to endocrine disruption [58]. Thus, there is significant risk of BPA exposure during critical developmental periods, which may have lasting effects on hormone responsiveness and homoeostatic control of various tissues later in

Table 2

Mean levels of total BPA in serum and urine in general population from different countries.

Matrix BPA (total) Country Year of sampling Reference

Serum (ng/mL) 0.32b Thailand 2009 [192]

Serum (ng/mL) 2.84 China NA [193]

Serum (ng/mL) 0.34a Thailand 2009 [194]

Urine (mg/g) 1.40 Canada 2007- 2009 [195]

Urine (mg/g) 5.88 Norway 2004 [196]

Urine (mg/g) 2.91 Netherlands 2004- -2006 [196]

Urine (mg/g) 2.75 USA 2001- 2004 [196]

Urine (mg/g) 10.5b China 2004- 2008 [197]

The data are expressed as arithmetic mean unless otherwise stated. a: Geometric mean; b: Median; NA: Not available.

life [59]. Exposure to endocrine disrupting chemicals (EDCs) during prenatal period may result in developmental effects and long term modification of organ systems in adults [60], and it is suggested that BPA exerts several estrogenic effects on rodent mammary gland [46]. Studies in mice showed exposure to BPA at 250 ng/kg body weight/day increased the risk for mammary tumorigenesis [61], however, in contrast to the mouse studies, fetal exposure of rats to BPA (0, 25 or 250 mg/kg body weight/day) lead to an increase in susceptibility to carcinogen-induced mammary tumors though no spontaneous tumors were observed [62]. These studies in rodents strongly suggest that BPA exposure during prenatal period possibly alters mammary gland susceptibility to other tumor promoting factors. A study in a mouse model, by Lozada and Keri in 2011, showed that prenatal exposure to BPA at 25 or 250 mg/kg body weight/day increased tumor susceptibility of the mammary gland, in a dose-dependent manner [63]. This same research team was able to observe an adult exposure to BPA directly promoting the growth of estrogen dependent tumors in vivo. Moreover, they also showed that BPA has pathological effects at very low doses and it follows a non-monotonic dose-response curve that is, with very high and low doses eliciting responses in vivo. These results indicate that BPA may increase mammary tumorigenesis by either alterations of the developing fetal mammary gland that increases susceptibility to carcinogenic stresses or promoting tumor cell growth through estrogenic signaling. Nevertheless, both suggest that exposure to BPA at various time points increases the risk of developing mammary cancer in mice. Thus, multiple lines of evidence demonstrate that fetal exposure to low doses of BPA alters cell proliferation, apoptosis and timing in the development of mammary glands, which may further predispose the mammary glands to carcinogenesis [63]. It is also reported that either disruption of the hypothalamic-pituitary-gonadal axis or direct actions on estrogen-sensitive organs by BPA may be involved in the susceptibility of mammary gland tissue to malignant transformation [64].

A human population-based study conducted in 2013 demonstrated that increases in serum BPA levels were associated with a significant increase in mammographic breast density after adjusting for potential confounders [65]. A small age-matched, case-control study conducted in Korea reported that the median level of total serum BPA was higher, though not statistically significant, in breast cancer cases than controls [p = 0.42; median level was 0.61 mg/L and 0.03 mg/L in cases and controls respectively] [66]. Another case-control study in postmenopausal Polish women [67], revealed no association between urinary BPA-G levels and risk of developing breast cancer, although, BPA-G levels were slightly higher in breast cancer cases than controls, results were not statistically significant and there was no indication of a trend. Nevertheless, the research group did observe an increased odds of breast cancer (Quartile 2 vs. Quartile 1: OR 1.70, 95% CI 1.15-2.52), comparing the second to first quartile of BPA-G, however, the OR was not significantly increased for the third or fourth quartiles of exposure.

2.4. Parabens

Parabens are a group of the alkyl esters of p-hydroxybenzoic acid and typically include methylparaben, ethylparaben, n-pro-pylparaben, n-butylparaben and isobutylparaben. Parabens are lipophilic compounds with increasing octanol/water partition coefficients (log Kow ranging from 1.66 to 3.24) and antimicrobial activity with increasing molecular weight and length of the alkyl side chain [68]. Due to their relatively low acute toxicity, parabens (or their salts) are used as preservatives in many consumer products [69] including cosmetics, toiletries, food and pharmaceuticals

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

[70]. Their extensive use as preservatives has led to their widespread distribution at measurable levels in the aquatic environment [71], house dust [72], human tissues [73], milk [74], blood [75], urine and semen [76]. The concentrations of parabens in different human matrices are presented in Table 3.

Parabens are mainly metabolized to p-hydroxybenzoic acid which elicit less estrogenic activity than the parent compounds. In vitro studies have demonstrated that parabens undergo hydrolysis in human plasma and that the rate of hydrolysis depends on the length of the chain [77]. However, a recent study in rats reported that parabens are metabolized before reaching plasma [78]. Hydrolysis of parabens has been demonstrated in human liver, small intestine and skin. Parabens undergo phase II reactions, and both glucuronide and sulfate conjugates have been detected in human urine [79]. Pathological conditions, such as liver or kidney dysfunctions, may increase their half-life and contribute further to their accumulation and hence toxicity [80]. In humans, parabens can reach the circulation upon oral administration or dermal application of cosmetic creams and their levels in blood depends on the extent of exposure [75]. Methylparaben, despite having the lowest lipophilicity, penetrates the skin to the greatest extent [81]. Parabens have weak estrogen activity and have been shown to induce the growth of MCF-7 human breast cancer cells in vitro which have made the researchers suggest their potential as initiators or promoters of breast cancer [82]. It has recently been reviewed that the in vivo estrogenicity of parabens are not as low as demonstrated by the in vitro assays [83].

In two studies conducted in the 1970s, investigators reported that parabens can cause chromosomal aberrations, particularly in the co-presence of PCBs. It has also been reported that subcutaneous administration of methylparaben causes mammary ade-nocarcinomas in rats [84] [85]. Parabens are also shown to be capable of disrupting cellular function through inhibiting secretion of lysosomal enzymes and causing mitochondrial dysfunction [86]. As the parabens possess estrogenic activity, and that the activity increases with the length and branching of the alkyl ester, they can bind to estrogen receptors and are able to mediate unwanted effects even at much lower concentrations and more specifically through non-receptor mediated mechanisms [87]. There is also a debate as to the estrogenic effect of parabens being too weak to be a concern, therefore, the current consensus is that cancer risk of exposure to parabens is due to more than estrogen mimicry [88]. An alternate mechanism has been proposed by which parabens can indirectly affect estrogen levels (i.e., an increase in the levels of estrogen is seen due to inhibition of sulfotransferase activity) [89]. Parabens with longer side-chains showed greater affinity for estrogen receptors and they had similar relative binding affinity (RBA) values to both ERa and ER$. Although their relative binding affinity (RBA) for the estrogen receptor is low, their efficacy in cell proliferation and intracellular signaling pathway is significant. The findings of parabens in their free form in both mammary gland and

in breast cancer tissues [87], support the general hypothesis that there may be a link between breast cancer and estrogenic compounds commonly used in underarm cosmetics and other consumer products.

In a study by Darbre et al. (2004) the paraben concentrations measured in tumors were unequivocally of the esters themselves [82]. This demonstrates that at least a portion of the parabens present in cosmetic, food and pharmaceutical products can be absorbed and retained in human body tissues without hydrolysis by tissue esterases to the common metabolite p-hydroxybenzoic acid. The results alone, however, do not suggest that these chemicals caused the tumors in these patients, but they support a potential for the parabens to exert an estrogenic stimulus in the human breast [90]. A relatively recent review of literature concludes that there is no link between parabens and breast cancer [91]. However, the potential effects of long-term low-level exposures or exposures to parabens in mixtures with other chemicals in humans has not been studied. Furthermore, the effects of early life exposure (directly in pre-adolescents) and effects in other sensitive or predisposed subgroups of the general population are currently not known.

2.5. Perfluoroalkyl substances

Perfluoroalkyl substances (PFASs) are a large group of emerging anthropogenic environmental contaminants. PFASs consist of an alkyl chain (4—14 carbons), which is partially or fully fluorinated, and have different functional groups attached. Among the PFASs are the perfluoroalkyl carboxylic acids such as perfluorooctanoic acid (PFOA) and perfluorodecanoic acid (PFDA), perfluoroalkyl sulfonic acids (PFSA) like perfluorooctane sulfonate (PFOS), per-fluoroalkyl sulfonamides such as perfluorooctylsulfonamide (PFOSA), and other polyfluorinated compounds, such as fluo-rotelomer alcohols (i.e., fluorotelomers). These chemicals have been produced since the 1950s and are used in many industrial and commercial applications (e.g. non-stick cookware, waterproof and breathable textiles, and protective coatings for paper, food packing materials, and carpets). PFASs are very resistant to biodegradation because of their carbon-fluorine bond, and thus they are very persistent in the environment. It is suggested that the transformation or biodegradation of precursor perfluorinated chemicals occurs by both abiotic and biotic degradation pathways where the typical final degradation products are PFOS and PFOA [92]. Long-chain perfluoroalkyl substances (PFAAs with eight or more carbons) are of specific concern because they are persistent, bio-accumulative and have shown toxicity in animal studies. PFASs have been used worldwide in a wide variety of industrial and consumer product applications. A ban on most uses of PFOS was imposed in the US in 2000 and in the European Union in 2008. In 2006, the manufacture, sale, and importation of PFOS as well as PFOS-containing products were banned in Canada. In January 2009,

Table 3

Mean concentrations of parabens in biological samples from published studies.

Matrix Methyl Paraben Ethyl Paraben Propyl Paraben Country Year of Sampling Reference

Breast Milk (ng/mL) 2.18 1.26 1.42 Switzerland 2004—2006 [74]

Urine (ng/mL) 142.7 23.0 48.1 USA 2003—2005 [79]

Urine (mg/g) 58.6a 8.63a USA NHANES study [198]

Urine (mg/mL) 17.7b 1.98b 3.6b Denmark 2006 [76]

Urine (mg/mL) 28.6a 3.67a USA 2000—2004 [199]

Urine (ng/mL) 6.5a 1.89a 3.57a China 2010 [200]

Breast tumor tissue (ng/g) 12.8 2.0 2.6 Scotland NA [82]

Breast tumor (ng/g) 17.9b 3.2b 14.2b UK 2005—2008 [73]

Serum (ng/mL) 1.53b <LODb 0.32b Denmark 2006 [76]

The data are expressed as arithmetic mean unless otherwise stated. a: Geometric mean; b: median; LOD: Limit of detection; NA: Not available.

PFOS and its salts were added to the Virtual Elimination List under Canadian Environment Protection Act (CEPA). PFOS and its salts have been added to Annex B of the Stockholm Convention on Persistent Organic Pollutants (POPs).

Unlike the other POPs (e.g. PBDEs and PCBs), which accumulate in lipid rich tissues, these PFAs bind to blood proteins and accumulate mainly in liver and kidney. PFASs have been detected in wildlife, human blood, breast milk, water (surface, ground, drinking), house dust and soils [93] [94,95]. The levels reported in biological matrices are presented in Table 4. Food intake appears to be the major factor contributing to background PFA levels in human sera, while exposure to contaminated water and soils results in elevated levels in both human and wildlife populations [96] [97]. In vivo studies have documented an array of toxicological outcomes including liver hypertrophy and tumors, thyroid hormone alterations, developmental toxicity, immunotoxicity, and carcinogenic potency [98]. Both animal as well as in vitro studies have shown that PFAAs may have potential genotoxic and neurotoxic effects [99] [100]. A two-year study in rats by Sibinski and his group, found effects of PFOA exposures on reproductive tissues at 5 mg/kg body weight, including a significant increase in mammary fibroadenomas and Leydig cell adenomas [101]. In addition, PFOA-exposed female pups at a concentration of 5 mg/kg body weight showed stunted mammary gland epithelial branching and growth [102]. It has been documented that PFOS affects antibody production in mice at levels found in the general human population [103]. Data on the immunotoxicity of PFASs in humans are limited, however, they studies suggest that exposure to PFASs may be associated with immunosuppressive effects [104]. A study published in 2006 showed estrogen-like properties of PFASs in human MCF-7 breast cancer cells suggesting endocrine potentials [105]. Studies suggest that some of the biological effects of the PFASs are mediated through peroxisome proliferator-activated receptors (PPARs). It is suggested that PPAR-a is the most likely target of PFOA and PFOS, with the former having the highest affinity in both human and mouse isoforms [106]. It is substantiated by numerous studies that PFCs are immunotoxic, thus, dysregulation(s) of immune system homeostasis can lead to adverse changes in immune functions, increasing the susceptibility to infections and cancer, as well as favoring the development of autoimmune diseases.

In humans, an elevated incidence of bladder cancer mortality among male workers exposed to PFOS has been reported [106]. In a

case control study on women from Greenland, it was found that the breast cancer risk, before and after adjusting for the confounders, was associated with serum level of PFOS (adjusted OR = 1.03, p = 0.05) and the sum of PFSA (adjusted OR = 1.03, p = 0.02) [107]. In another case-control study on Danish women, a weak positive association was found between breast cancer risk and the continuous serum PFOSA data (unadjusted RR 1.03, 95% CI 1.00-1.07; adjusted RR 1.04, 95% CI 0.99-1.08), a significant elevated breast cancer risk for PFOSA in the 5th quintile was seen, with a RR of 1.89 (95% CI 1.01-3.54) with adjustment for confounding. Interestingly, upon stratification for age at diagnosis, the positive association of PFOSA in the highest quintile and BC risk was stronger for women 40 years of age or younger (adjusted RR 3.42, 95%C1 1.25-9.36) [107].

2.6. Phthalates

Phthalates have been used extensively as plasticizers to improve the flexibility of polymers, particularly polyvinyl chloride (PVC), since first introduced into commercial markets nearly 75 years ago. 1n addition, phthalates have found widespread applications in the cosmetic and personal care product sectors, where they are used as antifoaming agents, dispersants, and emulsifiers. Worldwide production of phthalates has been estimated at 3.5 million tons/year, and currently, over 80% of all plasticizers are phthalates [108]. Phthalates are ubiquitous environmental contaminants potentially leading to adverse human health outcomes given their wide range of industrial and personal applications [109]. Because most phthalates are not chemically bound to the polymer, they can easily leach or outgas into the surrounding air, foodstuff and other matrices. Contaminated foodstuff is considered one of the major sources of exposure [110].

