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Health and Safety Issues with Plasticizers and Plasticized Materials 17.1 ADJUVANT EFFECT OF PLASTICIZERS Søren Thor Larsen National Research Centre for the Working Environment, Copenhagen, Denmark

17.1.1 INTRODUCTION Several studies have proposed that exposure to chemical with so-called adjuvant effect may contribute to the increase in prevalence of allergic diseases, such as rhinitis and asthma. Chemicals with adjuvants effect usually do not induce allergy per se, but they may increase the potency of allergens. Among the numerous chemicals with adjuvant effect are members of the group of phthalate plasticizers, including di-(2-ethylhexyl) phthalate, DEHP. The effect of phthalates on the immune system and their ability to promote allergy has been studied in both epidemiological studies, in laboratory animal studies and using in vitro (cellular) models. The present chapter discusses the possible role for phthalates in the development of allergic airway diseases. 17.1.2 AIRWAY ALLERGY The two main functions of the immune system are to distinguish “self” from “non-self” material and to distinguish harmful from harmless. A failure to distinguish self from nonself is seen in autoimmune diseases such as diabetes and arthritis, where the immune system attacks tissue belonging to its own host. Failure to distinguish harmful from harmless may lead to an overreaction to innocuous substances such as allergens which may lead to development of allergy1. Allergens belong to a diverse group of substances such as proteins from pollen, furred pets, house dust mites as well as other chemicals which may act as skin sensitizers. Allergic reactions can be categorized into four types (Type I-IV reactions) 1, but most cases of airway allergy belong to the Type I allergy. Briefly, Type I allergic reactions are mediated via production of allergen-specific IgE antibodies, which upon stimulation with the appropriate allergen, cause mast cell degranulation, that is release of inflammatory mediators including histamine, prostaglandins and leukotrienes.1 These

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mediator substances are key factors in the allergic inflammation reaction and are to a large extent responsible for the observed clinical symptoms. For allergic asthma, symptoms include difficulties in breathing due to bronchoconstriction and for rhinitis. The typical symptoms include sneezing, runny nose and increased tear flow.1 Apart from being hypersensitive to specific allergens, subjects with airway allergy are often also hypersensitive to non-specific exposures, including airway irritants, organic solvent vapor, tobacco smoke and dry or cold air.2 Since some of these exposures may occur at work places, subjects with allergic airway diseases may be considered as susceptible population in the working environment. The prevalence of allergic airway diseases has been increasing in Western Europe and the US since the Second World War.3-4 Although inheritable predisposition to allergy is the main risk factor, increasing evidence point toward environmental factors and conditions may play important roles. These changes in environmental factor include changes in diet, changes in the microbial environment or exposure to substances with impact on the immune system, including the so-called environmental adjuvants. Environmental chemicals with well-known adjuvant effect include tobacco smoke, ozone and diesel exhaust particles.5-7

17.1.3 ADJUVANT EFFECT An immunological adjuvant (from Latin adjuvare, which means to assist or help) may be defined as any substance that, when given in combination with an immunogen (e.g., a vaccine antigen) acts generally to direct, accelerate, prolong or enhance the quantity of the immune response.8 Adjuvants do not necessarily by themselves elicit any antigenic response; their effect is rather to increase the potency of e.g., immunogenic substances such as vaccine antigens. The group of immunological adjuvants is very diverse and includes aluminium salts (“alum”), oil-based formulations (including Freund's incomplete adjuvant and Freund's complete adjuvant) and cholera toxin.9-10 The mechanisms through which adjuvants exert their influence on immune responses are complex, but it is believed that enhanced antigen presentation or generation of “danger” signals that increase the alertness of the immune system may be involved.11 Some adjuvants seem to specifically promote the so-called T-helper cell type 2 (Th2) response12-13 which is closely linked to IgE-mediated allergies such as rhinitis and allergic asthma. 17.1.4 ADJUVANT EFFECT OF PHTHALATE PLASTICIZERS? Several studies have addressed the question whether phthalate plasticizers may possess adjuvant effect and, if so, whether this adjuvant effect increases the risk for IgE-mediated allergies. Studies include both epidemiological investigations, in vivo (laboratory animal) and in vitro (cell culture) studies. Some of the most important epidemiological and in vivo studies are presented and discussed in the following sections. In vitro studies are not included due to their limited interpretability in relation to human risk assessment. 17.1.4.1 Epidemiological studies The first study describing a possible association between exposure to phthalate and respiratory symptoms was published in 1997 by Øie and co-workers.14 Two years later, a study demonstrating an association between exposure to phthalates and the exacerbation of respiratory symptoms, such as bronchial obstruction or wheeze in children was published15. A later study on this topic found a correlation between presence of plastic

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wall coverings at work and an increased risk of asthma.16 Another study investigated the association between DEHP in indoor dust and wheezing among preschool children.17 In general, the epidemiological studies reported an increased risk of developing various respiratory symptoms in the presence of the plastic materials in the indoor environment. Although valuable, these studies also have some limitations: Firstly, the studies demonstrate an association rather than a causal link between exposure to phthalates and development of respiratory symptoms. Secondly, most of the studies provide imprecise information on exposure levels since the exposure assessment was done after onset of symptoms, i.e. the true exposure conditions during the period where symptoms developed are not known. Thirdly, in most of the published studies, levels of allergen in the indoor environment were not measured. Since phthalate is adsorbed by the dust grain over time, it could be speculated that a high phthalate concentration in the dust is a marker of low cleaning frequency, which is likely also associated with a higher allergen concentration, which is a known risk factor for the development of respiratory allergy.18

17.1.4.2 In vivo (animal) studies Several in vivo studies have been performed to assess an adjuvant effect of the phthalates. Adjuvant effect was in most cases based on increase in the production of antibodies against a co-administered antigen. Most studies are based on mouse models where a protein antigen in combination with a phthalate plasticizer is administered through subcutaneous injection, intraperitoneal injection, dermal application19-26 or, using the human relevant exposure routes, namely oral administration or inhalation of phthalate and antigen particles.27-28 Adjuvant effect can be assessed by several immune parameters, but with respect to allergic airway diseases, the measurement of allergen-specific IgE or total IgE antibodies as well as eosinophilic airway inflammation are the most relevant and interpretable parameters. In the mouse, the production of IgE antibodies is often closely linked to the formation of IgG1 antibodies, wherefore also this antibody may be useful for risk assessment. 29 However, it is important for the interpretation of data derived from mouse models to mention that IgG1 is much less effective at stimulating mast cell degranulation (i.e., release of inflammatory mediators) than is IgE.30 Furthermore, the productions of IgG1 and IgE antibodies do not always follow the same trend,31 wherefore IgE production is the most clinical relevant and most interpretable parameter in relation of risk assessment. A structure-activity study24 of a series of phthalate plasticizers and related substances (Figure 17.1.1) has been performed in order to rule out structural and physicochemical parameters of importance for the adjuvant effect. The substances were injected intraperitoneally in combination with the model allergen OVA in BALB/c mice and the potency of the phthalates were assessed based on their ability to increase the level of OVA-specific IgG1 (Table 17.1.1). It was concluded that the most potent phthalate plasticizers had two vicinal (neighbor) alkyl chains with a sum of 16 carbon atoms, which is the case for e.g., DEHP. Another study21 investigated adjuvant effect and development of allergic lung inflammation upon inhalation of DEHP and OVA aerosols. Since DEHP and allergen are both associated with dust particles, this study mimics human exposure and is consequently useful for risk assessment. The study was a 14-week repeated dose inhalation study using different concentrations (0.022, 0.094, 1.7 or 13 mg/m3) of DEHP in combination with

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Figure 17.1.1. Structures and names of substances studied in ref. 24.

0.14 mg/m3 OVA. The study showed that DEHP upon inhalation had adjuvant effect similar to that seen after intraperitoneal or subcutaneous injection, that increased anti-OVA IgG1 antibody levels, whereas no effect was seen on the OVA-specific IgE antibody production. Eosinophilic and lymphocytic lung inflammation, indicators of allergic airway inflammation, was seen at the highest DEHP exposure level. The “margin-of-safety”, which is the distance from the highest DEHP exposure level not giving rise to an effect in mice to the actual human exposure levels, was calculated to be in the range 50-100.28 Consequently it was concluded that “realistic” DEHP exposure levels likely to be encountered in the indoor environment would not be expected to cause adjuvant effects in humans, or to result in allergic inflammation of the lung Since the majority of phthalate intake is through the diet32 it is relevant also to study the effect of orally administered phthalate on the immunological effects in mice. This was recently done by Guo et al.27 who administered DEHP (30, 300 or 300 μg/kg) daily for 52 days and immunized mice by the i.p. route (day 25, 39 and 47) followed by OVA aerosol

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17.1 Adjuvant effect of plasticizers

challenge. The authors found an association between combined DEHP/OVA exposures and serum total IgE whereas no effect was seen on OVA-specific IgE. The highest dose of DEHP furthermore increased the number of eosinophils in the lungs. Also, the highest dose of DEHP increased the bronchial hyperactivity, an indicator of asthma-like conditions, in the animals. Table 17.1.1. Adjuvant effect of test compounds based on the IgG1 levels. Compound

Adjuvant factor* (dose of test compound)

1: Di-n-butyl phthalate

24 (10 μg)

2: Benzyl butyl phthalate

1

3: Di-(2-ethylhexyl) phthalate (DEHP)

13 (10 μg) 61 (100 μg)

4: Butyl dodecyl phthalate (BDP)

20 (10 μg) 68 (100 μg)

5: Di-n-octyl phthalate

61 (100 μg)

6: Bis-(2-ethylhexyl) terephthalate (DOTP)

4 (100 μg)

7: Diisononyl phthalate

42 (both 10 and 100 μg)

8: Diisodecyl phthalate

1

9: Trioctyl trimellitate (TOTM)

1

10: Methyl palmitate (MP)

1

* Adjuvant factor = ratio between the IgG1 level in the test group and the OVA only control group.