The metabolism of most phthalates in humans occurs by hydrolysis of one ester bond to form the hydrolytic phthalate monoesters. Some phthalates may undergo a phase 1 biotransformation where oxidative metabolites are formed. Both metabolites (monoester and oxidative) may react with glucuronic acid in a phase 11 biotransformation to form their respective glucuronic conjugates [111]. The widespread exposure of the general population to several phthalates has been documented in a number of human biomonitoring studies through the assessment of the urinary concentrations of their metabolites, all over the world [109].

Table 4

Level of PFOS and PFOA in serum and milk from different countries.

Matrix PFOS PFOA Country Year of sampling Reference

Serum (ng/mL) 32.65 5.09 USA 2000—2002 [201]

Serum (ng/mL) 8.28 6.16 Colombia 2003

Serum (ng/mL) 11.74 <20 Brazil 2003

Serum (ng/mL) 4.32 <3 Italy 2001

Serum (ng/mL) 42.14 21.34 Poland 2003

Serum (ng/mL) 15.66 4.82 Belgium 1998—2000

Serum (ng/mL) 1.85 2.64 India 2000

Serum (ng/mL) 12.74 <10 Malaysia 2004

Serum (ng/mL) 21.1 61.8 Korea 2003

Serum (ng/mL) 16.15 4.21 Japan 2002

Serum (ng/mL) 20.7 3.7 USA 2003—2004 [202]

Serum (ng/mL) 13.2 3.3 Sweden 2001—2004 [203]

Serum (ng/mL) 28.8 3.4 Canada 2002 [204]

Milk (mg/L) ND 0.11 China 2004 [205]

Milk (mg/L) ND <0.21—0.49 Sweden 1996—2004 [206]

Milk (mg/L) ND <0.004—0.34 Japan 2007 [96]

Milk (mg/L) ND 0.077 Germany 2006 [207]

Milk (mg/L) ND 0.044 USA 2004 [208]

Milk (mg/L) ND 0.25 Canada 2003—2004 [209]

The data are expressed as arithmetic mean unless otherwise stated. ND: Not detected.

Table 5

Concentration of low molecular weight urinary phthalate metabolites in adults from published studies.

MEHP MBzP MBP MEP Country Reference

Urine (creatinine corrected, mg/g)

15.1a 6.9a 81.9a 130.0a Poland [210]

11.0 a 29.4a 33.8a 771a USA [211]

2.8 a 2.6a 20.9a NA Japan [212]

5.1 6.3 82.5 106.8 Mexico [124]

39.6 a NA NA NA Korea [213]

3.7 5.5 27.5 42.3 Germany [214]

6.4 a 5.5 a 67.0a 64.0 a Denmark [215]

Urine (not corrected for creatinine, ng/mL)

5.4 NA 22.9 6.5 China [216]

3.0 2.3 53.6 22.5 China [217]

3.4 1.4 13.1 132 India

1.7 1.4 13.1 12.1 Japan

4.0 3.9 17.0 13.7 Korea

4.6 3.2 74.1 272 Kuwait

4.7 2.4 14.5 38.6 Malaysia

2.4 1.2 17.0 6.5 Vietnam

7.2 4.4 94.7 399.2 Egypt [218]

The data are expressed as geometric mean (GM) unless otherwise stated. a: Mean. NA: Data not available.

Selected biomonitoring data for the low molecular weight phthalate metabolites is presented in Table 5. Phthalates are generally less potent than endogenous estrogens but are a cause for concern because they are persistent in the environment, resistant to chemical or enzymatic degradation, and are able to sequester and accumulate in the adipose tissues. Hence, their role in steroid hormone-dependent cancers, such as breast cancer, should not be overlooked [112].

Estrogen regulates cellular responses by binding to ERs, regulating transcription of target genes in the nucleus and activating signaling pathway in the cytoplasm. The most commonly used phthalates [i.e., di-n butyl phthalate (DBP), benzyl butyl phthalate (BzBP), bis(2-ethylhexyl) phthalate (DEHP)] in consumer products are capable of binding to ERs. Thus the type of effects that may potentially be induced as a result of these phthalates and their ERbinding capabilities is a cause of concern [113]. A recent in vitro study (cells treated at 10~10 to 10~4 M/L) by Chen and Chien (2014) demonstrated that BzBP, DBP and DEHP stimulate breast cancer cells and may induce proliferation even at low concentrations [112]. These chemicals stimulate the P13K/AKT signaling pathway; akin to the effects that 17$-estradiol has on breast cancer cells. Another in vitro study by Kim et al. showed that the phthalates BBP, DBP, and DEHP mimic estrogen in the inhibition of TAM-induced apoptosis in MCF-7 cells by a mechanism involving Bcl-2 and Bax gene expression [114]. Recent studies have shown toxic effects of phthalates at environmentally relevant levels (96.5 ng/mL in cases vs. 26.4 ng/mL in controls) to which the general population is exposed [115] [116]. Studies have shown that even at very low concentrations, BBP, DBP and DEHP were capable of inducing proliferative effects in vitro and it has been shown that BBP at a concentration of 10 mM increases the proliferation ability and induces mitosis in MCF-7 and ZR-75, which are ER-positive breast cancer cell lines and that these effects may be related to estrogenic activity as BBP has been shown to compete with estradiol for binding to the estrogen receptor [117]. A study on MCF-7 cells, has also shown an altered ER mRNA expression by BBP and this might be related to aberrant DNA methylation in the promoter region of ER [118]. At 1 mM/L, BBP has also been shown to induce expression of the oncogenes c-Myc and HDAC6 (encoding histone deacetylase 6) in ERnegative breast cancer cell lines, which implies a novel oncogenic mechanism of these chemicals in breast cancer independent from their estrogenic activities [119]. However, any of these effects do not

display similar estrogenic effects in vivo [120]. There are little data on the effects of BBP in female rats. However, MEHP at 100 mM and DEHP at 1000 mg/kg body weight (dosed for 8—9 days), have been observed to decrease serum estradiol levels, and prolong estrous cycles acting through a receptor-mediated signaling pathway to suppress estradiol production in the ovary [121]. These chemicals may have an effect in other hormone-sensitive organs such as the mammary gland. A recent study showed that in utero exposure to BBP, at a concentration range of 120 mg—500 mg/kg body weight/ day, significantly affected post-natal maturation of female rats including changes in morphology of the mammary gland, and that these effects were dose-dependent [122]. Numerous in vivo (at concentration of 600 mg/kg body weight) and in vitro (cells treated at concentration range of 10~6 to 10~4 M) studies have also found that phthalates may contribute to breast cancer development through estrogen receptor-independent mechanisms [123]. One such mechanism proposed is the activation of Aryl hydrocarbon receptor (AhR), which is a ligand-activated transcription factor. On binding with phthalates, AhR, undergoes conformational change, is activated and regulates gene expression through both nongenomic and genomic mechanisms and thus confers a significant effect on the pathophysiology of human cancer [119].

In a case-control study in Mexico, the median urinary concentration of mono-ethyl phthalate (MEP), the main diethyl phthalate (DEP) metabolite, was reported to be significantly associated with breast cancer, after adjusting for risk factors and other phthalates [the odds ratio (OR), highest vs. lowest tertile = 2.20; 95% confidence interval (CI), 1.33—3.63; p for trend < 0.01]. This association became stronger when estimated for premenopausal women (OR, highest vs. lowest tertile = 4.13; 95% CI, 1.60—10.70; p for trend <0.01) [124]. The same study reported a negative association of breast cancer with the metabolites of butyl benzyl phthalate (BBzP) and of dinoctyl phthalate (DOP). Studies in humans have also shown that MEP has the potential to induce DNA damage and, therefore, increase cancer risk [125—127]. An occupational study on premenopausal women working in automotive (OR = 2.68; 95% CI, 1.47—4.88; p = 0.0013), plastics (OR = 2.43; 95% CI, 1.39—4.22; p = 0.0018) and food-canning industries (OR = 2.35; 95% CI, 1.00—5.53; p = 0.050) revealed excess risk for developing breast cancer, more likely due to their exposure to phthalates [128]. However, in a case-control study conducted 1998 and examining occupational exposure to estrogenic chemicals (including BBzP) found no association with breast cancer risk [129]. The discrepancy in results of these epidemiological studies can be attributed to differences in chemical exposure profiles across studies, difference in premenopausal and postmenopausal status of the participants, different study data collection methods, etc. These differences are common in epidemiologic studies and make interpretation of health risks more difficult. If biomonitoring data are available in these studies, it tends to strengthen the study validity and gener-alizability to external populations.

2.7. Polybrominated diphenyl ethers

Polybrominated diphenyl ethers (PBDEs) are an important class of brominated flame retardants that have widely been used as additives in electronic equipment, casings for computers and television sets, building materials, polyurethane foams, carpets and upholstery and a variety of other plastic products, to reduce the risk of fire [130]. There are 209 possible congeners of PBDEs and their physical and chemical properties are very similar to those of pol-ychlorinated biphenyls. Three major commercial mixtures of the PBDEs are produced, which vary in the degree of bromine substitution on the aromatic rings [130]. Commercial PBDE mixtures are produced at three degrees of bromination to give i) deca-BDE,

which consists of decabromodiphenyl ether (97—98%); ii) octa-BDE, which consists of hexabromodiphenyl ethers (10—12%), heptabro-modiphenyl ethers (43—44%), and octabromodiphenyl ethers (31—35%); and iii) penta-BDE, which consists of pentabromodi-phenyl ethers (50—62%) and tetrabromodiphenyl ethers (24—38%). According to the Bromine Science Environmental Forum (BSEF), the Deca-BDE mixture is the most widely used formulation, accounting for approximately 83% of the total PBDEs produced worldwide. According to the 2010 BSEF report, the three most commonly used commercial mixtures have been produced in large volumes with recent global production rates of more than 50,000 tonnes/year, while the total historical production of PBDEs between 1970 and 2005 was estimated in 1.3—1.5 million tonnes [131].

As additive flame retardants, PBDEs are not covalently bound to the polymers, therefore, they easily leach out into the environment. In the last twenty years, PBDEs have become ubiquitous persistent organic pollutants; some bioaccumulate in the environment, bio-magnify in the food chain, and are being detected in significant amounts in animals as well as humans. In response to the growing concern about their persistence, bioaccumulation and potential health effects the penta- and octa-PBDE, these formulations were banned in the European Union in 2004 [132] and are phased out in the US [133]. Recently, these two formulations were added to the list of banned chemicals in Annex A (elimination of production and use of all intentionally produced Persistent Organic Pollutants) of the Stockholm Convention in 2009 [134]. The deca-PBDE on the other hand, is still in use in the US [135] and was annulled from the European Union restriction in 2008 [136] The commercial PBDE mixtures are used without regulation in most Asian countries [137] and are still ubiquitous in the environment in the developed countries, partially due to their persistence and partially owing to their ability for long-range transport.

The reported levels of PBDEs in different populations are presented in Table 6. The hydroxylated PBDE (OH-PBDE) metabolites and methoxylated PBDE (MeO-PBDE) analogues have also been detected in human blood [138]. Interestingly, the fully brominated deca-BDE congener is poorly absorbed, rapidly eliminated, and does not bioaccumulate; it is likely one of the least bioactive congeners of the PBDEs [139]. In contrast, the lower molecular weight congeners, tri-to hexa-BDEs, are almost completely absorbed, slowly eliminated, highly bioaccumulative, and much more bioac-tive than deca-BDE. However, an important consideration is that

deca-BDE, when exposed to sunlight, is converted to the lower molecular weight, bioaccumulative congeners [140,141]. It has been shown that, the lower molecular weight congeners are selectively taken up and retained in the liver, adrenal cortex, and ovaries after PBDE exposure in adult C57BL mice [142]. The mechanisms for PBDE toxicity are complex and most of the studies have focussed on the toxic effects through some nuclear receptor mediated pathways [143].

Because of their persistence in the environment, and bioaccumulation in the food chain, even low levels of PBDEs might pose a risk to human health. The toxicological endpoints likely to be of the greatest concern for low, environmental concentrations of PBDEs are thyroid hormone disruption, neurodevelopmental effects, decreased female fecundability and for some congeners, cancer [144,145]. Moreover, it has also been reported that the metabolism of PBDEs enhances its toxicity [146]. Experimental evidence as well as in vitro studies have shown that PBDEs are endocrine-active compounds with the potential to interfere with thyroid hormone homeostasis, as well as to interact with steroid receptors (e.g. estrogens, androgens) and aryl hydrocarbon receptors (dioxin-like activity) [147] [148]. Exposure of MCF-7 breast cancer cells to low doses (10~12 to 10~9 M) of PBDE resulted in changes in growth kinetics, micronucleus formation and cell proliferation [149]. An in vitro study on MCF-7 cell line, have shown penta-BDE in the concentration range of 10~8 to 10~4 mol/L exerted estrogenic actions and stimulated proliferation in estrogen responsive breast cancer cells (MCF-7) via up-regulation of Bcl-2 mRNA and protein expression [150]. Another study has remarkably demonstrated similarity in function of BDE-99 and estrogen, whereby exposure to BDE-99 in the range of 10~4 to 10~7 mol/L stimulated proliferation in MCF-7 cells, promoted c-Myc mRNA and protein, and inhibited p53 mRNA and protein expression [151].

Numerous in vivo studies have established that PBDEs affect both male and female reproductive systems [147]. Moreover, it has also been demonstrated that exposure to low concentrations of PBDE-47 in utero and during lactation decrease the offspring's ovarian weight and size of tertiary follicles [152]. In contrast to the common notion that BDE-209 is less toxic, one study showed that PBDE-209 at a high dose of 5—100 nM induced not only prolifera-tive effects but also antiapoptotic effects in different cancer cell lines and activated PKCa and ERK1/2 phosphorylation [153]. One limitation of this study, however, was that the cell lines were

Table 6

Concentration (ng/g lipid) of BDE congeners in serum from different study population.