17.1.5 CONCLUSIONS Results from epidemiological studies suggest an association between exposure to phthalates and development of respiratory symptoms related to asthma. However, it remains unclear whether the phthalate exposure has actively contributed to the development of respiratory diseases or whether the phthalate exposure may exacerbate an already existing respiratory allergy. Also, it could be speculated that the phthalate exposure could act as a surrogate marker of other exposure. Finally, although thorough phthalate exposure assessments have been made in some of the epidemiological studies, measurements are often made after onset of symptoms and it is therefore not possible to obtain exposure data for the period wherein a possible sensitization has occurred. Regarding the animal studies, there are several studies demonstrating an adjuvant effect of some of the phthalates when these are injected, ingested or inhaled. The main adjuvant effect in mice was an increase in the IgG1 response, which is less interpretable than IgE in relation to human allergic sensitization. Furthermore, the only long-term inhalation study performed showed that adjuvant effect of DEHP occurred only at very high exposure concentrations, giving a margin-of-exposure of 50-100, suggesting that realistic DEHP exposure levels, i.e., those that can be found in private homes and offices, are not likely to cause allergic sensitization and promote allergic lung inflammation. A single animal study demonstrated adjuvant effect of orally administered phthalate at dose levels not far from human exposure levels.27 The role of orally administered phtha-

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late on the risk of sensitization needs further investigation, including elucidation of mechanisms to assess the human relevance of these observations. Bearing in mind that most of the inhaled phthalate is bound to dust particles, one effective method to reduce the amount of inhaled phthalate is to keep the indoor environment clean. Since removal of dust furthermore reduces the allergen level, cleaning seems to be an effective prevention for both phthalate and allergen exposures and it further reduces any possible “cocktail effect” between the two exposures as also proposed by Nielsen et al.33

REFERENCES 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33

Mygind N, Dahl R, Pedersen S, Thestrup-Pedersen K, Essential Allergy, Blackwell Science Ltd., Oxford, 1996. Peden, D, Reed, C E, J. Allergy Clin. Immunol., 125, S150-60, 2010. Anderson H R, Ruggles R, Strachan D P, Austin B R, Burr M, Jeffs D, Standring P, Steriu A, Goulding R, Br. Med. J., 328, 1052–1053, 2004. Latvala J, von Hertzen L, Lindholm H, Haahtela T, Br. Med. J., 330, 1186–1187, 2005. Peterson B, Saxon A, Ann. Allergy Asthma Immunol., 77, 263-270, 1996. Platts-Mills T A, Erwin E, Heymann P, Woodfolk J, Allergy, 60, 25-31, 2005. von Mutius E, Immunobiology, 212 433-439, 2007. Vogel F R, Dev. Biol. Stand., 92, 241-248 (1998) Schijns V E, Curr. Opin. Immunol., 12, 456-463, 2000. Lindblad E B, Vaccine, 22, 3658-68, 2004. Gallucci S, Lolkema M, Matzinger P, Nat. Med., 5, 1249-1255, 1999. De Gregorio E, Tritto E, Rappuoli R, Eur. J. Immunol., 38, 2068-2071, 2008. Freytag L C, Clements J D, Vaccine, 23, 1804-1813, 2005. Øie L, Hersoug L-G, Madsen J Ø, Environ Health Perspect, 105, 972-978, 1997. Jaakkola J J K, Øie L, Nafstad P, Botten G, Samuelsen S O, Magnus P. Am J Publ Health, 89, 188-192, 1999. Jaakkola J J K, Ieromnimon A, Jaakkola M S, Am. J. Epidemiol., 164, 742-749, 2006. Kolarik B, Naydenov K, Larrson M, Bornehag C-G, Sundell J, Environ. Health Perspect., 116, 98-103, 2008. Nielsen G D, Hansen J S, Lund R M, Bergqvist M, Larsen S T, Clausen S K, Thygesen P, Poulsen O M., Pharmacol. Toxicol., 90, 231-242, 2002. Larsen S T, Lund R M, Nielsen G D, Thygesen P, Poulsen O M, Toxicol. Lett., 125, 11-18, 2001. Larsen S T, Lund R M, Nielsen G D, Thygesen P, Poulsen O M, Pharmacol. Toxicol., 91, 264-272, 2002. Larsen S T, Lund R M, Thygesen P, Poulsen O M, Nielsen G D, Food Chem. Toxicol., 41, 439-446, 2003. Lee M H, Park J, Chung S W, Kang B Y, Kim S H, Kim T S, Int. Arch. Allergy Immunol., 134, 213-222, (2004) Larsen S T, Nielsen G D, Toxicol. Lett., 170, 223-228, 2007. Larsen S T, Nielsen G D, BMC Immunol., 9, 61-69, 2008. Dearman R J, Betts C J, Beresford L, Bailey L, Caddick H T, Kimber I, J. Appl. Toxicol., 29, 118-125, 2009. Dearman R J, Beresford L, Bailey L, Caddick H T, Betts C J, Kimber I, Toxicology, 244, 231-241, 2008. Guo J, Han B, Qin L, Li B, You H, Yang J, Liu D, Wei C, Nanberg E, Bornehag C-G, Yang X, PLoS one, 7, 2012. Larsen S T, Hansen J S, Hansen E W, Clausen P A, Nielsen G D, Toxicology, 235, 119-129, 2007. Snapper C M, Finkelman F D, Paul W E, Immunol. Rev., 102, 51-75, 1988. Ovary Z, Int. Arch. Allergy Appl. Immunol., 69, 385-392, 1982. Sarlo K, Dearman R J, Kimber I. Guinea pig, mouse and rat models for safety assessment of protein allergenicity. In: Tryphonas H, Fournier M, Blakley B R, Smits J E G, Brousseau P (Eds.), Investigative Immunotoxicology, CRC Press LLC, Taylor and Francis, Boca Raton, 2005. Wormuth M, Scheringer M, Vollenweider M, Hungerbuhler K, Risk Anal., 26, 803-824, 2006. Nielsen G D, Larsen S T, Olsen O, Lovik M, Poulsen S K, Glue C, Wolkoff P. Indoor Air, 17, 236-255, 2007.

17.2 The rodent hepatocarcinogenic response to phthalate plasticizers: basic biology and human

17.2 THE RODENT HEPATOCARCINOGENIC RESPONSE TO PHTHALATE PLASTICIZERS: BASIC BIOLOGY AND HUMAN EXTRAPOLATION Claire Sadler,1 Ann-Marie Bergholm,2 Nicola Powles-Glover,1 and Ruth A Roberts,1~ ~ Corresponding Author AstraZeneca Global Safety Assessment, 1Alderley Park, Macclesfield, SK10 4TJ, UK and 2Pepparedsleden 1, 431 83, Mölndal, Sweden.

17.2.1 INTRODUCTION Certain phthalate plasticizers such as di-(2-ethylhexyl) phthalate, DEHP, belong to the peroxisome proliferator, PP, family of rodent liver carcinogens.1-4 Here, the evidence for peroxisome proliferator-mediated rodent carcinogenesis in response to PPs will be considered together with an evaluation of the molecular basis for rodent-human species differences in response. Specifically, this chapter will focus on the role and mechanisms of peroxisome proliferator-induced rodent peroxisomal gene expression and the evidence for lack of relevance of the mechanism to humans.

17.2.2 GENE EXPRESSION AND CANCER TOXICOLOGY 17.2.2.1 GENE EXPRESSION Within an organism such as a human or a rodent, there are many different types of cells with diverse appearances and functions. However, since they are all derived from a single fertilized egg, it is generally accepted that they all share the same genetic information. Thus, diversity of function and appearance between, for example, a muscle and a skin cell is derived from the expression of different parts of the genetic information in different tissues. In addition to diverse gene expression between cell and tissues types, certain genes are only expressed at certain times and in response to particular stimuli. For example, the hormone estrogen peaks at certain times in the female reproductive cycle, temporarily switching on certain genes in certain tissues. Each gene consists of two principle parts; the coding sequence and the promoter that acts as an on/off switch for that particular gene (Figure 17.2.1). In turn, certain genes encode regulatory proteins that control expression of the structural genes. These regulatory proteins control gene expression by operating the switch found in the gene promoter region. 17.2.2.2 CANCER BIOLOGY: SOME BASIC CONSIDERATIONS Functioning of the normal human body requires exquisite control of cell survival and proliferation; unwanted cells die whereas others proliferate just enough to maintain health or to repair injury. Cancer occurs when this regulation breaks down causing inappropriate cell proliferation, sometimes in just one cell of the billions of cells in the body. Thus, one shouldn't ask “Why does cancer occur?” but rather “Why doesn't cancer occur more frequently?” The answer to this lies in the multiple checks and balances that operate in the human body to maintain healthy function against the wealth of internal and external challenges from natural and man-made sources.