Study Population BDE-47 BDE-99 BDE-100 BDE-153 BDE-183 BDE-209 Country Year of sampling Reference

General 23.0 5.4 4.4 10.0 0.46 1.9 Canada 1992—2005 [219]

General 17.7a 5.0a 3.0a 6.0a ND ND USA 2003—2005 [220]

General 2.3b 2.3b 1.6b 0.81b 0.6b 1.1b Spain NA [221]

General 0.65 0.17 0.23 0.59 0.07 2.94 Greece 2007 [222]

General 16.0a 2.6a 3.0a 7.1a ND ND USA 2007—2008 [223]

General 0.37a 0.1a 0.11a 0.31a 0.06a 1.2a Japan 2005 [224]

General 1.6b ND ND 0.57b 0.12b ND Sweden NA [225]

General 0.82b <0.16b 0.76b 1.7b 0.3b <15b UK 2003 [226]

Pooled 31.0 6.8 6.8 13.8 ND ND USA 2003—2008 [227]

General

Pregnant women 15.8a 4.4a 2.8a 2.4a <LOD ND USA 1999—2000 [228]

Pregnant women 1.3b 0.33b 0.51b 1.0b ND 0.77b Faroe Island 1994—1995 [229]

Pregnant women 0.8b 0.2b 0.2b 1.6b ND ND Netherlands 2001 —2002 [230]

Pregnant women 26.94a 9.75a 3.69a 3.88a 2.5a ND Canada 2004—2005 [231]

Children 5.6a 1.6a 1.6a 2.4a ND ND Mexico 2007—2008 [232]

Residents of coastal area 21.4 23.4 14.8 32.5 46.9 403 China 2007 [233]

High consumption of fish 48.3 11.6 9.69 9.82 ND ND USA 2001 —2003 [234]

e-waste recycling workers 3.3 3.2 1.6 7.4 8.6 260 India 2007 [235]

Electronics dismantlers 2.9b ND ND 4.5b 7.8b 4.8b Sweden NA [225]

The data are expressed as arithmetic mean unless otherwise stated. a: Geometric mean; b: median; NA: Not available; LOD: Limit of detection; ND: Not detected.

treated at a dose not representative of environmental exposure. Similarly, the majority of environmentally relevant dosing studies fail to consider the chronic and prolonged exposure effects of these environmentally ubiquitous chemicals. These animal studies indicate that PBDEs may significantly affect the reproductive system and be responsible for increasing risks of cancer in the mammary glands, uterus, and ovary.

Elevated cancer rates have been reported in human populations residing in areas of intense electronic waste recycling plants where high levels of PBDEs have been observed. However, in a case-control study by Hurley et al. [154] to evaluate the risk of breast cancer associated with adipose PBDE concentrations among women undergoing surgical breast biopsies in the San Francisco Bay Area (CA), no association could be made between PBDE levels in adipose tissues and breast cancer risk. The median concentrations for five major congeners measured were slightly higher in the controls than in cases; however, it was not statistically significant as the P-value ranged from 0.16 to 0.80, and adjusted Odds Ratios also did not significantly differ from 1.0. The above study had some limitations; the controls were possibly overmatched to cases, a choice of small sample size and the age distribution was skewed such that controls were significantly younger than the cases. 1nterestingly, a recent case-control study among Alaskan native women suggested a possible association between breast cancer and BDE-47, with a Univariate Odds Ratio of 1.79 (p = 0.06), although a multivariable adjusted analysis did not support this association (OR = 1.58, p = 0.23) [155].

2.8. Synthetic musks

Synthetic musks are used extensively in a number of personal-care products in lieu of natural musks because of their low production costs and easy preparation. They are also used in technical products such as herbicide formulations [156] [157]. Synthetic musks are often divided into three major classes: (!) nitromusks, including musk ketone (4-tertbutyl 3,5-dinitro-2,6-dimethylacetophenone), musk xylene (1-tert-butyl-3,5-dimethyl-2,4,6-trinitrobenzene), musk ambrette (MA, 1-tert-butyl-2-methoxy-4-methyl-3,5- dinitrobenzene), and musk moskene (MM, 1,1,3,3,5-pen tamethyl-4,6-dinitroindane); (2) polycyclic musk compounds, including HHCB (Galaxolide®, 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-(g)-2-benzopyran), AHTN

(Tonalide®, 7-acetyl-1,1,3,4,4,6- hexamethyl-1,2,3,4-

tetrahydronaphthalene), ADB1 (Celestolide®, 4-acetyl-1,1-dimethyl-6-tert-butylindane), AHM1 (Phantolide®, 6-acetyl-1,1,2,3,3,5-hexamethylindane), AT11 (Traseolide®, 5-acetyl-1,1,2,6-tetramethyl-3-isopropylin dane), DPM1 (Cashmeran®, 6,7-dihydro-1,1,2,3,3-pentam ethyl-4(5H)-indanon); and (3) macrocy-clic musks such as MT (Musk T®, 1,4-dioxacycloheptadecane-5,17-dione). These compounds, along with OTNE (marketed as 1so E Super®, (1,2,3,4,5,6,7,8-octahydro-2,3,8,8-tetramethyl naphthalen-2yl)ethan-1-one), are widely used in various consumer products such as perfumes, body lotions, soaps, shampoos, shower gels,

bubble bath, facial creams, and other cosmetics, air fresheners, detergents, fabric softeners, and household cleaners; they can also be used in food additives, cigarettes and fish bait. Synthetic musks can be released into the environment and they have been detected in water, sediment, house dust, tissues of various organisms, air, and snow [158] [159] [160]. The production of HHCB and AHTN was approximately 1800 tons/year in Europe in 2000 [161] which henceforth demonstrates its high commercial demand. 1n the United States, the production of HHCB was more than 4500 tons/ year and is thus listed by the USEPA as a high production-volume chemical for uses that are reportable under the Toxic Substances Control Act [162].

Their endocrine toxicity has been demonstrated by a number of researchers [163] [91]. Musk xylene is also listed as a hazardous substance to be phased out by the European Union [164]. There have been records of significant bioaccumulation and detection of most commonly used polycyclic musks such as HHCB (Galaxolide) and AHTN (Tonalide) in aquatic organisms which might be attributed to their lipophilicity (log Kow value of 5.4-6.3) [156] [165]. Similarly, occurrence of these two compounds in human adipose tissues has been reported in Germany and Switzerland [166] and recently in the United States [162] and they have been detected in human milk [167]. The levels of synthetic musks in human milk reported in the literature are presented in Table 7.

Musk ketone is a suspected comutagen, whereas, musk xylene is a suspected comutagen and carcinogen. Musk xylene has been classified in group 3 (not classifiable with respect to its carcino-genicity to humans) by 1ARC [168]. Although musk xylene and musk ketone are structurally similar and possess nearly identical physico-chemical properties, they differ significantly in their biological properties [169]. In an in vitro study, musk xylene and musk ketone were shown to probably possess estrogenic activity, and musk ketone has shown an affinity to the ER three times greater than muskxylene [169]. The first step in the transformation of nitro fragrances is reduction to the amine [161], interestingly, when musk ketone is reduced the estrogenic activity is lost; whereas, reduction of musk xylene leads to an increase of estrogenic potency [169]. Moreover, reduction to the amine in para position of musk xylene leads to an enhanced proliferative effect, in contrast, reduction of the amine group in ortho position results in a cytotoxic effect [169]. Thus, it can be observed that reduction of a nitro group of musk fragrances appears to result in a position-dependent variable influence on the biological effect of the metabolite [169]. In a study on rodents, musk ketone showed a strong induction of enzyme activities, which was species specific and distinct from musk xylene, thus, as a result of its enzyme modulating activity, musk ketone acts as a co-genotoxicant in combination with several human carcinogens, even at low doses [168].

As of date, very few studies have reported concentration of synthetic musks in biological matrices such as serum, plasma or whole blood. The scarcity of exposure data on musks in those matrices is mainly due to the large volume (5-10 mL) required to do the analysis and the lack of isotopically labelled standards to use

Table 7

Level of synthetic musks in breast milk.

Matrix Musk Xylene Musk ketone Country Year of sampling Reference

Milk (ng/g) 30.0 10.0 Germany NA [167]

Milk (ng/g) 9.44 14.9 Denmark 1999 [236]

Milk (ng/g) 17.0 58.2 USA 2004 [237]

Milk (ng/g) 9.5 <5.0 Sweden 1996—2003 [157]

Milk (ng/g) 8.0 ND Germany 2005 [238]

Milk (ng/g) 17.0 4.0 China 2006—2007 [239]

The data are expressed as arithmetic mean unless otherwise stated. a: Geometric mean; b: median; NA: Not available.

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

Table 8

Levels of Triclosan in different biological matrices from across the world.

Matrix Triclosan Country Year of sampling Reference

Urine (mg/g) 19.9 USA 2007- -2008 [240]

Urine (ng/mL) 13.0a USA 2003- 2004 [241]

Serum (ng/mL) 14.9 USA 1998 2003 [242]

Serum (ng/g) 19.0 Australia 2002 2003 [243]

Serum (ng/mL) 8.5 Hong Kong NA [244]

Serum (ng/mL) 1.71b Belgium NA [245]

Plasma (ng/mL) 16.0 Sweden 2003 2004 [246]

Milk (ng/mL) 0.54 Sweden 2003 2004 [246]

Milk (ng/mL) 9.3 USA NA [247]

The data are expressed as arithmetic mean unless otherwise stated. a: Geometric mean; b: Median; NA: Not available.

in quantitative analysis. As a result, there are no human studies that have been reported to date indicating a breast cancer risk with exposure to musks; however, with the in vitro studies documenting strong estrogenic activity of the synthetic musks, they could play a significant role in hormone dependent cancer, including breast cancer.

2.9. Triclosan

A number of pharmaceuticals and personal care products (PPCPs) are emerging environmental pollutants that are of increasing concern owing to their persistence and bioaccumulative ability. One such synthetic, broad spectrum antibacterial agent that is widely used in PPCPs, in polymers and fibers to give various textiles (toys, undergarments, cutting boards) antibacterial properties is triclosan (TCS, 2,4,4'-trichloro-2'-hydroxydiphenyl ether) [170]. TCS is a lipophilic and phenolic compound with a log octanol/ water partition coefficient of 4.76. Due to its widespread use over the past 30 years, it is one of the most frequently detected compounds in wastewater effluent in USA [171], Canada [172], Europe [173], Australia [174], and Japan [175]. The detection of TCS in house dust reflects its widespread use in the domestic environment [72]. It has also been reported in sewage [176], human breast milk and in plasma [177]. The concentrations of TCS documented in different human biological matrices are presented in Table 8. The presence of TCS in sewage, human milk and plasma demonstrates systemic distribution in humans but local absorption from dermal application of cosmetics to the breast and use of treated undergarments are potential routes of exposure. A correlation has been reported between the levels of TCS in plasma and milk and the use of personal care products such as toothpaste, soap and underarm deodorants containing TCS [177]. The detection of TCS in human breast milk raises new questions of potential adverse effects to humans at a sensitive life stage and has implications for risk assessment for infants via oral ingestion [178]. Moreover, the presence of TCS in human milk also implies passage from the human breast and this raises concerns for its possible role in breast cancer development [179]. Although TCS has not been measured in human breast tissue, its presence in human milk implies its passage through the breast epithelial cells without metabolic alteration [180].

TCS has been reported to cause endocrine disruption and cyto-toxicity, to have weak estrogenic properties (agonist) and to possess estrogen antagonist activity at higher concentration [181]. In vitro effects at environmentally relevant concentrations (10~5 M) have further authenticated the intrinsic estrogenic and androgenic activity of TCS [180]. The endocrine activity of TCS at lower concentrations together with other adverse cellular actions of TCS including their ability to undergo photodegradation into dioxins [182] raises the question whether TCS might contribute to the

development of breast cancer. Clearance and/or persistence of TCS in human breast tissues from small regular inputs could add further to the estrogenic burden of cosmetic chemicals around the breast [179]. In a very recent in vitro study it was shown that exposure of breast cancer cells to TCS at the concentration range of 10~7 to 10~3 M, accelerate breast cancer cell growth through an ER-mediated pathway [38]. In addition the same study showed that TCS also up regulated cell cycle factors, cyclin D1 and p21 in both in vitro and in vivo models where rats were dosed at 100 mg/kg body weight.

A recent review on studies of human with regards to exposures to triclosan and cancer risk concluded that human studies are lacking in both number and scope [183]. It was particularly noted that epidemiologic studies of risk associated with exposure to tri-closan are needed.

3. Conclusion

Many chemicals interfere with normal, hormonally regulated biological processes and adversely affect development and/or reproductive functions by mimicking or inhibiting endogenous hormone action, modulating the production of endogenous hormones or altering hormone receptor populations. These endocrine disrupting chemicals exert their effects via nuclear receptors, membrane receptors, neurotransmitter receptors, and enzymatic pathways involved in the synthesis of hormones. Exposure to chemicals during windows of susceptibility can alter mammary gland structure, development and function, thereby increasing cancer risk. Thereby, these factors warrant the need to better understand the complex relationship between exposure to endocrine disrupting chemicals and the alterations in mammary gland morphology and gene expression that ultimately increase disease risk. Multiple combinations of one or more chemicals can yield functional significance; furthermore, the influence of time on enabling estrogenic responses from low doses of the chemicals, is an important factor relevant to environmental stimulation. In dose-response studies, higher doses of chemicals might amplify tumor growth, but dose alone would not be the sole determinant as different breast tumors develop and grow at different rates. These numerous studies on chemicals and their influence on the development of the disease, have implications for future considerations of environmental influences in breast cancer development as environmental compounds with estrogenic activity have been measured in human breast tissue and other biological matrices. The use of consumer products varies for each individual, so each will have a different body burdens with mixtures of different chemicals which may have a common mechanism of action and subsequently functional relevance. Furthermore, multiple compounds may be circulating and also present in the human breast at low levels, and their combined levels could be capable in the long term of producing estrogenic stimulus. Thus, a more holistic model of total estrogenic burden in the human breast is needed, in which total xenoestrogen burden is considered, not in isolation but in the context of levels of physiological hormones. Most epidemiologic studies report levels of one or one type of chemical at a time and yet other factors such as age, lifestyle, geographical location or genetic background may lead to differences in individual chemical loadings which can mask causative factors with a common mechanism of action such as estrogenic activity. Thus, measurements of multiple chemicals within a human breast (or a representative biological matrix) and its complex interactions with endogenous hormones has to be considered which may together alter endocrine homeo-stasis in the human breast with the potential to lead to development of cancer.