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Figure 17.2.1. DNA, genes and proteins. Each gene consists of a coding sequence and a promoter sequence. The coding sequence contains the information or “blueprint” for new proteins and the promoter contains a regulatory sequence or “switch”. This switch can be turned on or off by regulatory proteins, controlling gene expression.

17.2.2.3 CHEMICAL CARCINOGENESIS Chemicals can cause cancer in one of two main ways: they can damage DNA or they can interfere with the normal regulation of cell proliferation and cell disposal. Chemicals that damage DNA are called genotoxic (toxic to the genome) and they cause cancer by altering or mutating the genetic code. Chemicals that do not mutate DNA yet cause cancer are called nongenotoxic carcinogens. These chemicals interfere with normal cell regulation, resulting in a proliferation of unwanted cells or in the persistence of “anarchic” cells that should have been eliminated. Genotoxic chemical carcinogens can be detected easily using a range of laboratory tests that detect the genetic mutations correlated with cancer. However, for nongenotoxic chemicals, there are no such assays and detection depends principally upon tests in laboratory animals such as rats and mice given the chemicals throughout their lifetime. Occasionally, cancer does occur in mice alone or sometimes in rats and mice, particularly in the liver. On the strength of the occurrence or not of cancer in one or two rodent species, some chemicals are classed as likely or unlikely human carcinogens. This seems a reasonable “default” approach if there is no evidence to the contrary. However, experimental and epidemiological evidence shows marked species differences in response to some chemicals between rodents and humans with humans failing to show the adverse response noted in rats and mice.

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689

The more we understand about how nongentoxic carcinogens cause cancer in rodents, the more sophisticated this experimental system can be and the more sophisticated the extrapolation to humans. Recent progress means that today we are able to explain many of these changes at the level of the sequence of DNA itself via the modulation of gene expression.

17.2.3 PEROXISOME PROLIFERATORS AND RODENT NONGENOTOXIC HEPATOCARCINOGENESIS 17.2.3.1 THE PEROXISOME PROLIFERATORS Peroxisome proliferators, PPs, constitute a large and chemically diverse family of nongentoxic rodent hepatocarcinogens.5-9 This family includes fibrate hypolipidaemic drugs such as bezafibrate and gemfibrozil,10-12 given to patients at risk of heart disease to lower blood cholesterol and restore lipid balance. Also, the PP class includes chemicals of environmental and industrial significance such as the plasticizer DEHP.1,4,13,14 In the rodent, the evidence for liver tumors in response to PPs is clear and unequivocal. In addition to this hepatocarcinogenesis, PPs induce peroxisome that are responsible for metabolism of fatty acids.6 One of the key enzymes in this pathway is acyl CoA oxidase, ACO.15-17 Levels of ACO are increased dramatically in the livers of rodents treated with PPs but there is no increase of this enzyme in humans. Because of the close association between peroxisome proliferation and ACO, this enzyme is used as a marker or indicator of the rodent response to PPs. The link between peroxisome proliferation and hepatocarcinogenesis remains to be elucidated. However, evidence suggests a commonality and there is consensus that peroxisome proliferation is necessary but not sufficient per se for the observed onset of rodent liver cancer after prolonged exposure to PPs.18-20 17.2.3.2 PPAR α The liver is a major site of biotransformation and is critical in modulating chemical and metabolically induced toxicity. Peroxisome proliferator-activated receptor (PPAR) subfamily of nuclear receptors, (PPAR)α, β (also known as δ), and γ, identified in the early 1990s, function as sensors for fatty acids and fatty acid derivatives and control important metabolic pathways involved in the maintenance of energy balance. PPARs also regulate other diverse biological processes such as development, differentiation, inflammation, and neoplasia. Specifically, PPARα and PPARβ participate in energy burning, whereas PPARγ is critical in regulating adipocyte differentiation, energy storage by adipocytes and in the immune system. PPARs exhibit differential expression patterns in the liver and there is evidence to suggest that PPARs may modulate hepatotoxicity. PPARα and PPARβ exhibit a protective function in liver toxicity and studies suggest that PPARβ/δ may enhance chemically induced liver toxicity.21-25 PPARs exhibit distinct and noninterchangeable functional roles in mammalian energy metabolism but display high levels of homologies at the protein level. The PPAR subfamily consists of three members namely PPARα (NR1C1), PPARβ/δ (NR1C2), and PPARγ (NR1C3) has two isoforms with a high degree of sequence conservation across the species. All three PPARs in the human and mouse are encoded by separate genes that are on different chromosomes.21,22

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PPARα is expressed in tissues with high fatty acid oxidation activities, which include predominantly liver, but also in kidney, small intestine, heart, and skeletal muscle, which is consistent with its predominant functional role in regulating lipid catabolism. In the liver, PPARα is the master regulator of mitochondrial, peroxisomal, and microsomal fatty acid oxidation systems where it is activated by synthetic peroxisome proliferators and in addition senses the influx of fatty acids during fasting to upregulate the fatty acid burning capacity. Also in the liver, PPARβ can be activated by plasma free fatty acids which influx during fasting conditions.22-28 Generally, a given nuclear receptor regulates the expression of a prescribed set of target genes, co-activators are likely to influence the functioning of many regulators and thus affect the transcription of many genes at different times and during different cellular processes. As depicted in Figure 17.2.1, these regulatory proteins can bind to DNA and switch on gene expression. PPARα is such a regulatory protein. It switches on genes by recognizing and binding to the gene promoter region via a specific DNA sequence known as a peroxisome proliferator response element, PPREs. These areas of DNA that can be recognized by PPARα are found in the promoter regions located upstream of PP-responsive genes such as that for the peroxisomal enzyme of β-oxidation, acyl-CoA oxidase (ACO) (Figure 17.2.2.). In the nucleus, PPARs exist as heterodimers with retinoid X receptor-α bound to DNA with corepressor molecules. Upon binding of a ligand, PPARs undergo shape changes that aid the removal of co-repressor molecules and invoke a space fitting recruitment of transcription co-factors including coactivators such as PPAR-binding protein (PBP/ PPARBP), thyroid hormone receptor-associated protein 220 (TRAP220), mediator complex subunit 1 (MED1) and co-activaFigure 17.2.2. PPARα mediates the rodent response to tor-associated proteins. These associations PPs. Binding sites for PPARα have been found in the promoters of genes associated with peroxisome prolifer- may exert a broader influence on the funcation such as acyl CoA oxidase, providing proof that tions of several nuclear receptors and their PPARα can operate the “switch” and turn on expression of rodent genes known to be responsive. The binding site target genes. Functional significance for within the gene promoter is called a peroxisome prolifer- the existence of over 200 nuclear receptor ator response element (PPREs) and is defined by the cofactors is not readily evident, but emergDNA sequence TGACCT repeated once with a one letter ing gene knockout mouse models show that “spacer” to give TGACCT n TGACCT. some of the coactivators are essential for embryonic growth and survival and for controlling receptor specific target gene expression in cell specific need based demands. PPARα activation is responsible for the pleiotropic effects of PPs seen in rodents such as enzyme induction, peroxisome proliferation, liver enlargement, and tumors.29-31 Evidence for this is strong and is derived from studies of mice that have had their DNA altered so that they no longer possess PPARα (Figure 17.2.3). These mice are referred to as PPARα null transgenic mice. The PPARα null mouse is refractory to the effects of PPs

17.2 The rodent hepatocarcinogenic response to phthalate plasticizers

691

Figure 17.2.3. The response to PPs is lost in a transgenic mouse that has had its DNA altered so that it no longer has the regulatory protein, PPARα. In the PPARα null mouse, there is no peroxisome proliferation, cell proliferation, liver enlargement nor tumors in response to PPs.

such as peroxisome proliferation, cell proliferation, liver enlargement, and tumorigenesis.29-31 Thus, data support the position that the pleiotropic effects of PPs in the rodent are mediated by PPARα. The validity of this conclusion has been tested rigorously in the PPARα null mouse using doses of DEHP sufficient to cause significant body weight loss and 100% morality in wild type mice by 16 weeks. In this study, PPARα null mice fed DEHP beyond the time at which the wild-type mice had died showed no liver effects.31 In summary, PPARα mediates the hepatocarcinogenic effects of PPs in the rodent; there are no data to support such effects independent of PPARα.

17.2.4 SPECIES DIFFERENCES IN RESPONSE TO PEROXISOME PROLIFERATORS It is well established that there are species difference in response to PPARα activation and peroxisome proliferation.6,32-36 Peroxisome proliferator chemicals are classic nongenotoxic carcinogens. These agents cause liver cancer when chronically administered to rats and mice (not hamster). Available data (both in vitro and in vivo) suggest the rat as the most sensitive and man as the non responsive species to this effect. Furthermore, studies with cultured human hepatocytes show that there is no peroxisome proliferation or induction of S-phase in response to PPs. The cascade of molecular events leading to liver cancer in rodents involves hepatocyte proliferation, oxidative stress, increase in proinflammatory cytokines, and inhibition of apoptosis. The direct target genes involved in the hepatocarcinogenic effect are not known but certainly there is induction of genes involved in lipid metabolism but not in hepatocellular proliferation.37-41 It has been shown42 that the mechanism of hepatocellular proliferation involves downregulation of the microRNA let-7c gene by PPARα. Let-7c controls levels of proliferative c-myc by destabilizing its mRNA. Thus, upon suppression of let-7c, c-myc mRNA and protein are elevated, resulting in enhanced hepatocellular proliferation. PPARα humanized mice are resistant to peroxisome proliferator-induced cell proliferation and

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Health and Safety Issues with Plasticizers and Plasticized Materials

cancer. These mice do not exhibit downregulation of let-7c gene expression thus resistance to hepatocellular carcinogenesis. Furthermore, in a study in cyno monkeys after given a high dose (1g/kg/day) orally of DEHP for 28 days32 a subtle increase in the numbers of peroxisomes (in the same magnitude as the control group given corn oil) with slight enlargements of the mitochondria was demonstrated. The activity of the mitochondrial enzyme CPT increased significantly in males (in the same magnitude as the control group given corn oil). No statistically significant differences were observed in the FAOS or CAT activities. This low sensitivity response to peroxisome proliferators in monkeys after a very high dose of DEHP was considered to be closer to the response in humans than that seen in rodents. From these data it could be concluded that DEHP induced hepatic peroxisome proliferation in cynomolgus monkeys, however, the degree of increase was very low, hepatomegaly or hepatic proliferation was not observed and the exposure level was extremely high. Despite a huge amount of investigations performed, the underlying mechanism behind PPARα agonist-induced hepatocarcinogenesis is not yet fully understood. In contrast to genotoxic carcinogens that are activated to electrophilic derivatives that can bind DNA and directly mutate genes, peroxisome proliferators are not metabolically activated. We can assume that hepatic peroxisome proliferation in human resulting from DEHP is subtle, just as in the case of cynomolgus monkeys. There have been no reports showing that peroxisome proliferators induce mitochondrial changes in humans. The issue for peroxisome proliferators is the risk of hepatocarcinogenesis, not peroxisome proliferation itself. Hoivik et al.43 suggested that the primate may be refractory to PPAR-induced hepatocarcinogenesis because cynomolgus monkeys responded to fibrates in a manner that is different from the rodent; that is to say, there was no indication of cell proliferation, and there was no remarkable increase in the mRNA levels for most proteins known to respond to oxidative stress. In addition, human liver does contain a functional PPARα44 although the expression of PPARα in humans is around 10-fold lower when compared with responsive species such as rat and mouse.45-46 In total, these data support a “quantitative” hypothesis whereby PPARα expression in humans is sufficient to mediate the beneficial effects of hypolipidaemic drugs via regulation of genes for enzymes and lipid transporters. Expression levels are too low, however, for modulation of the full battery of genes that are activated in rats and mice such as those involved in peroxisome proliferation and perturbation of hepatocyte growth control. The second hypothesis to explain lack of human response is based on quality of the PPARα-mediated response. Thus, even in the presence of sufficient human PPARα, genes associated with rodent peroxisome proliferation and cancer would not be switched on. Evidence in support of this hypothesis arises from work that shows species difference in the sequence of the ACO gene promoter,47 a marker for rodent peroxisome proliferation (Figure 17.2.4). The rat ACO gene promoter contains binding sites for PPARα known as PPREs and, as expected, rodent ACO levels are increased in the presence of PPs. In contrast, the human gene sequence differs from the rat gene sequence resulting in an inactive “switch” and no ACO increase in human hepatocytes. Thus, lack of human response to PPs may be attributed to a non-functional “switch” in the genes associated with rodent peroxisome proliferation.

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The mode of action for PPARα agonist induced liver cancer is relatively well established. There is good evidence showing a profound species difference in the response to PPARα agonism in liver, with rodent models consistently showing enhanced sensitivity as compared to nonhuman primate and human models, which typically show a diminished response.48 In total, there is good reason to suggest that humans are refractory to PPARα agonist-induce liver cancer, but there are clearly some data gaps that should be filled to specifically delineate the mechanisms underlying the species differences. In fact, up to today there have been no reports showing that peroxisome proliferators cause hepatocarcinogenesis in nonhuman primates or humans.

17.2.5 CHEMICAL REGULATION Since 2007 all chemicals are assessed under the EU chemical legislation REACH (Registration, Evaluation, Authorization and restriction of CHemicals). The aim of this legislative activity is to ensure a high level of protection for workers, consumers and the environment against dangerous chemicals.49 The category the chemical is placed in is represented on the label and can have significant economic consequences, restriction of use and progressive substitution of the most dangerous chemicals when suitable alternatives have been identified. The risk assessment, and therefore the categorization, is based on hazard identification through in vivo studies and exposure calculations. DEHP is a good example of how scientific evidence contributes to the risk assessment and can affect the categorization. Although DEHP causes liver tumors in rats after prolonged exposure, the mechanistic understanding and in vivo evidence of peroxisome proliferator-mediated rodent carcinogenesis explains why these are not considered to be relevant to humans, and therefore DEHP is not classified as a human carcinogen.

17.2.6 SUMMARY In summary, the adverse response of rodents to PPs is mediated by PPARα. The Figure 17.2.4. A. Rat, B. Human. Species differences in scientific evidence demonstrates that ACO gene promoter sequence and activity. The rat ACO humans are less sensitive to peroxisome gene is switched on when PPs activate their receptor proliferation and non-responsive to tumors PPARα since PPARα can bind to a specific DNA sequence (TCACCT T TGTCCT) found in the rat gene induced by PPs such as DEHP. These spepromoter. This results in rat ACO gene expression. In cies differences may be attributed to both contrast, the DNA that makes up the human gene prodifferences in the quantity of PPARα and moter has a different sequence that cannot be switched to DNA sequence differences in the proon and the human ACO gene is not expressed in response to PPs. moter regions of genes found to be respon-

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sive to PPs in the rodent. At least for one gene that is a marker of rodent peroxisome proliferation, these sequence differences result in a non-functional switch that cannot be activated. These data suggest that PPs such as the phthalates DEHP and DINP pose no significant risk of cancer to humans. Recent results can be concluded as follows:50 “Although some effects of CAR and PPARα activators can be observed in human liver, the major species difference between rodents and humans is that while these compounds are mitogenic agents in rodents, they do not stimulate replicative DNA synthesis in human hepatocytes. Hence the MOAs for rodent liver tumor formation by these chemicals are not plausible for humans. This conclusion is supported by epidemiologic studies with phenobarbital and hypolipidemic drugs where no increased risk of liver tumors has been reported.”50

REFERENCES 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32

Doull J, Cattley R, Elcombe C, Lake B, Swenberg J, Wilkinson C, Williams G, van Gamert M, Regulatory Pharmacology and Toxicol., 29, 327, 1999. Kurata Y, Kidachi F, Yokoyama M, Naoto T, Masanobu K, Toxicol. Sci., 42, 49, 1998. Lake B, Gangoli S, Grasso P, Lloyd A, Toxicol. Appl. Pharmacology, 32, 355, 1975. National T P Publication, No, 82, 1982. Bentley P, Calder I, Elcombe C R, Grasso P, Wiegand H G, Stringer D A, ECETOC Monograph, 17, 1992. Ashby J, Brady A, Elcombe C R, Elliot B M, Ishmael J, Odum J, Tugwood J D, Kettle S, Purchase I F H, Human Experimental Toxicol., 13, S1, 1994. ECETOC; European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC); Brussels, 1992; Vol. 17. Reddy J, Azarnoff D, Svodoba D, Prasad J, J. Cell Biol., 61, 344, 1974. Reddy J K, Azarnoff D L, Highnite C E, Nature, 283, 397, 1980. Frick H, Elo Haapa K, Heinonen O P, New Eng. J. Medicine, 317, 1235, 1987. Giometti C S, Taylor J, Gemmell M A, Tollaksen S L, Lalwani N D, Reddy J K, Appl. Theor Electropher, 2, 101, 1991. IARC, 1996 Ward J M, Hagiwara A, Anderson L M, Linsey K, Diwan B A, Toxicol, Appl. Pharmacology, 96, 494, 1988. Ward J M, Diwan B A, Ohshima M, Hu H, Schuller H M, Rice J M, Environ, Health Perspectives, 65, 279, 1986. Bell D R, Bars R G, Gibson G G, Elcombe C R, Biochem J., 275, 247, 1991. Berthou L, Saladin R, Yaqoob P, Branellec D, Calder P, Fruchart J-C, Densfle P, Auwerx J, Staels B, Eur. J. Biochem., 232, 179, 1995. Lazarow P B, De Duve C, Cell Biol., 73, 2043, 1976. Chevalier S, Macdonald N, Roberts R, J. Cell Sci., 112, 4785, 1999. Roberts R, James, N, Hasmall S, Holden P, Lambe K, Macdonald N, West D, Whitcombe D, Woodyatt N, Toxicol. Lett., 112-113, 49, 2000. Roberts R, Moffat G, Comments Toxicol., 7, 259, 2001. Michalik L, Auwerx J, Berger J P, Pharmacol. Rev., 58, 4, 726-41, 2006. Bookout A L, Jeong Y, Downes M, Yu R T, Evans R M, Mangelsdorf D J, Cell, 126, 4, 789-99, 2006. Mukherjee R, Locke K T, Miao B, J. Pharmacol. Exp. Therapeutics, 327, 3, 716-26, 2008. Mei C-L, He P, Cheng B, Liu W, Wang Y-F, Wan J-J, Cell Biol. Int., 33, 3, 301-8, 2009. Crisafulli C, Cuzzocrea S, Shock, 32, 1, 62-73, 2009. Pyper S R, Viswakarma N, Yu S, Reddy J K, Nuclear Receptor Signalling, 16, 8, article e002, 2010. Sanderson L M, Degenhardt T, Koppen A, Molec. Cellular Biol., 29, 23, 6257-67, 2009. Conaway J W, Florens L, Sato S, FEBS Lett., 579, 4, 904-8, 2005. Lee S S-T, Pineau T, Drago J, Lee E J, Owens J O, Kroetz D L, Fernandez-Salguero P M, Westphal H, Gonzalez F J, Molec. Cellular Biol., 15, 3012, 1995. Peters J M, Cattley R C, Gonzalez F J, Carcinogensis, 18, 2029, 1997 Ward J, Peters J, Perella C, Gonzalez F, Toxicol. Pathology, 26, 240, 1998 Satake S, Nakamura C, Minamide Y, Kudo S, Maeda H, Chihaya Y, Kamimura Y, Miyajima H, Sasaki J, Goryo M, Okada K, J. Toxicol. Pathol., 23, 75-83, 2010.