The exposure data compiled in this review shows substantial

variability in contaminant concentrations. This variability can be attributed to a variety of factors, including: sample matrix (i.e., serum vs urine), year of sample collection, geographical location, occupational or non-occupational exposure, etc. Furthermore, differences in the chemical/physical properties, toxicokinetics and toxicodynamics of these contaminants will be reflected in their presence and concentrations in various biological matrices. The conventional approach of assessing external exposure has been by environmental monitoring, however, there is a growing interest and focus in evaluating integrated exposure to environmental chemicals using human biomonitoring (HBM) to measure the parent compounds or their metabolites in human tissues or specimens. 1deally, to conduct a proper HBM study, it is important to: a) choose an appropriate biomarker which will reflect the exposure to the parent compound, b) select a suitable biological matrix, and c) choose a sensitive and specific biomarker of exposure for the chemical of interest. The most commonly used matrices in HBM studies are blood (serum/plasma/whole blood) and urine. However, the main advantage of using urine as a desired matrix is that it is non-invasive and samples can be accessible in large volumes. The routes (oral, inhalation, dermal) of exposure in the general population to the chemicals can be quiet varied, and sources can range from dietary sources to contaminants in household air or dust. These exposures will vary by individual and susceptible subgroups such as infants and children, so it is important to conduct biological monitoring and comprehensive exposures assessment studies in these groups to identify trends (ideally reductions) in exposures over time.

1n developed countries there has been a decrease in common environmental contaminants (e.g. heavy metals, DDT, Dioxins, PCBs), because of a concerted effort of strict regulation, improved monitoring, cleaner industrial processes and increased public awareness. As there is a decline in the levels of common (legacy) compounds, the focus has shifted to the chemicals that are mainly synthetic by nature and produced primarily to replace phased-out chemicals. Unlike common contaminants, 'emerging chemical contaminants' mostly find their way into the environment via diffuse sources (i.e., domestic, commercial, and industrial uses). Moreover, the development of more sensitive and new analytical capabilities allows scientists to identify contaminants which are typically present in ultra-low concentrations (parts per billion to parts per trillion). Exposure in-utero and in early life, trans-generational effects and the concerted play between the low-dose mechanistic effects of chemical cocktail in the environment and the vulnerable population who are predisposed to cancer, be it either by genetics or other lifestyle influences, must be considered. Still a significant amount of work needs to be done to assess the importance of these emerging chemicals in the environment with breast cancer risk and public health significance. 1n this review, the toxicology and available human epidemiology of some important emerging environmental contaminants have been summarized. The toxicological evidence indicates that these compounds may increase the risk of cancer, and specifically breast cancer, but the human evidence is lacking in part due to the paucity of epidemi-ologic data. Given their widespread use, multiple sources of human exposures and persistence in the environment, there is an urgent need to conduct a biomarker-based research in human populations to help identify potentially widespread population health risks, including breast cancer.

Acknowledgments

This project was funded by Canadian Cancer Society Grant # 700865. The authors wish to thank Ivana Kosarac and Xinghua Fan at Health Canada for their review of the draft manuscript.

References

[1] M.P. Curado, Breast cancer in the world: incidence and mortality, Salud publica Mex. 53 (2011) 372-384.

[2] ACS, American, Cancer Society. Breast Cancer Facts and Figures 2013-2014, American Cancer Society, Atlanta, 2013.

[3] CBCF, in.

[4] A.N. Monteiro, BRCA1: exploring the links to transcription, Trends Biochem. Sci. 25 (2000) 469-474.

[5] N. Mavaddat, A.C. Antoniou, D.F. Easton, M. Garcia-Closas, Genetic susceptibility to breast cancer, Mol. Oncol. 4 (2010) 174-191.

[6] R. Kaaks, S. Rinaldi, T.J. Key, F. Berrino, P.H. Peeters, C. Biessy, L. Dossus, A. Lukanova, S. Bingham, K.T. Khaw, N.E. Allen, H.B. Bueno-de-Mesquita, C.H. van Gils, D. Grobbee, H. Boeing, P.H. Lahmann, G. Nagel, J. Chang-Claude, F. Clavel-Chapelon, A. Fournier, A. Thiebaut, C.A. Gonzalez, J.R. Quiros, M.J. Tormo, E. Ardanaz, P. Amiano, V. Krogh, D. Palli, S. Panico, R. Tumino, P. Vineis, A. Trichopoulou, V. Kalapothaki, D. Trichopoulos, P. Ferrari, T. Norat, R. Saracci, E. Riboli, Postmenopausal serum androgens, oestrogens and breast cancer risk: the European prospective investigation into cancer and nutrition, Endocrine-related cancer 12 (2005) 1071-1082.

[7] J. Russo, I.H. Russo, The role of estrogen in the initiation of breast cancer, J. steroid Biochem. Mol. Biol. 102 (2006) 89-96.

[8] J.L. Kelsey, M.D. Gammon, E.M. John, Reproductive factors and breast cancer, Epidemiol. Rev. 15 (1993) 36-47.

[9] L. Bernstein, R.K. Ross, Endogenous hormones and breast cancer risk, Epidemiol. Rev. 15 (1993) 48-65.

[10] A. Howell, A.H. Sims, K.R. Ong, M.N. Harvie, D.G. Evans, R.B. Clarke, Mechanisms of Disease: prediction and prevention of breast cancer-cellular and molecular interactions, Nat. Clin. Pract. Oncol. 2 (2005) 635-646.

[11] J. Weyandt, R.E. Ellsworth, J.A. Hooke, C.D. Shriver, D.L. Ellsworth, Environmental chemicals and breast cancer risk-a structural chemistry perspective, Curr. Med. Chem. 15 (2008) 2680-2701.

[12] S.M. Snedeker, Pesticides and breast cancer risk: a review of DDT, DDE, and dieldrin, Environ. health Perspec. 1 (109 Suppl) (2001) 35-47.

[13] P.K. Mandal, Dioxin: a review of its environmental effects and its aryl hydrocarbon receptor biology, J. comp. Phys. B, Biochem. Syst. Environ. physiology 175 (2005) 221-230.

[14] E. Negri, C. Bosetti, E. Fattore, C. La Vecchia, Environmental exposure to polychlorinated biphenyls (PCBs) and breast cancer: a systematic review of the epidemiological evidence, Eur. J. cancer Prev. official J. Eur. Cancer Prev. Organ. 12 (2003) 509-516.

[15] J.A. McLachlan, Environmental signaling: what embryos and evolution teach us about endocrine disrupting chemicals, Endocr. Rev. 22 (2001) 319-341.

[16] J.D. Yager, Endogenous estrogens as carcinogens through metabolic activation, J. Natl. Cancer Inst. Monogr. (2000) 67-73.

[17] S.M. Choi, S.D. Yoo, B.M. Lee, Toxicological characteristics of endocrine-disrupting chemicals: developmental toxicity, carcinogenicity, and mutage-nicity, Journal of toxicology and environmental health. Part B, Crit. Rev. 7 (2004) 1-24.

[18] K. Yoon, S.J. Kwack, H.S. Kim, B.M. Lee, Estrogenic endocrine-disrupting chemicals: molecular mechanisms of actions on putative human diseases, Journal of toxicology and environmental health. Part B, Crit. Rev. 17 (2014) 127-174.

[19] W.H. Goodson 3rd, L. Lowe, D.O. Carpenter, M. Gilbertson, A. Manaf Ali, A. Lopez de Cerain Salsamendi, A. Lasfar, A. Carnero, A. Azqueta, A. Amedei, A.K. Charles, A.R. Collins, A. Ward, A.C. Salzberg, A. Colacci, A.K. Olsen, A. Berg, B.J. Barclay, B.P. Zhou, C. Blanco-Aparicio, C.J. Baglole, C. Dong,

C. Mondello, C.W. Hsu, C.C. Naus, C. Yedjou, C.S. Curran, D.W. Laird, D.C. Koch,

D.J. Carlin, D.W. Felsher, D. Roy, D.G. Brown, E. Ratovitski, E.P. Ryan,

E. Corsini, E. Rojas, E.Y. Moon, E. Laconi, F. Marongiu, F. Al-Mulla,

F. Chiaradonna, F. Darroudi, F.L. Martin, F.J. Van Schooten, G.S. Goldberg,

G. Wagemaker, G.N. Nangami, G.M. Calaf, G. Williams, G.T. Wolf, G. Koppen,

G. Brunborg, H.K. Lyerly, H. Krishnan, H. Ab Hamid, H. Yasaei, H. Sone,

H. Kondoh, H.K. Salem, H.Y. Hsu, H.H. Park, I. Koturbash, I.R. Miousse, A.I. Scovassi, J.E. Klaunig, J. Vondracek, J. Raju, J. Roman, J.P. Wise Sr., J.R. Whitfield, J. Woodrick, J.A. Christopher, J. Ochieng, J.F. Martinez-Leal, J. Weisz, J. Kravchenko, J. Sun, K.R. Prudhomme, K.B. Narayanan, K.A. Cohen-Solal, K. Moorwood, L. Gonzalez, L. Soucek, L. Jian, L.S. D'Abronzo, L.T. Lin, L. Li, L. Gulliver, L.J. McCawley, L. Memeo, L. Vermeulen, L. Leyns, L. Zhang, M. Valverde, M. Khatami, M.F. Romano, M. Chapellier, M.A. Williams, M. Wade, M.H. Manjili, M.E. Lleonart, M. Xia, M.J. Gonzalez, M.V. Karamouzis, M. Kirsch-Volders, M. Vaccari, N.B. Kuemmerle, N. Singh, N. Cruickshanks, N. Kleinstreuer, N. van Larebeke, N. Ahmed, O. Ogunkua, P.K. Krishnakumar, P. Vadgama, P.A. Marignani, P.M. Ghosh, P. Ostrosky-Wegman, P.A. Thompson, P. Dent, P. Heneberg, P. Darbre, P. Sing Leung, P. Nangia-Makker, Q.S. Cheng, R.B. Robey, R. Al-Temaimi, R. Roy, R. Andrade-Vieira, R.K. Sinha, R. Mehta, R. Vento, R. Di Fiore, R. Ponce-Cusi, R. Dornetshuber-Fleiss, R. Nahta, R.C. Castellino, R. Palorini, R. Abd Hamid, S.A. Langie, S.E. Eltom, S.A. Brooks, S. Ryeom, S.S. Wise, S.N. Bay, S.A. Harris, S. Papagerakis, S. Romano, S. Pavanello, S. Eriksson, S. Forte, S.C. Casey, S. Luanpitpong, T.J. Lee, T. Otsuki, T. Chen, T. Massfelder, T. Sanderson, T. Guarnieri, T. Hultman, V. Dormoy, V. Odero-Marah, V. Sabbisetti, V. Maguer-Satta, W.K. Rathmell, W. Engstrom, W.K. Decker, W.H. Bisson, Y. Rojanasakul, Y. Luqmani, Z. Chen, Z. Hu, Assessing the carcinogenic

potential of low-dose exposures to chemical mixtures in the environment: the challenge ahead, Carcinogenesis 1 (36 Suppl) (2015) S254—S296.

[20] NORMAN, in, 2015.

[21] USEPA, Contaminants of Emerging Concern, 2013.

[22] R.W. Klaper, L.C, in: S. Campbell (Ed.), Emerging Contaminant Threats and the Great Lakes: Existing Science, Estimating Relative Risk and Determining Policies, 2011.

[23] G.G. Ying, Fate, behavior and effects of surfactants and their degradation products in the environment, Environ. Int. 32 (2006) 417—431.

[24] K. Guenther, V. Heinke, B. Thiele, E. Kleist, H. Prast, T. Raecker, Endocrine disrupting nonylphenols are ubiquitous in food, Environ. Sci. Technol. 36 (2002) 1676—1680.

[25] R. White, S. Jobling, S.A. Hoare, J.P. Sumpter, M.G. Parker, Environmentally persistent alkylphenolic compounds are estrogenic, Endocrinology 135 (1994) 175—182.

[26] L.B. Clark, R.T. Rosen, T.G. Hartman, J.B. Louis, I. Suffet, R. Lippincott, J.D. Rosen, Determination of alkylphenol ethoxylates and their acetic acid derivatives in drinking water by particle beam liquid chromatography/mass spectrometry, Int. J. Environ. Anal. Chem. 47 (1992) 167—180.

[27] F.H. Kayama, H. Hamamatsu, A, Potential health effects of alkylphenols in Japan, Jpn. Med. Assoc. J. 46 (2003) 108—114.

[28] W. Dekant, W. Volkel, Human exposure to bisphenol A by biomonitoring: methods, results and assessment of environmental exposures, Toxicol. Appl. Pharmacol. 228 (2008) 114—134.

[29] M. Choi, H.-B. Moon, J. Yu, S.-S. Kim, A.S. Pait, H.-G. Choi, Nationwide monitoring of nonylphenolic compounds and coprostanol in sediments from Korean coastal waters, Mar. Pollut. Bull. 58 (2009) 1086—1092.

[30] M. Petrovic, E. Eljarrat, M.J. Lopez De Alda, D. Barcelo, Endocrine disrupting compounds and other emerging contaminants in the environment: a survey on new monitoring strategies and occurrence data, Anal. Bioanal. Chem. 378 (2004) 549—562.

[31] M.R. Taylor, P.T. Harrison, Ecological effects of endocrine disruption: current evidence and research priorities, Chemosphere 39 (1999) 1237—1248.

[32] R. Acevedo, P.G. Parnell, H. Villanueva, L.M. Chapman, T. Gimenez, S.L. Gray, W.S. Baldwin, The contribution of hepatic steroid metabolism to serum estradiol and estriol concentrations in nonylphenol treated MMTVneu mice and its potential effects on breast cancer incidence and latency, J. Appl. Toxicol. JAT 25 (2005) 339—353.

[33] A. Soares, B. Guieysse, B. Jefferson, E. Cartmell, J.N. Lester, Nonylphenol in the environment: a critical review on occurrence, fate, toxicity and treatment in wastewaters, Environ. Int. 34 (2008) 1033—1049.

[34] Y. Tabira, M. Nakai, D. Asai, Y. Yakabe, Y. Tahara, T. Shinmyozu, M. Noguchi, M. Takatsuki, Y. Shimohigashi, Structural requirements of para-alkylphenols to bind to estrogen receptor, Eur. J. Biochem./FEBS 262 (1999) 240—245.