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17.3 THE INFLUENCE OF MATERNAL NUTRITION ON PHTHALATE TERATOGENICITY Janet Y. Uriu-Adams1 and Carl L. Keen1,2 1

2

Departments of Nutrition and Internal Medicine, University of California at Davis, One Shields Avenue, Davis, California, 95616-8669, USA

17.3.1 INTRODUCTION It has been estimated that 2-3% of the world's annual 140 million births will have a major congenital malformation. Despite improvements in the infant mortality rate, birth defects remain the leading cause of infant death in the United States followed by prematurity/low birth weight.1 The World Health Organization, WHO, defines low birth weight as a birth weight <2500 g; low birth weight can be due to preterm birth and intrauterine growth retardation, IUGR. The prevalence of IUGR varies among countries and can be six times higher in developing countries compared to developed countries.2,3 In addition, both preterm infants and IUGR infants have higher risks of early mortality and morbidity. Low birth weight has consistently been associated with chronic diseases in adulthood such as coronary heart disease, hypertension and type-2 diabetes.4-10 Determining how epigenetic events that occur in utero can lead to persistent changes in gene expression and increased risk of disease in later life is an area of increasing interest. If the concept that chronic diseases can have their origin during fetal development is substantiated, we may well need to redefine what is intended by the words “birth defects”, and the calculated frequency of “birth defects” may need to be substantially elevated. Despite extensive investigative efforts, the mechanism(s) involved in the development of adverse pregnancy outcomes are still poorly understood. It is likely that a variety of environmental, maternal and embryo/fetal factors interact and contribute to the formation of birth defects and the development of IUGR. We and others have hypothesized that the teratogenicity of diverse compounds are mediated, in part, through the modulation of maternal/embryo/fetal nutrient metabolism. This subchapter will first briefly review the teratogenicity of two phthalates used in the manufacturing of plastic consumer products; di-(2-ethylhexyl) phthalate, DEHP, and butyl benzyl phthalate, BBP. These phthalates were chosen because they are widely used, are found in the environment, and there is public concern regarding their safety. The National Toxicology Program, NTP, and the Center for the Evaluation of Risks to Human Reproduction, CERHR, Phthalates Expert Panel have recently published extensive reviews of the reproductive and developmental toxicity of seven phthalate compounds, including DEHP and BBP.11-17 In the second part of this subchapter, data will be presented delineating how the induction of an acute phase response in the mother by physiological stressors and diverse developmental toxicants, including certain phthalates and their metabolites, can alter maternal zinc metabolism. The sequestration of zinc in maternal tissues can subsequently precipitate a zinc deficiency in the embryo/fetus and result in teratogenesis. A brief discussion will follow of acute phase response-induced changes in other nutrients, along with a discussion on how these changes may also represent a reproductive challenge.

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17.3.2 REPRODUCTIVE TOXICITY OF BBP AND DEHP Phthalates are a group of phthalic acid diester compounds with straight or branch chain alcohols that are commonly used as plasticizers to add flexibility to plastic consumer products. Phthalates are used in the manufacturing of diverse products such as artificial leather upholstery for cars and homes, medical devices, shower curtains, vinyl tiles, clothing such as raincoats and footwear, cosmetics, food packaging, and children's toys.11-17 Phthalates or their metabolites can be cytotoxic at high concentrations. While the exact mode of action for the toxicity of phthalates is not known, it has been suggested that their toxicity is due to multiple factors including their ability to induce oxidative stress,18-21 their ability to modulate steroid hormone metabolism,20,22-24 and their ability to stimulate peroxisome proliferation.20,25-29 Since phthalates are not covalently bound to plastics, exposure through oral and dermal routes by the use of phthalate-containing products, or through their presence in the environment is a public concern. At high doses, phthalates have been shown to be developmental and reproductive toxicants in animal models; the severity of the effects differs depending on the type of chemical, dose, route of administration, timing of exposure, and species tested.11-17,30 Manifestations of adverse reproductive effects such as reduced testosterone and spermatocytes in males are observed at later stages of the life cycle. Thus, high exposures during pregnancy can lead to persistent and adverse effects in adulthood. Data on the effects of phthalate exposure on human reproduction and development are limited. It has been estimated that human exposures to phthalates are below the NOAELs (no observable adverse effect level) obtained from animal reproductive and developmental toxicity studies, thus there is generally negligible or minimal concern for adverse reproductive or developmental effects from phthalates in adult humans.31 However, as discussed below, the nutritional status of the mother can significantly affect her response to teratogens. In addition, while the majority of current interest in the fetal origins of adult disease hypothesis is centered on maternal nutritional insults, developmental exposure to xenobiotics needs to be equally considered, as do potential xenobiotic-nutrient interactions. This point is discussed below for 2-ethylhexanoic acid (2-EHA), a metabolite of the parent compound DEHP. DEHP is used to increase flexibility of polyvinylchloride, PVC, and is found in a wide array of products including medical devices, building and car products, food packaging, and children's products. For the general adult human population, exposures are estimated to be 3-30 μg/kg/day.15 Exposures may be higher in infants and toddlers due to mouthing behaviors of DEHP-containing toys and products, or to parenteral DEHP exposure through therapeutic medical interventions such as replacement of blood products, exchange transfusion, extracorporeal membrane oxygenation, ECMO, and other procedures that are vital for the care of critically ill infants.29 In these two cases, the Phthalate Expert Panel expressed concern and serious concern, respectively, that high exposure to DEHP may adversely affect male reproductive tract development.15 The Panel also expressed concern that “ambient oral DEHP exposures to pregnant or lactating women may adversely affect the development of their offspring”.15 Currently, the European Union does not allow the use of several phthalates (greater than 0.1%) in toys and products made of PVC that are intended to be put in the mouth by children under the age of three.32 Similarly, phthalates are not being used in the manufacturing of infant bottle nipples, teethers, and toys intended for mouthing in the United States and Canada.29

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When DEHP is given to mice during pregnancy, fetal toxicity is observed including embryo or fetal death, decreased growth, and malformations of the skeletal and cardiovascular systems, the neural tube and eye.33,34 Postnatal loss and eye defects have also been observed in offspring of rats exposed to high doses of DEHP during days 6-15 of gestation.35 The testes (specifically the Sertoli cells) are particularly vulnerable to DEHPinduced toxicity. Consistent pathological findings include decreased testes weight and atrophy, alterations in Sertoli cell ultrastructure and function, and reduced sperm numbers.15 High doses of DEHP given by oral gavage to young or adult rats for up to 21 days resulted in atrophy of the seminiferous tubules.36,37 DEHP has been shown to decrease antioxidant levels including glutathione, ascorbic acid and free thiols in a dose-dependent manner, and increase production of reactive oxygen species leading to increased apoptosis.18 Chronic exposures (13-104 weeks) of DEHP in young and mature rodents result in atrophy of the seminiferous tubules and loss of spermatogenesis.38,39 In rats, DEHP exposure throughout the prenatal and postnatal periods leads to decreased fertility in males and complete infertility in females indicating a functional reproductive deficit.40 Histological changes of testes from offspring of rat dams exposed to DEHP in drinking water during pregnancy and lactation included disorganization of tubular epithelium, spermatogonial cell detachment from the basal membrane, and disrupted spermatogenesis.41 Pre- and post-natal exposure of DEHP has also been shown to reduce testosterone production in male fetuses.42 Mature male rats gavaged with DEHP for 4 weeks showed a dose-related reduction in male fertility and severe testicular lesions.43 A recent report from the US Food and Drug Administration indicates that highly exposed male infants (e.g., exposure from life-saving medical procedures) could be at risk of DEHP-induced testicular toxicity.44 High doses of DEHP (2000 mg/kg/day) have also been shown to negatively affect estradiol synthesis and ovulation in adult female rats.26,30,45 In rodents, orally administered DEHP is rapidly converted to monoethylhexyl phthalate, MEHP, the recognized toxic metabolite, which is then absorbed across the intestine. Similar to DEHP, MEHP is a developmental toxicant and a teratogen.15 Using co-cultures of Sertoli cells and gonocytes from neonatal rats, Li and coworkers have shown that MEHP (but not DEHP) has an anti-proliferative effect in vitro46 that may contribute to the reduced testes weights in vivo. Apoptosis may also play a role in testes atrophy as MEHP has been shown to increase the number of apoptotic cells in vitro,47 and it can up-regulate the expression and activity of death receptors (DR4, 5 and 6) in the testis in vivo.48 The concentrations of DEHP monoester metabolites in urine have been found to be high in young children (samples collected from 328 children from 6-11 years of age),49 suggesting that environmental exposure to DEHP may be high for this sensitive age group, although the extent to which this represents a health risk is a subject of considerable controversy.29 Butyl benzyl phthalate, BBP, is largely used as a plasticizer for vinyl floor tiles.50 It is also used in sealants, vinyl foams, adhesives and carpet backing, and food conveyor belts.50 Human exposure to BBP has been estimated to be 2-4 μg/kg/day.17,51 While human data with regard to developmental toxicity or reproductive toxicity of BBP are lacking, numerous studies in rodents show that high levels of BBP (typically greater than 750 mg/kg body weight/day) given orally by gavage or fed in the diet during gestation can negatively affect development and result in embryo/fetal death, reduced growth, and gross, skeletal and soft tissue malformations.52-62 The teratogenic effects of BBP have also

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been observed in mice and chickens.53,63 In a two-generation reproductive toxicity study, male and female rats were gavaged with BBP (0, 20, 100, or 500 mg/kg body weight/ day).64 At the 500 mg/kg dose, adverse developmental effects were observed in the offspring including low birth weights (but no effects on viability), decreased anogenital distance and delayed preputial separation in males, and reduced serum testosterone and spermatocytes in the testes of mature males. It should be noted that when these animals reached sexual maturity, there were no effects of BBP on reproductive capacity. Other studies confirm that large doses of BBP (typically greater than 1000 mg BBP/ kg/day) adversely affect development of the male reproductive tract resulting in reduced sperm counts and testicular atrophy in adult rats.65,66 High doses of BBP administered during pregnancy reduced the anogenital distance of males and increased the incidence of undescended testes.67 In addition, BBP exposure from gestation day 14 to postnatal day 3 altered sexual differentiation in the male offspring, and affected males exhibited femalelike areolas/nipples.68 The developing male reproductive system seems to be a more sensitive target of phthalate toxicity than adult testis.69 The NTP stated that based on the strong animal data, there is a “potential for similar or other adverse effects in human populations if exposures are sufficiently high”.31 However, based on the human exposure estimates of Kohn and coworkers,51 the NTP concluded that there is “minimal concern for developmental effects in fetuses and children” and “negligible concern for adverse reproductive effects in exposed men”.31

17.3.3 ACUTE PHASE RESPONSE-INDUCED ALTERATIONS IN MATERNAL AND CONCEPTUS NUTRIENT METABOLISM It is well established that maternal nutrition can significantly impact pregnancy outcome. Inadequate calorie intake, and the consumption of “poor diets” (typically classified as low in protein, dairy products and fresh fruits and vegetables) have been associated with a high incidence of in utero and early postnatal death, and central nervous system and behavioral abnormalities.70-72 Several variables can influence an individual's nutritional status including diet, genetics, environment, lifestyle habits, the presence of disease or physiological stressors, and drug-toxicant exposures. Of particular interest to this subchapter is that drug-toxicant exposures can induce an acute phase response and alter maternal nutrient metabolism leading to developmental defects. One of the micronutrients that can be altered by the acute phase response is zinc. In rodent and non-human primate models, congenital malformations, intrauterine growth retardation, high perinatal mortality, and persistent behavioral and immune defects can be consequences of prenatal zinc deficiency.73-75 Typical malformations associated with severe zinc deficiency in animal models include cleft lip and palate, brain and eye malformations, lung, urogenital and skeletal anomalies, and numerous abnormalities of the heart.74 Even transitory, short-term zinc deficiency can result in severe malformations.76 Altered zinc status in humans is also associated with pregnancy complications, malformations, and low birth weight.77-82 In humans, suboptimal zinc status can occur through a variety of means. The average dietary intake of zinc during pregnancy is well below the RDA83-86 suggesting that suboptimal zinc status is common in many population groups. Dietary factors such as phytate and fiber can also reduce the bioavailability of zinc.87 Prasad and coworkers first documented that zinc deficiency was a major contributor to “nutritional dwarfism” in humans

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in parts of the Middle East.88 Male villagers in Iran exhibited retarded growth, anorexia, lethargy and hypogonadism due in part to consumption of diets high in phytate-containing cereals.88 Growth rate and gonadal changes were reversed with zinc supplementation. Widespread zinc deficiency was subsequently documented in several other countries, and the WHO recently identified zinc deficiency as one of the twelve leading risk factors for early death in developing countries.89 Given the prevalence of mild to moderate zinc deficiency, it is reasonable to predict that a woman's zinc status can influence her susceptibility to various teratogens. In addition to primary zinc deficiency, a secondary zinc deficiency can also be precipitated when micronutrients, such as iron, interact with the absorption and metabolism of zinc.90 Additionally, gene mutations can result in a higher than normal nutrient requirement. For example, individuals with acrodermatitis enteropathica, a genetic disorder in zinc absorption, require a high amount of zinc in their diet.91 Alterations in zinc metabolism can also occur when an acute phase response is triggered by a wide variety of factors, including numerous diverse drugs and toxicants, physical stimuli such as infection, inflammation, tissue trauma, ischemic necrosis, and myocardial infarction, psychological stress, and several disease conditions.90 The acute phase reaction is a non-specific metabolic response to re-establish homeostasis and to repair injured tissues.92 Many chemicals or infectious agents can cause hepatocellular injury by inducing infiltration of inflammatory cells to the liver resulting in hepatocyte cell death.93 Inflammatory mediators released by immune cells and tissues include cytokines such as tumor necrosis factor alpha, TNF-α, interleukin-1, IL-1, interleukin-6, IL-6, and gamma interferon, IFNγ, reactive oxygen and reactive nitrogen species, as well as metabolites of arachidonic acid (e.g. leukotrienes, prostaglandins, and thromboxane A2), all of which promote the inflammatory process.94 The body's response to this complex interaction among stress hormones and cytokines is to produce fever, increase circulating glucocorticoids, alter fuel metabolism, and up- or down-regulate the hepatic production of a number of acute-phase proteins. Several of these acute-phase proteins are involved in the metabolism of metals including zinc, copper and iron. The acute phase response protein, metallothionein, MT, belongs to a family of low molecular weight cysteine-rich proteins that can bind nutritionally important bivalent cations such as zinc and copper, as well as the environmental toxicants, cadmium and mercury.95 Two of the MT proteins (MT-1 and MT-2) are widely expressed in the body with highest concentrations in the liver, kidney, intestine and pancreas. The MT-3 isoform is predominantly found in brain while expression of the MT-4 isoform has been detected in skin and tongue squamous epithelium and maternal deciduum.96,97 Postulated functions of MT include protection against heavy metal toxicity, as a contributor to zinc homeostasis and metabolism, and as an antioxidant and free radical scavenger. The promoter region of the MT-gene contains a number of response elements that are sensitive to glucocorticoids (glucocorticoid responsive element; GRE), reactive oxygen species (antioxidant response element; ARE), and metals such as zinc and cadmium (metal regulatory elements; MREs).95 Thus, the production of the MT protein can be increased by various stimuli including the presence of metals, as well as conditions that result in oxidative stress, tissue damage, inflammation or infection. Our group has proposed the hypothesis that diverse chemicals, toxicants and stresses mediate their teratoge-

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nicity by inducing an acute phase response in the mother and ultimately altering zinc metabolism in the conceptus.74 In this scenario, as a consequence of the maternal acute phase response, there is a marked increase in the synthetic rates of several proteins including MT in maternal liver. The increase in the concentration of MT results in a sequestration of zinc in maternal liver, a drop in plasma zinc, and reduced zinc transfer to the embryo. If the decreased zinc transfer to the embryo occurs during critical periods of organogenesis, zinc-dependent processes can be disrupted in the growing embryo leading to developmental defects and malformations.74,90,95 Numerous diverse compounds have been shown to mediate developmental toxicity by altering maternal and embryo/fetal zinc metabolism. For example, the saponin, α-hederin, which is a component of oriental herbs used to treat hepatitis and infectious diseases, has been shown to induce maternal liver MT and zinc concentrations and result in decreased embryonic zinc, and increased malformations.98,99 Using an oral dose of the radioactive isotope 65Zn, Duffy and coworkers showed that the tissue distribution of 65Zn reflected the increase in hepatic MT.98 When rat embryos were cultured in serum taken from rats two hours after α-hederin treatment (e.g. before the increase in maternal liver MT production), embryo development was normal. In contrast, embryos cultured in serum taken from rats 18 hours after α-hederin treatment (e.g., peak MT production) developed multiple abnormalities. The addition of zinc to the 18-hour post-α-hederin treated serum resulted in normal embryonic development99 indicating that the low zinc content of the serum was directly responsible for the α-hederin-induced teratogenicity. Similar maternally-mediated zinc deficiency-induced teratogenicity has been noted for the cancer drug 6-mercaptopurine,100,101 the anti-convulsant drug valproic acid,102 the anesthetic urethane,103 ethanol administration,104-107 and chronic disease states such as diabetes.108,109 Other conditions that increase pro-inflammatory cytokines such as maternal stress induced by restraint (immobilization)110,111 have been shown to be teratogenic.112 It has been suggested that many chemicals may become teratogenic when they are present at doses that are sufficiently high to induce overt maternal toxicity and maternal stress.113 Additionally, maternal stress can also enhance the teratogenicity of other agents.114 Lipopolysaccharide (LPS or endotoxin) can also activate the release of cytokines such as TNF-α, IL-1, and IL-6, and promote the acute phase response. LPS administration to pregnant mice results in high embryo lethality.115 In contrast, LPS added to embryo culture media was not developmentally toxic indicating that the teratogenic effects of LPS in vivo are maternally mediated.115 One of the cytokines whose production is increased early in response to tissue injury and inflammation is tumor necrosis factor alpha, TNF-α.116-118 TNF-α is a key mediator in the inflammatory response and increased TNF-α levels can induce genes involved in inflammation including cytokines, chemokines, nitric oxide synthase and adhesion molecules. We have shown that a teratogenic dose of TNF-α elicits an increase in MT synthesis and zinc levels in maternal liver, and a reduction in zinc transport into the mouse embryo, even in instances where the mother is fed a zinc adequate diet.119 With regard to phthalates, DEHP has been shown to induce maternal liver MT and lead to a functional zinc deficiency in the embryo indicating that altered Zn metabolism may contribute to DEHP-induced reproductive toxicity and teratogenicity.120 Intubation of