[35] M. Isidori, M. Cangiano, F.A. Palermo, A. Parrella, E-screen and vitellogenin assay for the detection of the estrogenic activity of alkylphenols and trace elements, Comparative biochemistry and physiology, Toxicol. Pharmacol. CBP 152 (2010) 51—56.

[36] S. Meier, T.E. Andersen, B. Norberg, A. Thorsen, G.L. Taranger, O.S. Kjesbu, R. Dale, H.C. Morton, J. Klungsoyr, A. Svardal, Effects of alkylphenols on the reproductive system of Atlantic cod (Gadus morhua), Aquat. Toxicol. 81 (2007) 207—218.

[37] L. Pisapia, G. Del Pozzo, P. Barba, L. Caputo, L. Mita, E. Viggiano, G.L. Russo, C. Nicolucci, S. Rossi, U. Bencivenga, D.G. Mita, N. Diano, Effects of some endocrine disruptors on cell cycle progression and murine dendritic cell differentiation, General Comp. Endocrinol. 178 (2012) 54—63.

[38] H.R. Lee, K.A. Hwang, K.H. Nam, H.C. Kim, K.C. Choi, Progression of breast cancer cells was enhanced by endocrine-disrupting chemicals, triclosan and octylphenol, via an estrogen receptor-dependent signaling pathway in cellular and mouse xenograft models, Chem. Res. Toxicol. 27 (2014) 834—842.

[39] H. Ajj, A. Chesnel, S. Pinel, F. Plenat, S. Flament, H. Dumond, An alkylphenol mix promotes seminoma derived cell proliferation through an ERalpha36-mediated mechanism, PloS one 8 (2013) e61758.

[40] U.S.E.P.A. (USEPA), Bisphenol a (BPA) Action Plan Summary, 2013.

[41] T. Geens, T.Z. Apelbaum, L. Goeyens, H. Neels, A. Covaci, Intake of bisphenol A from canned beverages and foods on the Belgian market, Food additives & contaminants. Part A, Chem. analysis, control, Expo. risk Assess. 27 (2010) 1627—1637.

[42] A. Goodson, H. Robin, W. Summerfield, I. Cooper, Migration of bisphenol A from can coatings—effects of damage, storage conditions and heating, Food Addit. Contam. 21 (2004) 1015—1026.

[43] C. Kubwabo, I. Kosarac, B. Stewart, B.R. Gauthier, K. Lalonde, P.J. Lalonde, Migration of bisphenol A from plastic baby bottles, baby bottle liners and reusable polycarbonate drinking bottles, Food Addit. Contam. Part A, Chem. analysis, control, Expo. risk Assess. 26 (2009) 928—937.

[44] A. Matsumoto, N. Kunugita, K. Kitagawa, T. Isse, T. Oyama, G.L. Foureman, M. Morita, T. Kawamoto, Bisphenol A levels in human urine, Environ. health Perspect. 111 (2003) 101—104.

[45] Y.B. Wetherill, B.T. Akingbemi, J. Kanno, J.A. McLachlan, A. Nadal, C. Sonnenschein, C.S. Watson, R.T. Zoeller, S.M. Belcher, In vitro molecular mechanisms of bisphenol A action, Reprod. Toxicol. 24 (2007) 178—198.

[46] C.A. Richter, L.S. Birnbaum, F. Farabollini, R.R. Newbold, B.S. Rubin, C.E. Talsness, J.G. Vandenbergh, D.R. Walser-Kuntz, F.S. vom Saal, In vivo effects of bisphenol A in laboratory rodent studies, Reprod. Toxicol. 24 (2007)

Contaminants xxx (2016) 1—16 199—224.

[47] K.L. Christensen, M. Lorber, X. Ye, A.M. Calafat, Reconstruction of bisphenol A intake using a simple pharmacokinetic model, J. Expo. Sci. Environ. Epidemiol. 25 (2015) 240—248.

[48] G.G. Kuiper, J.G. Lemmen, B. Carlsson, J.C. Corton, S.H. Safe, P.T. van der Saag, B. van der Burg, J.A. Gustafsson, Interaction of estrogenic chemicals and phytoestrogens with estrogen receptor beta, Endocrinology 139 (1998) 4252—4263.

[49] W.V. Welshons, S.C. Nagel, F.S. vom Saal, Large effects from small exposures. III. Endocrine mechanisms mediating effects of bisphenol A at levels of human exposure, Endocrinology 147 (2006) S56—S69.

[50] J.S. Lakind, J. Levesque, P. Dumas, S. Bryan, J. Clarke, D.Q. Naiman, Comparing United States and Canadian population exposures from National Biomonitoring Surveys: bisphenol A intake as a case study, J. Expo. Sci. Environ. Epidemiol. 22 (2012) 219—226.

[51] J.G. Teeguarden, A.M. Calafat, X. Ye, D.R. Doerge, M.I. Churchwell, R. Gunawan, M.K. Graham, Twenty-four hour human urine and serum profiles of bisphenol a during high-dietary exposure, Toxicol. Sci. official J. Soc. Toxicol. 123 (2011)48—57.

[52] R.E. Chapin, J. Adams, K. Boekelheide, L.E. Gray Jr., S.W. Hayward, P.S. Lees,

B.S. McIntyre, K.M. Portier, T.M. Schnorr, S.G. Selevan, J.G. Vandenbergh, S.R. Woskie, NTP-CERHR expert panel report on the reproductive and developmental toxicity of bisphenol A, Birth defects research. Part B, Dev. reproductive Toxicol. 83 (2008) 157—395.

[53] T.A. Patterson, N.C. Twaddle, C.S. Roegge, R.J. Callicott, J.W. Fisher, D.R. Doerge, Concurrent determination of bisphenol A pharmacokinetics in maternal and fetal rhesus monkeys, Toxicol. Appl. Pharmacol. 267 (2013) 41 —48.

[54] W. Volkel, N. Bittner, W. Dekant, Quantitation of bisphenol A and bisphenol A glucuronide in biological samples by high performance liquid chromatography-tandem mass spectrometry, Drug metabolism Dispos. Biol. fate Chem. 33 (2005) 1748—1757.

[55] N.J. Cabaton, C. Canlet, P.R. Wadia, M. Tremblay-Franco, R. Gautier, J. Molina,

C. Sonnenschein, J.P. Cravedi, B.S. Rubin, A.M. Soto, D. Zalko, Effects of low doses of bisphenol A on the metabolome of perinatally exposed CD-1 mice, Environ. health Perspect. 121 (2013) 586—593.

[56] J.A. Taylor, F.S. Vom Saal, W.V. Welshons, B. Drury, G. Rottinghaus, P.A. Hunt, P.L. Toutain, C.M. Laffont, C.A. VandeVoort, Similarity of bisphenol A phar-macokinetics in rhesus monkeys and mice: relevance for human exposure, Environ. health Perspect. 119 (2011) 422—430.

[57] M. Nishikawa, H. Iwano, R. Yanagisawa, N. Koike, H. Inoue, H. Yokota, Placental transfer of conjugated bisphenol A and subsequent reactivation in the rat fetus, Environ. health Perspect. 118 (2010) 1196—1203.

[58] M.S. Golub, K.L. Wu, F.L. Kaufman, L.H. Li, F. Moran-Messen, L. Zeise, G.V. Alexeeff, J.M. Donald, Bisphenol A: developmental toxicity from early prenatal exposure, Birth defects research. Part B, Dev. reproductive Toxicol. 89 (2010) 441—466.

[59] L.N. Vandenberg, M.V. Maffini, C. Sonnenschein, B.S. Rubin, A.M. Soto, Bisphenol-A and the great divide: a review of controversies in the field of endocrine disruption, Endocr. Rev. 30 (2009) 75—95.

[60] H.B. Adewale, W.N. Jefferson, R.R. Newbold, H.B. Patisaul, Neonatal bisphenol-a exposure alters rat reproductive development and ovarian morphology without impairing activation of gonadotropin-releasing hormone neurons, Biol. reproduction 81 (2009) 690—699.

[61] L.N. Vandenberg, M.V. Maffini, P.R Wadia, C. Sonnenschein, B.S. Rubin, A.M. Soto, Exposure to environmentally relevant doses of the xenoestrogen bisphenol-A alters development of the fetal mouse mammary gland, Endocrinology 148 (2007) 116—127.

[62] A.M. Betancourt, I.A. Eltoum, R.A. Desmond, J. Russo, C.A. Lamartiniere, In utero exposure to bisphenol A shifts the window of susceptibility for mammary carcinogenesis in the rat, Environ. health Perspect. 118 (2010) 1614—1619.

[63] K. Weber Lozada, R.A. Keri, Bisphenol A increases mammary cancer risk in two distinct mouse models of breast cancer, Biol. reproduction 85 (2011) 490—497.

[64] H. Gao, B.J. Yang, N. Li, L.M. Feng, X.Y. Shi, W.H. Zhao, S.J. Liu, Bisphenol A and hormone-associated cancers: current progress and perspectives, Medicine 94 (2015) e211.

[65] B.L. Sprague, A. Trentham-Dietz, C.J. Hedman, J. Wang, J.D. Hemming, J.M. Hampton, D.S. Buist, E.J. Aiello Bowles, G.S. Sisney, E.S. Burnside, Circulating serum xenoestrogens and mammographic breast density, Breast cancer Res. BCR 15 (2013) R45.

[66] M. Yang, J.H. Ryu, R. Jeon, D. Kang, K.Y. Yoo, Effects of bisphenol A on breast cancer and its risk factors, Archives Toxicol. 83 (2009) 281—285.

[67] B. Trabert, R.T. Falk, J.D. Figueroa, B.I. Graubard, M. Garcia-Closas, J. Lissowska, B. Peplonska, S.D. Fox, L.A. Brinton, Urinary bisphenol A-glucuronide and postmenopausal breast cancer in Poland, Cancer causes control CCC 25 (2014) 1587—1593.

[68] R. Golden, J. Gandy, G. Vollmer, A review of the endocrine activity of parabens and implications for potential risks to human health, Crit. Rev. Toxicol. 35 (2005) 435—458.

[69] K. Yazar, S. Johnsson, M.L. Lind, A. Boman, C. Liden, Preservatives and fragrances in selected consumer-available cosmetics and detergents, Contact Dermat. 64 (2011) 265—272.

[70] P.D. Darbre, J.R. Byford, L.E. Shaw, R.A. Horton, G.S. Pope, M.J. Sauer,

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

[71 [72

[73 [74

[75 [76 [77 [78

[82 [83

[88 [89

[91 [92

[93 [94 [95

Oestrogenic activity of isobutylparaben in vitro and in vivo, J. Appl. Toxicol. JAT 22 (2002) 219-226. [96

J.M. Brausch, G.M. Rand, A review of personal care products in the aquatic environment: environmental concentrations and toxicity, Chemosphere 82 [97 (2011) 1518-1532.

X. Fan, C. Kubwabo, P. Rasmussen, H. Jones-Otazo, Simultaneous quantitation [98 of parabens, triclosan, and methyl triclosan in indoor house dust using solid phase extraction and gas chromatography-mass spectrometry, J. Environ. Monit. JEM 12 (2010) 1891-1897.

L. Barr, G. Metaxas, C.A. Harbach, L.A. Savoy, P.D. Darbre, Measurement of [99 paraben concentrations in human breast tissue at serial locations across the breast from axilla to sternum, J. Appl. Toxicol. JAT 32 (2012) 219-232. M. Schlumpf, K. Kypke, M. Wittassek, J. Angerer, H. Mascher, D. Mascher, [100 C. Vokt, M. Birchler, W. Lichtensteiger, Exposure patterns of UV filters, fragrances, parabens, phthalates, organochlor pesticides, PBDEs, and PCBs in human milk: correlation of UV filters with use of cosmetics, Chemosphere 81 (2010) 1171-1183. [101

T.M. Sandanger, S. Huber, M.K. Moe, T. Braathen, H. Leknes, E. Lund, Plasma concentrations of parabens in postmenopausal women and self-reported use of personal care products: the NOWAC postgenome study, J. Expo. Sci. En- [102 viron. Epidemiol. 21 (2011) 595-600.

H. Frederiksen, N. Jorgensen, A.M. Andersson, Parabens in urine, serum and seminal plasma from healthy Danish men determined by liquid chromatography-tandem mass spectrometry (LC-MS/MS), J. Expo. Sci. Environ. Epidemiol. 21 (2011) 262-271. [103 S. Abbas, H. Greige-Gerges, N. Karam, M.H. Piet, P. Netter, J. Magdalou, Metabolism of parabens (4-hydroxybenzoic acid esters) by hepatic esterases and UDP-glucuronosyltransferases in man, Drug metabolism Pharmacokinet. 25 (2010) 568-577.

N. Aubert, T. Ameller, J.J. Legrand, Systemic exposure to parabens: pharma- [104 cokinetics, tissue distribution, excretion balance and plasma metabolites of [14C]-methyl-, propyl- and butylparaben in rats after oral, topical or subcutaneous administration, Food Chem. Toxicol. 1nt. J. Publ. Br. 1ndustrial Biol. Res. Assoc. 50 (2012) 445-454. [105

X. Ye, A.M. Bishop, J.A. Reidy, L.L. Needham, A.M. Calafat, Parabens as urinary biomarkers of exposure in humans, Environ. health Perspect. 114 (2006) 1843-1846.

Z. Dagher, M. Borgie, J. Magdalou, R. Chahine, H. Greige-Gerges, p-Hydrox- [106 ybenzoate esters metabolism in MCF7 breast cancer cells, Food Chem. Tox-icol. Int. J. Publ. Br. Industrial Biol. Res. Assoc. 50 (2012) 4109-4114. S. El Hussein, P. Muret, M. Berard, S. Makki, P. Humbert, Assessment of [107 principal parabens used in cosmetics after their passage through human epidermis-dermis layers (ex-vivo study), Exp. Dermatol 16 (2007) 830-836. P. Darbre, A. Aljarrah, W. Miller, N. Coldham, M. Sauer, G. Pope, Concentrations of parabens in human breast tumours, J. Appl. Toxicol. 24 (2004) 5-13. [108 L. Sun, T. Yu, J. Guo, Z. Zhang, Y. Hu, X. Xiao, Y. Sun, H. Xiao, J. Li, D. Zhu, L. Sai, J. Li, The estrogenicity of methylparaben and ethylparaben at doses close to the acceptable daily intake in immature Sprague-Dawley rats, Sci. Rep. 6 (2016) 25173. [109

M. 1shidate Jr., M. Hayashi, M. Sawada, A. Matsuoka, K. Yoshikawa, M. Ono, M. Nakadate, Cytotoxicity test on medical drugs-chromosome aberration tests with Chinese hamster cells in vitro (author's transl), Eisei Shikenjo [110 hokoku. Bull. Natl. Inst. Hyg. Sci. (1977) 55-61.