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high doses of 2-EHA (a secondary metabolite of DEHP) administered acutely (single dose of 6.25, 9.38 or 12.5 mmol/kg body weight on gestation day 11.5) or over several days (single dose of 3.5 mmol/kg body weight on gestation days 8-15) has been shown to induce maternal liver MT, increase the percentage of an oral dose of 65Zn in maternal liver, decrease embryonic zinc uptake and retention, and increase fetal malformations including encephalocele and tail defects.121 When exposure to 2-EHA was coupled with a low zinc diet, the incidences of encephalocele and tail defects were increased whereas dietary zinc supplementation decreased 2-EHA-induced teratogenicity121 indicating that a mother's diet and nutritional status can modulate her response to teratogens. In contrast, the phthalate BBP administered at high doses (1000, 1500 or 2000 mg/kg on gestation days 11-13) was teratogenic and maternally toxic but did not alter maternal MT concentrations, or maternal or embryo 65Zn metabolism.62 These data indicate that there are multiple mechanisms of phthalate teratogenicity. The direct toxicity of several of the compounds used in the acute-phase response experiments has been investigated using the in vitro embryo culture technique. Urethane and α-hederin were not directly developmentally toxic as embryos cultured in their presence exhibited normal development.99,103 Embryos cultured in conditioned serum (e.g., serum from 2-EHA-treated rats or TNF-α-treated rats) developed abnormally while supplementation of zinc to the above conditioned embryo culture media reduced their teratogenicity.121,122 Similar to results from the 2-EHA studies,121 the teratogenic effects of chemical or stress-induced acute phase response in vivo were further amplified in animals fed low zinc diets while they were reduced in animals fed zinc supplemented diets.74,101,103,119 A series of papers investigating the effects of the acute phase inducers ethanol and LPS in mice that lack MT show that the teratogenicity of alcohol and LPS is reduced in MT-/- mice.106,107,123 Taken together, these data support the concept that the teratogenicity of diverse compounds is mediated through alterations in maternal and embryonic/fetal zinc metabolism, and that the induction of maternal hepatic MT is an important feature in this process. In addition to phthalate-induced alterations in zinc metabolism during pregnancy, postnatal and adult exposure to phthalate esters including di-n-butyl phthalate, DBP, di-npentyl phthalate, DEHP and various isomers of monobutyl-o-phthalate can alter zinc metabolism, and induce tissue damage, particularly in the testes.124-130 Oral administration of DEHP to neonatal and adult rats showed a decrease in Sertoli cell numbers, a doserelated decrease in maturation of the spermatids in the tubules, and decreased testicular zinc concentrations in the older animals.131 Chronic dietary exposure of DEHP resulted in degenerative changes in the testis, decreased testicular zinc concentrations, lower sperm density and motility, and an increase in abnormal sperm morphology.129 Thomas and coworkers have reported that phthalate-induced zinc deficiency is consistent with germinal epithelial damage.124 Agarwal et al.128 examined the effects of dietary zinc on DEHP-induced testicular atrophy in male rats. DEHP was gavaged at three doses representing relatively nontoxic, mildly toxic and moderately toxic doses. Rats were fed a diet that was low (2 ppm), adequate (20 ppm), or high (200 ppm) in zinc. There was no effect of DEHP on testis, seminal vesicle, prostate, or epididymis weight from rats on normal or high zinc diets, but when rats were fed low zinc diets, DEHP reduced the weights of all organs in a dose-dependent

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manner. In the low zinc group, testes were characterize604d by degeneration, decreased lactate dehydrogenase activity, decreased total and free sulfhydryl contents, and decreased zinc concentrations. These data support the hypothesis that the effects of phthalates can be modulated by the zinc status of the animal. It is important to note that the developmental effects of zinc deficiency are attributed in part to an increase in tissue oxidative stress. Zinc has multiple roles in the oxidative defense system, and increases in reactive oxygen species concentrations are an early sign of cellular zinc deficiency.132 Increased oxidative damage and apoptosis have been observed in zinc deficient embryos, tissues (including the testes) and cells.133-136 Given that phthalates can induce oxidative damage and apoptosis, it is reasonable to speculate that suboptimal zinc status may increase the sensitivity of target cells to certain phthalates. The acute phase response can also induce changes in other nutrients in addition to zinc. For example, during infection, serum retinal, vitamin E, carotenoids (α-carotene and β-carotene) and folate can be significantly decreased during the acute phase response.137,138 Exposure to DEHP and DBP also leads to reductions in blood and soft tissue vitamin E concentrations.139,140 The severity of the testicular lesions induced by DEHP can be reduced in animals given supplemental vitamins C and E.141 Supplemental vitamin E can also reduce the developmental toxicity of DBP in rat embryonic limb bud cells.19 The above results are not surprising given the concept that certain phthalates, including DEHP and MEHP, are thought to be cytotoxic in part due to their ability to induce tissue oxidative stress.18,21 Thus, it is reasonable to suggest that acute phase response-induced changes in the antioxidant vitamins could result in an increased susceptibility to phthalate toxicity and teratogenicity. Additional consequences of an acute phase response are the up-regulation of the copper-binding protein, ceruloplasmin, and the down-regulation of the iron-binding protein, transferrin, resulting in hypercupremia and hypoferremia, respectively.74 The extent to which hypercupremia and hypoferremia may modulate the teratogenicity of phthalates has not been established. Low blood selenium, which is also noted during the acute phase response, is often associated with low cellular glutathione peroxidase activities.142 Given the critical roles of glutathione peroxidases in the oxidative defense system, it can be predicted that low selenium status will increase the risk for phthalate teratogenicity. This hypothesis, though, has not been tested. While the acute exposure of many drugs and agents that produce an acute phase response has been studied with regard to their teratogenic effects, there is considerably less information of the impact of chronic acute phase response stimulation on the developmental toxicity of teratogens. It should be noted that persistence of the acute phase response by chronic or repeated exposures could promote chronic inflammation. Repetitive or chronic episodes of acute phase response associated with inflammation and stress have been shown to be important independent predictors of the risk of developing chronic disease including cardiovascular diseases, diabetes and cancer.143-145 The effects of longterm alterations of mineral and vitamin levels induced by chronic, or prolonged, inflammatory acute phase response on health and pregnancy outcome clearly merit further investigation.

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17.3.4 CONCLUDING COMMENTS That a suboptimal diet can influence pregnancy outcome is now well established. Maternal micronutrient deficiencies can increase the risk for congenital malformations, including biochemical abnormalities that may result in an increased risk for chronic disease later in life. The extent to which maternal nutritional status modulates the response to certain reproductive and developmental insults is a subject of debate. Maternal suboptimal nutritional status is common, at times occurring even in the face of a seemingly adequate diet. Given the above we suggest that increased consideration should be given to the idea that potential teratogens should be examined in animals fed marginal diets, as well as in animals fed “control” nutritionally adequate diets. The extent to which phthalates induce teratogenesis in humans is of considerable controversy, and they represent one class of chemicals that in our opinion should be studied in nutritionally compromised, as well as non-compromised populations. 17.3.5 RECENT FINDINGS Sprague-Dawley rats were administered 0, 0.25, 0.50, or 1 g/kg/day of diundecyl phthalate or ditridecyl phthalate, by gavage, on gestation days 6–20.146 Diundecyl phthalate and ditridecyl phthalate had no adverse effects on maternal body weight and food consumption.146 The number of live fetuses, percent of post-implantation loss and of resorptions, fetal sex, and fetal body weights were not affected by either phthalate.146 There was no evidence of teratogenicity, whatever treatment.146 17.3.6 ACKNOWLEDGEMENTS This work was funded by a National Institute of Health grant HD-01743.