M.M. Mason, C. Cate, J. Baker, Toxicology and carcinogenesis of various chemicals used in the preparation of vaccines, Clin. Toxicol. 4 (1971) [111 185-204.

Y. Nakagawa, P. Moldeus, Mechanism of p-hydroxybenzoate ester-induced [112 mitochondrial dysfunction and cytotoxicity in isolated rat hepatocytes, Biochem. Pharmacol. 55 (1998) 1907-1914. P.D. Darbre, J.R. Byford, L.E. Shaw, S. Hall, N.G. Coldham, G.S. Pope, M.J. Sauer, [113 Oestrogenic activity of benzylparaben, J. Appl. Toxicol. JAT 23 (2003) 43-51. W.J. Crinnion, Toxic effects of the easily avoidable phthalates and parabens, Altern. Med. Rev. a J. Clin. Ther. 15 (2010) 190-196.

P.D. Darbre, P.W. Harvey, Paraben esters: review of recent studies of endo- [114 crine toxicity, absorption, esterase and human exposure, and discussion of potential human health risks, J. Appl. Toxicol. JAT 28 (2008) 561-578. P.W. Harvey, D.J. Everett, Regulation of endocrine-disrupting chemicals: [115 critical overview and deficiencies in toxicology and risk assessment for human health, Best Pract. Res. Clin. Endocrinol. metabolism 20 (2006) 145-165. [116

R.J. Witorsch, J.A. Thomas, Personal care products and endocrine disruption: a critical review of the literature, Crit. Rev. Toxicol. 3 (40 Suppl) (2010) 1-30. S. Dimitrov, V. Kamenska, J.D. Walker, W. Windle, R. Purdy, M. Lewis, O. Mekenyan, Predicting the biodegradation products of perfluorinated chemicals using CATABOL, SAR QSAR Environ. Res. 15 (2004) 69-82. [117

S. Rayne, K. Forest, Perfluoroalkyl sulfonic and carboxylic acids: a critical review of physicochemical properties, levels and patterns in waters and [118 wastewaters, and treatment methods, J. environ. sci. and health. Part A, Toxic/hazardous Subst. Environ. Eng. 44 (2009) 1145-1199. [119

P. Zareitalabad, J. Siemens, M. Hamer, W. Amelung, Perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) in surface waters, sediments, soils and wastewater-A review on concentrations and distribution coefficients, Chemosphere 91 (2013) 725-732. [120] C. Kubwabo, B. Stewart, J. Zhu, L. Marro, Occurrence of perfluorosulfonates and other perfluorochemicals in dust from selected homes in the city of

Ottawa, Canada, J. Environ. Monit. JEM 7 (2005) 1074-1078.

R. Vestergren, 1.T. Cousins, Tracking the pathways of human exposure to

perfluorocarboxylates, Environ. Sci. Technol. 43 (2009) 5565-5575.

J.P. Giesy, K. Kannan, Global distribution of perfluorooctane sulfonate in

wildlife, Environ. Sci. Technol. 35 (2001) 1339-1342.

M.E. Andersen, J.L. Butenhoff, S.C. Chang, D.G. Farrar, G.L. Kennedy Jr., C. Lau,

G.W. Olsen, J. Seed, K.B. Wallace, Perfluoroalkyl acids and related chemis-tries-toxicokinetics and modes of action, Toxicol. Sci. official J. Soc. Toxicol. 102 (2008) 3-14.

S. Fuentes, P. Vicens, M.T. Colomina, J.L. Domingo, Behavioral effects in adult mice exposed to perfluorooctane sulfonate (PFOS), Toxicology 242 (2007) 123-129.

N. Johansson, P. Eriksson, H. Viberg, Neonatal exposure to PFOS and PFOA in mice results in changes in proteins which are important for neuronal growth and synaptogenesis in the developing brain, Toxicol. Sci. official J. Soc. Toxicol. 108 (2009) 412-418.

L.J.A. Sibinski, J.L. Erickson, E.E., 3M Company/Riker Exp. No 0281CR0012 M., Two year oral (diet) toxicity/carcinogenicity study of fluorochemical FC-143 in rats, 3, 1987.

S.S. White, A.M. Calafat, Z. Kuklenyik, L. Villanueva, R.D. Zehr, L. Helfant, M.J. Strynar, A.B. Lindstrom, J.R. Thibodeaux, C. Wood, S.E. Fenton, Gesta-tional PFOA exposure of mice is associated with altered mammary gland development in dams and female offspring, Toxicol. Sci. official J. Soc. Tox-icol. 96 (2007) 133-144.

M.M. Peden-Adams, J.E. Stuckey, K.M. Gaworecki, J. Berger-Ritchie, K. Bryant, P.G. Jodice, T.R. Scott, J.B. Ferrario, B. Guan, C. Vigo, J.S. Boone, W.D. McGuinn, J.C. DeWitt, D.E. Keil, Developmental toxicity in white leghorn chickens following in ovo exposure to perfluorooctane sulfonate (PFOS), Reprod. Toxicol. 27 (2009) 307-318.

E.A. Emmett, H. Zhang, F.S. Shofer, D. Freeman, N.V. Rodway, C. Desai, L.M. Shaw, Community exposure to perfluorooctanoate: relationships between serum levels and certain health parameters, J. Occup. Environ. Med./ Am. Coll. Occup. Environ. Med. 48 (2006) 771-779.

M. Maras, C. Vanparys, F. Muylle, J. Robbens, U. Berger, J.L. Barber, R. Blust, W. De Coen, Estrogen-like properties of fluorotelomer alcohols as revealed by mcf-7 breast cancer cell proliferation, Environ. health Perspect. 114 (2006) 100-105.

C. Lau, K. Anitole, C. Hodes, D. Lai, A. Pfahles-Hutchens, J. Seed, Perfluoroalkyl acids: a review of monitoring and toxicological findings, Toxicol. Sci. official J. Soc. Toxicol. 99 (2007) 366-394.

E.C. Bonefeld-Jorgensen, M. Long, R. Bossi, P. Ayotte, G. Asmund, T. Kruger, M. Ghisari, G. Mulvad, P. Kern, P. Nzulumiki, E. Dewailly, Perfluorinated compounds are related to breast cancer risk in Greenlandic 1nuit: a case control study, Environ. health a Glob. access Sci. source 10 (2011) 88.

C. Kubwabo, P.E. Rasmussen, X. Fan, 1. Kosarac, F. Wu, A. Zidek, S.L. Kuchta, Analysis of selected phthalates in Canadian indoor dust collected using household vacuum and standardized sampling techniques, 1ndoor air 23

(2013) 506-514.

M. Wittassek, H.M. Koch, J. Angerer, T. Bruning, Assessing exposure to phthalates - the human biomonitoring approach, Mol. Nutr. food Res. 55 (2011) 7-31.

H.M. Koch, A.M. Calafat, Human body burdens of chemicals used in plastic manufacture, Philosophical transactions of the Royal Society of London. Series B, Biol. Sci. 364 (2009) 2063-2078.

H. Frederiksen, N.E. Skakkebaek, A.M. Andersson, Metabolism of phthalates in humans, Mol. Nutr. food Res. 51 (2007) 899-911.

F.P. Chen, M.H. Chien, Lower concentrations of phthalates induce proliferation in human breast cancer cells, Climacteric J. 1nt. Menopause Soc. 17

(2014) 377-384.

Y. Kang, W. Den, H. Bai, F.H. Ko, Direct quantitative analysis of phthalate esters as micro-contaminants in cleanroom air and wafer surfaces by autothermal desorption-gas chromatography-mass spectrometry, J. Chromatogr. A 1070 (2005) 137-145.

I.Y. Kim, S.Y. Han, A. Moon, Phthalates inhibit tamoxifen-induced apoptosis in MCF-7 human breast cancer cells, J. Toxicol. Environ. health. Part A 67 (2004) 2025-2035.

Y.Y. Chou, P.C. Huang, C.C. Lee, M.H. Wu, S.J. Lin, Phthalate exposure in girls during early puberty, J. Pediatr. Endocrinol. metabolism JPEM 22 (2009) 69-77.

M.S. Wolff, S.L. Teitelbaum, S.M. Pinney, G. Windham, L. Liao, F. Biro, L.H. Kushi, C. Erdmann, R.A. Hiatt, M.E. Rybak, A.M. Calafat, C. Breast, C. Environment Research, 1nvestigation of relationships between urinary biomarkers of phytoestrogens, phthalates, and phenols and pubertal stages in girls, Environ. health Perspect. 118 (2010) 1039-1046. N.P. Moore, The oestrogenic potential of the phthalate esters, Reprod. Tox-icol. 14 (2000) 183-192.

S.C. Kang, B.M. Lee, DNA methylation of estrogen receptor alpha gene by phthalates, J. Toxicol. Environ. health. Part A 68 (2005) 1995-2003. T.H. Hsieh, C.F. Tsai, C.Y. Hsu, P.L. Kuo, J.N. Lee, C.Y. Chai, S.C. Wang, E.M. Tsai, Phthalates induce proliferation and invasiveness of estrogen receptor-negative breast cancer through the AhR/HDAC6/c-Myc signaling pathway, FASEB J. official Publ. Fed. Am. Soc. Exp. Biol. 26 (2012) 778-787. T.R. Zacharewski, M.D. Meek, J.H. Clemons, Z.F. Wu, M.R. Fielden, J.B. Matthews, Examination of the in vitro and in vivo estrogenic activities of eight commercial phthalate esters, Toxicol. Sci. official J. Soc. Toxicol. 46

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

(1998) 282-293.

T. Lovekamp-Swan, B.J. Davis, Mechanisms of phthalate ester toxicity in the female reproductive system, Environ. health Perspect. 111 (2003) 139-145. R. Moral, J. Santucci-Pereira, R. Wang, I.H. Russo, C.A. Lamartiniere, J. Russo, In utero exposure to butyl benzyl phthalate induces modifications in the morphology and the gene expression profile of the mammary gland: an experimental study in rats, Environ. health a Glob. access Sci. source 10 (2011) 5.

E.J. Hong, Y.K. Ji, K.C. Choi, N. Manabe, E.B. Jeung, Conflict of estrogenic activity by various phthalates between in vitro and in vivo models related to the expression of Calbindin-D9k, J. reproduction Dev. 51 (2005) 253-263. L. Lopez-Carrillo, R.U. Hernandez-Ramirez, A.M. Calafat, L. Torres-Sanchez, M. Galvan-Portillo, L.L. Needham, R. Ruiz-Ramos, M.E. Cebrian, Exposure to phthalates and breast cancer risk in northern Mexico, Environ. health Per-spect. 118 (2010) 539-544.

N.H. Kleinsasser, E.R. Kastenbauer, H. Weissacher, R.K. Muenzenrieder, U.A. Harreus, Phthalates demonstrate genotoxicity on human mucosa of the upper aerodigestive tract, Environ. Mol. Mutagen. 35 (2000) 9-12. G.A. Martinez-Nava, A.I. Burguete-Garcia, L. Lopez-Carrillo, R.U. Hernandez-Ramirez, V. Madrid-Marina, M.E. Cebrian, PPARgamma and PPARGC1B polymorphisms modify the association between phthalate metabolites and breast cancer risk, Biomarkers Biochem. Indic. Expo. response, susceptibility Chem. 18 (2013) 493-501.

R. Hauser, J.D. Meeker, N.P. Singh, M.J. Silva, L. Ryan, S. Duty, A.M. Calafat, DNA damage in human sperm is related to urinary levels of phthalate monoester and oxidative metabolites, Hum. Reprod. 22 (2007) 688-695. J.T. Brophy, M.M. Keith, A. Watterson, R. Park, M. Gilbertson, E. Maticka-Tyndale, M. Beck, H. Abu-Zahra, K. Schneider, A. Reinhartz, R. Dematteo, I. Luginaah, Breast cancer risk in relation to occupations with exposure to carcinogens and endocrine disruptors: a Canadian case-control study, Environ. health a Glob. access Sci. source 11 (2012) 87. A. Aschengrau, P.F. Coogan, M. Quinn, L.J. Cashins, Occupational exposure to estrogenic chemicals and the occurrence of breast cancer: an exploratory analysis, Am. J. industrial Med. 34 (1998) 6-14.

P.O. Darnerud, G.S. Eriksen, T. Johannesson, P.B. Larsen, M. Viluksela, Poly-brominated diphenyl ethers: occurrence, dietary exposure, and toxicology, Environ. health Perspect. 1 (109 Suppl) (2001) 49-68. Brominated BSEF, Science and Environmental Forum, 2010. K. Prevedouros, K.C. Jones, A.J. Sweetman, Estimation of the production, consumption, and atmospheric emissions of pentabrominated diphenyl ether in Europe between 1970 and 2000, Environ. Sci. Technol. 38 (2004) 3224-3231.

J. Ward, S.P. Mohapatra, A. Mitchell, An overview of policies for managing polybrominated diphenyl ethers (PBDEs) in the Great Lakes basin, Environ. Int. 34 (2008) 1148-1156.

S.C. Secretariat, Listing of POPs in the Stockholm Convention, Annex a, 2009. Polybrominated USEPA, Diphenyl Ethers (PBDEs) Action Plan Summary, 2013.

European EU, Court of Justice Annuls Deca-BDE Exemption, 2008. J. Tan, S.M. Cheng, A. Loganath, Y.S. Chong, J.P. Obbard, Polybrominated diphenyl ethers in house dust in Singapore, Chemosphere 66 (2007) 985-992.

X. Qiu, R.M. Bigsby, R.A. Hites, Hydroxylated metabolites of polybrominated diphenyl ethers in human blood samples from the United States, Environ. health Perspect. 117 (2009) 93-98.

K. Hooper, T.A. McDonald, The PBDEs: an emerging environmental challenge and another reason for breast-milk monitoring programs, Environ. health Perspect. 108 (2000) 387-392.

I. Watanabe, R. Tatsukawa, Formation of brominated dibenzofurans from the photolysis of flame retardant decabromobiphenyl ether in hexane solution by UV and sun light, Bull. Environ. Contam. Toxicol. 39 (1987) 953-959.

G. Soderstrom, U. Sellstrom, C.A. de Wit, M. Tysklind, Photolytic debromi-nation of decabromodiphenyl ether (BDE 209), Environ. Sci. Technol. 38 (2004) 127-132.

P.O. Darnerud, S. Risberg, Tissue localisation of tetra- and pentabromodi-phenyl ether congeners (BDE-47, -85 and -99) in perinatal and adult C57BL mice, Chemosphere 62 (2006) 485-493.

E.K. Shanle, W. Xu, Endocrine disrupting chemicals targeting estrogen receptor signaling: identification and mechanisms of action, Chem. Res. Tox-icol. 24 (2011) 6-19.

K.G. Harley, A.R. Marks, J. Chevrier, A. Bradman, A. Sjodin, B. Eskenazi, PBDE concentrations in women's serum and fecundability, Environ. health Perspect. 118 (2010) 699-704.

T.A. McDonald, A perspective on the potential health risks of PBDEs, Che-mosphere 46 (2002) 745-755.

M.M. Dingemans, A. de Groot, R.G. van Kleef, A. Bergman, M. van den Berg,

H.P. Vijverberg, R.H. Westerink, Hydroxylation increases the neurotoxic potential of BDE-47 to affect exocytosis and calcium homeostasis in PC12 cells, Environ. health Perspect. 116 (2008) 637-643.

H. Lilienthal, A. Hack, A. Roth-Harer, S.W. Grande, C.E. Talsness, Effects of developmental exposure to 2,2 ,4,4 ,5-pentabromodiphenyl ether (PBDE-99) on sex steroids, sexual development, and sexually dimorphic behavior in rats, Environ. health Perspect. 114 (2006) 194-201.

A.K. Peters, S. Nijmeijer, K. Gradin, M. Backlund, A. Bergman, L. Poellinger, M.S. Denison, M. Van den Berg, Interactions of polybrominated diphenyl

161 162

ethers with the aryl hydrocarbon receptor pathway, Toxicol. Sci. official J. Soc. Toxicol. 92 (2006) 133—142.

J. Ukpebor, V. Llabjani, F.L. Martin, C.J. Halsall, Sublethal genotoxicity and cell alterations by organophosphorus pesticides in MCF-7 cells: implications for environmentally relevant concentrations, Environ. Toxicol. Chem./SETAC 30 (2011) 632—639.

L. Yu, P. Zhan, Molecular mechanisms underlying proliferation and apoptosis in breast cancer MCF-7 cells induced by pentabrominated diphenyl ethers, Toxicol. Environ. Chem. 91 (2009) 665—670.

Y.-H. Cui, P. Zhan, D. Luo, Y.-y. Xia, Molecular mechanisms underlying pen-tabrominated diphenyl ether-induced proliferation in breast cancer MCF-7 cells, Toxicol. Environ Chem. 92 (2010) 1177—1185.

C.E. Talsness, S.N. Kuriyama, A. Sterner-Kock, P. Schnitker, S.W. Grande, M. Shakibaei, A. Andrade, K. Grote, I. Chahoud, In utero and lactational exposures to low doses of polybrominated diphenyl ether-47 alter the reproductive system and thyroid gland of female rat offspring, Environ. health Perspect. 116 (2008) 308—314.

Z.H. Li, X.Y. Liu, N. Wang, J.S. Chen, Y.H. Chen, J.T. Huang, C.H. Su, F. Xie, B. Yu,

D.J. Chen, Effects ofdecabrominated diphenyl ether (PBDE-209) in regulation of growth and apoptosis of breast, ovarian, and cervical cancer cells, Environ. health Perspect. 120 (2012) 541—546.

S. Hurley, P. Reynolds, D. Goldberg, D.O. Nelson, S.S. Jeffrey, M. Petreas, Adipose levels of polybrominated diphenyl ethers and risk of breast cancer, Breast cancer Res. Treat. 129 (2011) 505—511.

A.K. Holmes, K.R. Koller, S.M. Kieszak, A. Sjodin, A.M. Calafat, F.D. Sacco, D.W. Varner, A.P. Lanier, C.H. Rubin, Case-control study of breast cancer and exposure to synthetic environmental chemicals among Alaska Native women, Int. J. circumpolar health 73 (2014) 25760.

J.L. Reiner, K. Kannan, A survey of polycyclic musks in selected household commodities from the United States, Chemosphere 62 (2006) 867—873. S. Lignell, P.O. Darnerud, M. Aune, S. Cnattingius, J. Hajslova, L. Setkova,

A. Glynn, Temporal trends of synthetic musk compounds in mother's milk and associations with personal use of perfumed products, Environ. Sci. Technol. 42 (2008) 6743—6748.

T. Kupper, J.D. Berset, R. Etter-Holzer, R. Furrer, J. Tarradellas, Concentrations and specific loads of polycyclic musks in sewage sludge originating from a monitoring network in Switzerland, Chemosphere 54 (2004) 1111 —1120. R. Salgado, J.P. Noronha, A. Oehmen, G. Carvalho, M.A. Reis, Analysis of 65 pharmaceuticals and personal care products in 5 wastewater treatment plants in Portugal using a simplified analytical methodology, Water Sci. Technol. a J. Int. Assoc. Water Pollut. Res. 62 (2010) 2862—2871.

C. Kubwabo, X. Fan, P.E. Rasmussen, F. Wu, Determination of synthetic musk compounds in indoor house dust by gas chromatography-ion trap mass spectrometry, Anal. Bioanal. Chem. 404 (2012) 467—477.

G.G. Rimkus, Polycyclic musk fragrances in the aquatic environment, Toxicol. Lett. 111 (1999) 37—56.

K. Kannan, J.L. Reiner, S.H. Yun, E.E. Perrotta, L. Tao, B. Johnson-Restrepo,

B.D. Rodan, Polycyclic musk compounds in higher trophic level aquatic organisms and humans from the United States, Chemosphere 61 (2005) 693—700.

T. Luckenbach, D. Epel, Nitromusk and polycyclic musk compounds as long-term inhibitors of cellular xenobiotic defense systems mediated by multi-drug transporters, Environ. health Perspect. 113 (2005) 17—24. EU, The European Commission bans musk xylene, 2011.

H. Nakata, Occurrence of synthetic musk fragrances in marine mammals and sharks from Japanese coastal waters, Environ. Sci. Technol. 39 (2005) 3430—3434.

S. Muller, P. Schmid, C. Schlatter, Occurrence of nitro and non-nitro benze-noid musk compounds in human adipose tissue, Chemosphere 33 (1996) 17—28.

G.G. Rimkus, M. Wolf, Polycyclic musk fragrances in human adipose tissue and human milk, Chemosphere 33 (1996) 2033—2043.

H.H. Schmeiser, R. Gminski, V. Mersch-Sundermann, Evaluation of health risks caused by musk ketone, Int. J. Hyg. Environ. health 203 (2001) 293—299.

N. Bitsch, C. Dudas, W. Korner, K. Failing, S. Biselli, G. Rimkus, H. Brunn, Estrogenic activity of musk fragrances detected by the E-screen assay using human mcf-7 cells, Archives Environ. Contam. Toxicol. 43 (2002) 257—264. W.C. Chau, J.L. Wu, Z. Cai, Investigation of levels and fate of triclosan in environmental waters from the analysis of gas chromatography coupled with ion trap mass spectrometry, Chemosphere 73 (2008) S13—S17.

D.W. Kolpin, E.T. Furlong, M.T. Meyer, E.M. Thurman, S.D. Zaugg, L.B. Barber, H.T. Buxton, Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams, 1999-2000: a national reconnaissance, Environ. Sci. Technol. 36 (2002) 1202—1211.

L. Lishman, S.A. Smyth, K. Sarafin, S. Kleywegt, J. Toito, T. Peart, B. Lee, M. Servos, M. Beland, P. Seto, Occurrence and reductions of pharmaceuticals and personal care products and estrogens by municipal wastewater treatment plants in Ontario, Canada, Sci. total Environ. 367 (2006) 544—558. D. Sabaliunas, S.F. Webb, A. Hauk, M. Jacob, W.S. Eckhoff, Environmental fate of triclosan in the river Aire basin, UK, Water Res. 37 (2003) 3145—3154. G.G. Ying, R.S. Kookana, Triclosan in wastewaters and biosolids from Australian wastewater treatment plants, Environ. Int. 33 (2007) 199—205. N. Nakada, T. Tanishima, H. Shinohara, K. Kiri, H. Takada, Pharmaceutical chemicals and endocrine disrupters in municipal wastewater in Tokyo and

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

180 181

200 201

their removal during activated sludge treatment, Water Res. 40 (2006) 3297-3303. [204

H.B. Lee, T.E. Peart, M.L. Svoboda, Determination of endocrine-disrupting phenols, acidic pharmaceuticals, and personal-care products in sewage by solid-phase extraction and gas chromatography-mass spectrometry, [205 J. Chromatogr. A 1094 (2005) 122-129.

M. Allmyr, M. Adolfsson-Erici, M.S. McLachlan, G. Sandborgh-Englund, Tri-closan in plasma and milk from Swedish nursing mothers and their exposure via personal care products, Sci. total Environ. 372 (2006) 87-93. [206

A.D. Dayan, Risk assessment of triclosan [Irgasan] in human breast milk, Food Chem. Toxicol. Int. J. Publ. Br. Industrial Biol. Res. Assoc. 45 (2007) 125-129.

P.D. Darbre, Environmental oestrogens, cosmetics and breast cancer, Best [207 practice & research, Clin. Endocrinol. metabolism 20 (2006) 121-143. R.H. Gee, A. Charles, N. Taylor, P.D. Darbre, Oestrogenic and androgenic activity of triclosan in breast cancer cells, J. Appl. Toxicol. JAT 28 (2008) 78-91. N.D. Henry, P.A. Fair, Comparison of in vitro cytotoxicity, estrogenicity and [208 anti-estrogenicity of triclosan, perfluorooctane sulfonate and per-fluorooctanoic acid, J. Appl. Toxicol. JAT 33 (2013) 265-272.

K. Aranami, J.W. Readman, Photolytic degradation of triclosan in freshwater [209 and seawater, Chemosphere 66 (2007) 1052-1056.

M.T. Dinwiddie, P.D. Terry, J. Chen, Recent evidence regarding triclosan and

cancer risk, Int. J. Environ. Res. public health 11 (2014) 2209-2217. [210

H.C. Environment Canada, Priority substances list assessment report, 2005.

P. R.J.B, Man made chemicals in Maternal and cord blood, 2005.

M.-L. Chen, W.-P. Lee, H.-Y. Chung, B.-R. Guo, I.-F. Mao, Biomonitoring of

alkylphenols exposure for textile and housekeeping workers, Int. J. Environ.

Anal. Chem. 85 (2005) 335-347. [211

B.L. Tan, M. Ali Mohd, Analysis of selected pesticides and alkylphenols in human cord blood by gas chromatograph-mass spectrometer, Talanta 61 (2003) 385-391.

E. Puy-Azurmendi, M. Ortiz-Zarragoitia, M. Villagrasa, M. Kuster, P. Aragon, [212 J. Atienza, R. Puchades, A. Maquieira, C. Dominguez, M. Lopez de Alda, D. Fernandes, C. Porte, J.M. Bayona, D. Barcelo, M.P. Cajaraville, Endocrine disruption in thicklip grey mullet (Chelon labrosus) from the urdaibai biosphere reserve (Bay of biscay, Southwestern Europe), Sci. total Environ. 443 (2013) 233-244. [213

N. Ademollo, F. Ferrara, M. Delise, F. Fabietti, E. Funari, Nonylphenol and octylphenol in human breast milk, Environ. Int. 34 (2008) 984-987. [214

H. Otaka, A. Yasuhara, M. Morita, Determination of bisphenol A and 4-non-ylphenol in human milk using alkaline digestion and cleanup by solid-phase extraction, Anal. Sci. Int. J. Jpn. Soc. Anal. Chem. 19 (2003) 1663-1666. G.W. Chen, W.H. Ding, H.Y. Ku, H.R. Chao, H.Y. Chen, M.C. Huang, S.L. Wang, [215 Alkylphenols in human milk and their relations to dietary habits in central Taiwan, Food Chem. Toxicol. Int. J. Publ. Br. Industrial Biol. Res. Assoc. 48 (2010) 1939-1944.

C. Sriphrapradang, L.O. Chailurkit, W. Aekplakorn, B. Ongphiphadhanakul, [216 Association between bisphenol A and abnormal free thyroxine level in men, Endocrine 44 (2013) 441-447.

Y. He, M. Miao, L.J. Herrinton, C. Wu, W. Yuan, Z. Zhou, D.K. Li, Bisphenol A levels in blood and urine in a Chinese population and the personal factors [217 affecting the levels, Environ. Res. 109 (2009) 629-633. W. Aekplakorn, L.O. Chailurkit, B. Ongphiphadhanakul, Relationship of serum bisphenol A with diabetes in the thai population, national health examina- [218 tion survey IV, 2009, J. diabetes 7 (2015) 240-249.

T. Bushnik, D. Haines, P. Levallois, J. Levesque, J. Van Oostdam, C. Viau, Lead and bisphenol A concentrations in the Canadian population, Health Rep. 21 (2010) 7-18.

X. Ye, F.H. Pierik, J. Angerer, H.M. Meltzer, V.W. Jaddoe, H. Tiemeier, [219 J.A. Hoppin, M.P. Longnecker, Levels of metabolites of organophosphate pesticides, phthalates, and bisphenol A in pooled urine specimens from pregnant women participating in the Norwegian Mother and Child Cohort [220 Study (MoBa), Int. J. Hyg. Environ. health 212 (2009) 481-491.

D.K. Li, Z. Zhou, M. Miao, Y. He, J. Wang, J. Ferber, L.J. Herrinton, E. Gao, W. Yuan, Urine bisphenol-A (BPA) level in relation to semen quality, Fertil.

Steril. 95 (2011) 625-630 e621-624. [221

A.M. Calafat, X. Ye, L.-Y. Wong, A.M. Bishop, L.L. Needham, Urinary concentrations of four parabens in the US population: NHANES 2005-2006 (Online), Environ. health Perspect. 118 (2010) 679. [222

J.D. Meeker, T. Yang, X. Ye, A.M. Calafat, R. Hauser, Urinary concentrations of parabens and serum hormone levels, semen quality parameters, and sperm DNA damage, Environ. health Perspect. 119 (2011) 252-257. [223

W.L. Ma, L. Wang, Y. Guo, L.Y. Liu, H. Qi, N.Z. Zhu, C.J. Gao, Y.F. Li, K. Kannan, Urinary concentrations ofparabens in Chinese young adults: implications for human exposure, Archives Environ. Contam. Toxicol. 65 (2013) 611-618. [224

K. Kannan, S. Corsolini, J. Falandysz, G. Fillmann, K.S. Kumar, B.G. Loganathan, M.A. Mohd, J. Olivero, N. Van Wouwe, J.H. Yang, K.M. Aldoust, Per-fluorooctanesulfonate and related fluorochemicals in human blood from several countries, Environ. Sci. Technol. 38 (2004) 4489-4495. A.M. Calafat, Z. Kuklenyik, J.A. Reidy, S.P. Caudill, J.S. Tully, L.L. Needham, [225 Serum concentrations of 11 polyfluoroalkyl compounds in the u.s. population: data from the national health and nutrition examination survey (NHANES), Environ. Sci. Technol. 41 (2007) 2237-2242. [226

L. Lind, B. Zethelius, S. Salihovic, B. van Bavel, P.M. Lind, Circulating levels of perfluoroalkyl substances and prevalent diabetes in the elderly, Diabetologia

57 (2014) 473-479.

C. Kubwabo, N. Vais, F.M. Benoit, A pilot study on the determination of perfluorooctanesulfonate and other perfluorinated compounds in blood of Canadians, J. Environ. Monit. JEM 6 (2004) 540-545.

M.K. So, N. Yamashita, S. Taniyasu, Q. Jiang, J.P. Giesy, K. Chen, P.K. Lam, Health risks in infants associated with exposure to perfluorinated compounds in human breast milk from Zhoushan, China, Environ. Sci. Technol. 40 (2006) 2924-2929.

A. Karrman, 1. Ericson, B. van Bavel, P.O. Darnerud, M. Aune, A. Glynn, S. Lignell, G. Lindstrom, Exposure of perfluorinated chemicals through lactation: levels of matched human milk and serum and a temporal trend, 1996-2004, in Sweden, Environ. health Perspect. 115 (2007) 226-230. W. Volkel, O. Genzel-Boroviczeny, H. Demmelmair, C. Gebauer, B. Koletzko,

D. Twardella, U. Raab, H. Fromme, Perfluorooctane sulphonate (PFOS) and perfluorooctanoic acid (PFOA) in human breast milk: results of a pilot study, Int. J. Hyg. Environ. health 211 (2008) 440-446.

L. Tao, K. Kannan, C.M. Wong, K.F. Arcaro, J.L. Butenhoff, Perfluorinated compounds in human milk from Massachusetts, U.S.A, Environ. Sci. Technol. 42 (2008) 3096-3101.

C. Kubwabo, 1. Kosarac, K. Lalonde, Determination of selected perfluorinated compounds and polyfluoroalkyl phosphate surfactants in human milk, Chemosphere 91 (2013) 771-777.

J. Jurewicz, M. Radwan, W. Sobala, D. Ligocka, P. Radwan, M. Bochenek, W. Hawula, L. Jakubowski, W. Hanke, Human urinary phthalate metabolites level and main semen parameters, sperm chromatin structure, sperm aneuploidy and reproductive hormones, Reprod. Toxicol. 42 (2013) 232-241.

R.W. Stahlhut, E. van Wijngaarden, T.D. Dye, S. Cook, S.H. Swan, Concentrations of urinary phthalate metabolites are associated with increased waist circumference and insulin resistance in adult U.S. males, Environ. health Perspect. 115 (2007) 876-882.

K. Kato, S. Shoda, M. Takahashi, N. Doi, Y. Yoshimura, H. Nakazawa, Determination of three phthalate metabolites in human urine using on-line solidphase extraction-liquid chromatography-tandem mass spectrometry, Journal of chromatography. B, Anal. Technol. Biomed. life Sci. 788 (2003) 407-411.

H.J. Koo, B.M. Lee, Human monitoring of phthalates and risk assessment, J. Toxicol. Environ. health. Part A 68 (2005) 1379-1392. M. Kasper-Sonnenberg, H.M. Koch, J. Wittsiepe, M. Wilhelm, Levels of phthalate metabolites in urine among mother-child-pairs - results from the Duisburg birth cohort study, Germany, 1nt. J. Hyg. Environ. health 215 (2012) 373-382.

H. Frederiksen, J.K. Nielsen, T.A. Morck, P.W. Hansen, J.F. Jensen, O. Nielsen, A.M. Andersson, L.E. Knudsen, Urinary excretion of phthalate metabolites, phenols and parabens in rural and urban Danish mother-child pairs, 1nt. J. Hyg. Environ. health 216 (2013) 772-783.

X. Han, Z. Cui, N. Zhou, M. Ma, L. Li, Y. Li, H. Lin, L. Ao, W. Shu, J. Liu, J. Cao, Urinary phthalate metabolites and male reproductive function parameters in Chongqing general population, China, 1nt. J. Hyg. Environ. health 217 (2014) 271-278.

Y. Guo, H. Alomirah, H.S. Cho, T.B. Minh, M.A. Mohd, H. Nakata, K. Kannan, Occurrence of phthalate metabolites in human urine from several Asian countries, Environ. Sci. Technol. 45 (2011) 3138-3144. J.A. Colacino, A.S. Soliman, A.M. Calafat, M.S. Nahar, A. Van Zomeren-Dohm,

A. Hablas, 1.A. Seifeldin, L.S. Rozek, D.C. Dolinoy, Exposure to phthalates among premenstrual girls from rural and urban Gharbiah, Egypt: a pilot exposure assessment study, Environ. health a Glob. access Sci. source 10 (2011)40.

J.J. Ryan, D.F. Rawn, The brominated flame retardants, PBDEs and HBCD, in Canadian human milk samples collected from 1992 to 2005; concentrations and trends, Environ. Int. 70 (2014) 1-8.

H.A. Anderson, P. 1mm, L. Knobeloch, M. Turyk, J. Mathew, C. Buelow, V. Persky, Polybrominated diphenyl ethers (PBDE) in serum: findings from a US cohort of consumers of sport-caught fish, Chemosphere 73 (2008) 187-194.

B. Gomara, L.R. Bordajandi, M.J. Gonzalez, Feasibility of two multidimensional techniques, heart-cut MDGC and GCxGC, for the separation of PCBs and PBDEs, J. Sep. Sci. 30 (2007) 1920-1929.

O.I. Kalantzi, T. Geens, A. Covaci, P.A. Siskos, Distribution of polybrominated diphenyl ethers (PBDEs) and other persistent organic pollutants in human serum from Greece, Environ. 1nt. 37 (2011) 349-353. P.1. Johnson, H.M. Stapleton, A. Sjodin, J.D. Meeker, Relationships between polybrominated diphenyl ether concentrations in house dust and serum, Environ. Sci. Technol. 44 (2010) 5627-5632.

K. 1noue, K. Harada, K. Takenaka, S. Uehara, M. Kono, T. Shimizu, T. Takasuga, K. Senthilkumar, F. Yamashita, A. Koizumi, Levels and concentration ratios of polychlorinated biphenyls and polybrominated diphenyl ethers in serum and breast milk in Japanese mothers, Environ. health Perspect. 114 (2006) 1179-1185.

A. Sjodin, L. Hagmar, E. Klasson-Wehler, K. Kronholm-Diab, E. Jakobsson, A. Bergman, Flame retardant exposure: polybrominated diphenyl ethers in blood from Swedish workers, Environ. health Perspect. 107 (1999) 643-648. G.O. Thomas, M. Wilkinson, S. Hodson, K.C. Jones, Organohalogen chemicals in human blood from the United Kingdom, Environ. Pollut. 141 (2006) 30-41.

S. Siddique et al. / Emerging Contaminants xxx (2016) 1—16

[227] A. Sjodin, R.S. Jones, S.P. Caudill, L.Y. Wong, W.E. Turner, A.M. Calafat, Poly-brominated diphenyl ethers, polychlorinated biphenyls, and persistent [238] pesticides in serum from the national health and nutrition examination

survey: 2003-2008, Environ. Sci. Technol. 48 (2014) 753-760.

[228] R. Castorina, A. Bradman, A. Sjodin, L. Fenster, R.S. Jones, K.G. Harley, E.A. Eisen, B. Eskenazi, Determinants of serum polybrominated diphenyl [239] ether (PBDE) levels among pregnant women in the CHAMACOS cohort, Environ. Sci. Technol. 45 (2011) 6553-6560.

[229] B. Fangstrom, A. Strid, P. Grandjean, P. Weihe, A. Bergman, A retrospective [240] study of PBDEs and PCBs in human milk from the Faroe Islands, Environ.

health a Glob. access Sci. source 4 (2005) 12.

[230] L. Meijer, J. Weiss, M. Van Velzen, A. Brouwer, A. Bergman, P.J. Sauer, Serum [241] concentrations of neutral and phenolic organohalogens in pregnant women

and some of their infants in The Netherlands, Environ. Sci. Technol. 42 (2008) 3428-3433. [242]

[231] W.G. Foster, S. Gregorovich, K.M. Morrison, S.A. Atkinson, C. Kubwabo, B. Stewart, K. Teo, Human maternal and umbilical cord blood concentrations

of polybrominated diphenyl ethers, Chemosphere 84 (2011) 1301-1309. [243]

[232] B. Eskenazi, L. Fenster, R. Castorina, A.R. Marks, A. Sjodin, L.G. Rosas, N. Holland, A.G. Guerra, L. Lopez-Carillo, A. Bradman, A comparison of PBDE serum concentrations in Mexican and Mexican-American children living in California, Environ. health Perspect. 119 (2011) 1442-1448. [244]

[233] J. Jin, Y. Wang, C. Yang, J. Hu, W. Liu, J. Cui, X. Tang, Polybrominated diphenyl ethers in the serum and breast milk of the resident population from production area, China, Environ. Int. 35 (2009) 1048-1052.

[234] E.F. Fitzgerald, B.A. Fletcher, E. Belanger, L. Tao, K. Kannan, S.A. Hwang, Fish consumption and concentrations of polybrominated diphenyl ethers (PBDEs) [245] in the serum of older residents of upper Hudson River communities, Archives Environ. Occup. health 65 (2010) 183-190.

[235] A. Eguchi, K. Nomiyama, G. Devanathan, A. Subramanian, K.A. Bulbule,

P. Parthasarathy, S. Takahashi, S. Tanabe, Different profiles of anthropogenic [246] and naturally produced organohalogen compounds in serum from residents living near a coastal area and e-waste recycling workers in India, Environ.

Int. 47 (2012) 8—16.

[236] L. Duedahl-Olesen, T. Cederberg, K.H. Pedersen, A. Hojgard, Synthetic musk fragrances in trout from Danish fish farms and human milk, Chemosphere 61 (2005) 422—431.

[237] J.L. Reiner, C.M. Wong, K.F. Arcaro, K. Kannan, Synthetic musk fragrances in human milk from the United States, Environ. Sci. Technol. 41 (2007)

3815-3820.

U. Raab, U. Preiss, M. Albrecht, N. Shahin, H. Parlar, H. Fromme, Concentrations of polybrominated diphenyl ethers, organochlorine compounds and nitro musks in mother's milk from Germany (Bavaria), Chemosphere 72 (2008) 87-94.

X. Zhang, G. Liang, X. Zeng, J. Zhou, G. Sheng, J. Ful, Levels of synthetic musk fragrances in human milk from three cities in the Yangtze River Delta in Eastern China, J. Environ. Sci. 23 (2011) 983-990.

E.S. Koeppe, K.K. Ferguson, J.A. Colacino, J.D. Meeker, Relationship between urinary triclosan and paraben concentrations and serum thyroid measures in NHANES 2007-2008, Sci. total Environ. 445-446 (2013) 299-305. A.M. Calafat, X. Ye, L.Y. Wong, J.A. Reidy, L.L. Needham, Urinary concentrations oftriclosan in the U.S. population: 2003-2004, Environ. health Perspect. 116 (2008) 303-307.

X. Ye, L.Y. Wong, L.T. Jia, L.L. Needham, A.M. Calafat, Stability of the conjugated species of environmental phenols and parabens in human serum, Environ. Int. 35 (2009) 1160-1163.

M. Allmyr, F. Harden, L.M. Toms, J.F. Mueller, M.S. McLachlan, M. Adolfsson-Erici, G. Sandborgh-Englund, The influence of age and gender on triclosan concentrations in Australian human blood serum, Sci. total Environ. 393 (2008) 162-167.

J.L. Wu, K.F. Leung, S.F. Tong, C.W. Lam, Organochlorine isotopic pattern-enhanced detection and quantification of triclosan and its metabolites in human serum by ultra-high-performance liquid chromatography/quadru-pole time-of-flight/mass spectrometry, Rapid Commun. mass Spectrom. RCM 26 (2012) 123-132.

T. Geens, H. Neels, A. Covaci, Sensitive and selective method for the determination of bisphenol-A and triclosan in serum and urine as pentafluorobenzoate-derivatives using GC-ECNI/MS, J. Chromatogr. B, Anal. Technol. Biomed. life Sci. 877 (2009) 4042-4046.

M. Allmyr, M.S. McLachlan, G. Sandborgh-Englund, M. Adolfsson-Erici, Determination of triclosan as its pentafluorobenzoyl ester in human plasma and milk using electron capture negative ionization mass spectrometry, Anal. Chem. 78 (2006) 6542-6546.

X. Ye, A.M. Bishop, L.L. Needham, A.M. Calafat, Automated on-line column-switching HPLC-MS/MS method with peak focusing for measuring parabens, triclosan, and other environmental phenols in human milk, Anal. Chim. acta 622 (2008) 150-156.