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17.4 PUBLIC HEALTH IMPLICATIONS OF PHTHALATES: A REVIEW OF FINDINGS FROM THE U.S. NATIONAL TOXICOLOGY PROGRAM'S EXPERT PANEL REPORTS Stephanie R. Miles-Richardson* Agency for Toxic Substances and Disease Registry, Office of Urban Affairs, Atlanta, Georgia, USA

17.4.1 INTRODUCTION Phthalates are ubiquitous, high-production-volume synthetic chemicals used in the manufacture of plastics and other consumer products. They are plasticizers which are compounded with plastic resins to increase their workability, flexibility, or extensibility.1 These compounds are used largely in the production of polyvinyl chloride, PVC,2 which is, in turn, used in a wide variety of consumer products. Di-(2-ethylhexyl) phthalate, DEHP, and dibutyl phthalate, DBP, the most commonly used plasticizers in PVC products, are found in items ranging from toys to swimming pool liners to medical tubing and blood storage bags.2,3 These plasticizers, as well as others, are also present in cosmetics, lubricants, floor carpets, and other consumer products.4 Globally, more than 18 billion pounds of phthalates are produced each year. Because of the high volume of production of phthalates and the wide variety of their commercial uses, they are ubiquitous and widely distributed throughout the environment.5-7 In water, phthalates will biodegrade, adsorb to sediments, and bioconcentrate in aquatic organisms.8 Though they have low estrogenic activity compared to estradiol and other potent estrogenic substances, phthalates may cause endocrine disruption though several mechanisms. At critical windows periods of development, they may disrupt biologic systems and also act as anti-androgens.9 Phthalate esters are a key additive in many plastics, therefore are an important constituent of numerous consumer products.10 Most of the mid- to high-molecular weight phthalate esters are used in the manufacturing of PVC to impart flexibility,11 while DBP is used in epoxy resins and cellulose esters and specialized adhesive formulations,12 DEHP is used extensively in disposable medical devices made of PVC.13 Plasticizers are also used in consumer products such as food packaging, bottled water, and children toys. Due to the widespread use of phthalates, it is important to consider human exposure and public health implications of such exposure. This review will focus on the following phthalates which were the subject of recent evaluations of risks to human reproduction by the U.S. National Toxicology Program, NTP, National Institutes of Environmental Health Sciences, NIEHS: butyl benzyl phthalate, BBP; di-isodecyl phthalate, DIDP; di-isononyl phthalate, DINP; DBP; di-n-hexyl phthalate, DnHP; di-n-octyl phthalate, DnOP; and DEHP.

*Writing performed by Stephanie R. Miles-Richardson in her private capacity. No official support or endorsement by CDC/ATSDR is intended or should be inferred.

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709

17.4.2 EXPOSURE TO ADULTS IN THE GENERAL POPULATION The primary route of human exposure to phthalates is oral. Food and foodstuffs are considered the largest source of exposure. Adult exposure to BBP, believed to be almost entirely due to food intake, ranges from 0.11-2.0 μg/kg bw/day. Estimates of foodborne exposure were obtained through a review of reports from the International Program on Chemical Safety, IPCS, and the United Kingdom's Ministry of Agriculture, Fisheries, and Food, MAFF,14 (whose duties have subsequently been overtaken by the Department for Environment, Food, and Rural Affairs).15 Of note, the IPCS report uses limited data that is nearly two decades old; the MAFF report was not clear with regard to descriptions of the calculations and assumptions used.14 Exposure to DBP, obtained from IPCS16 and Health Canada17 (both of which were based on a 1986 Canadian market-basket survey14), estimated DBP exposure at 7.0 and 1.9 μg/kg bw/day, respectively. After converting the MAFF-estimated mean (13 μg/person/day) and high level (31 μg/person/day) values using the IPCS body weight assumptions, exposure values obtained from MAFF were calculated at 0.20-0.48 μg/kg bw/day. Thus, according to three different sources, the range of exposure of adults to DBP was estimated at 0.2-7.0 μg/kg bw/day. The Agency for Toxic Substances and Disease Registry's, ATSDR, dietary estimate of 0.007-0.02 μg/kg bw/day, based on fish studies published between 1973 and 1987,18 is significantly lower than other exposure estimates available. The largest source of exposure of the general population to DEHP is also food. The second greatest exposure is via indoor air.19 The range of exposure from all sources except non-dietary, medical, and occupational was estimated to be 3-30 μg/kg bw/day.20 Fatty foods appear to provide the greatest dietary exposure to DEHP.16,19,21 Although DEHP can leach from food processing equipment and food wrappings into fatty foods and dairy products,19 because of international differences in food processing and packaging, dietary exposures may vary.19 Further, in recent years, plastics materials have been improved to minimize the leaching of plasticizers from the plastic containers by the food contents.7 The literature contains several examples of dietary exposure to DEHP. In Japan, when 63 samples of hospital food were examined for the presence of phthalates, DEHP was found at the highest level (10-4400 ng/g). Investigators found that on 2 days out of 21 days, the level of DEHP (1850 μg/day) exceeded the European Union's, EU, Total Dietary Intake, TDI, of 1.85 mg for a 50 kg person.22 Another study which evaluated plasticizer contamination of foods sold in prepared pack lunches found that 5 of 16 packed lunches had levels of DEHP which exceeded the EUs TDI.23 In both instances, disposable PVC gloves used during food preparation were suspected as the source of the high levels of DEHP. A survey evaluating the period from 1997-1999 demonstrated the presence of plasticizers used in cap-sealing resins for bottled foods in seven out of 21 samples on the Japanese domestic market and in 10 out of 61 imported samples.24 Few exposure data were available for DIDP, DnOP, DINP, and DnHP, the latter of which is used in PVC intended for food applications (i.e., bottle cap liners, seam cements).25 Therefore, the NTP Expert Panel did not calculate exposure estimates for these phthalates. It is anticipated that exposure to each does not exceed exposure levels calculated for DEHP.26-28 Other phthalate exposure in adults may result from environmen-

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tal uptake during crop cultivation or by migration from processing equipment or packaging materials for foods.14,25,28,29

17.4.3 EXPOSURE OF VULNERABLE SUB-POPULATIONS 17.4.3.1 Children In general, children are exposed to environmental chemicals at levels which exceed adult exposures. Pound for pound, children breathe more air, eat more food, and drink more water. Also, the likelihood of environmental exposure is increased because children play close to the ground, they spend lots of time outdoors, and they engage in a great deal of hand to mouth activities. Further, children are not as efficient in metabolizing and excreting environmental toxins and their brains, immune systems, and endocrine, and reproductive organs are growing and developing.30 Since phthalates are widely used to make plastics pliable, the migration of phthalates, particularly as babies and infants suck, chew, and otherwise mouth toys, is of concern. Because PVC used in toys and nipples is not bound to the polymeric matrix, there is the risk of ingestion of plasticizer leached by the sucking action of babies and the biting action that typically occurs in older infants who are beginning to teethe.1 As a result of this concern, in 1986, US toy manufacturers began voluntarily removing DEHP from pacifiers and nipples.27,31 DEHP is also no longer used in Canada in items such as pacifiers, rattles, teethers, and other toys intended for mouthing. However, it may still be present in larger toys used by older children.19 Estimated exposure levels for children are presented (Table 17.4.1). Table 17.4.1. Estimated exposure of children to phthalates Phthalate

Estimated exposure to children, mg/kg bw/day

Source of exposure

Age of child

BBP

0.2A 0.1A

formula formula

birth 6 months

DBP

5.0C 2.3C 2.4A 1.4A 0.11B 0.022B

food food formula formula drinking water drinking water

6 months to 4 years 12 to 19 years birth 6 months 0 to 6 months 12 to 19 years

DINP

54.4 to 84.5*

toys

not applicable#

A

NTP Expert Panel calculated using MAFF exposure level, manufacturer intake rates, and IPCS body weight assumption of 2.5-3.5 kg at birth and 7.5 at 6 months of age. B ICPS estimate C Health Canada estimate * maximum exposure assuming 8 kg body weight; 3 hours of exposure; 10 cm2 mouthing area (Fiala et al., 2000) # Adult volunteers used to simulate mouthing by children, (weight estimate for calculation was 8 kg; Fiala et al., 2000).

Because of concerns regarding potential health effects from DEHP exposure, many toy manufacturers have discontinued using DEHP in their products. It is also no longer used in the US in baby teethers and rattles and is no longer used as a plasticizer in plastic food wrap products. However, children may still be exposed to DEHP as a result of medi-

17.4 Public health implications of phthalates

711

cal procedures such as blood transfusions or hemodialysis during which DEHP may leach from plastic equipment into biological fluids. This may be particularly significant to neonatal children and effects may be more adverse in neonatal males.3 Although the largest source of phthalate exposure for the general public may be from DEHP, the most important plasticizer in terms of children's exposure is DINP since it is the primary plasticizer used in children's products.27,32 It is widely used to impart softness and flexibility to otherwise rigid PVC products.33 Several studies in the UK seeking to measure concentrations of DINP and DIDP in children's toys found varying levels of phthalates.26 However, DIDP concentrations have decreased from 30% (1990 UK survey), to 15% (1991 UK survey). Latter surveys of children's toys (1992 and 1996) demonstrated no detectable DIDP.26,34 In 1996, the US Consumer Product and Safety Commission, CPSC, reported DINP as the predominant phthalate in a sample of 35 toys. DIDP was not detected.26,35 Using both in vitro systems and adult volunteers, several studies have attempted to estimate the actual exposure of children to the phthalates present in children's products.27 Studies using 10 and 20 adult volunteers who mouthed on DINP-containing specimens reported mean migration rates of DINP at 120 and 241.3 μg/11 cm2/hour, respectively.27,32,36 Another similar study in Austria using 10 adult volunteers demonstrated greater migration rates when samples were chewed.27 Despite such estimations, no correlation between the release rate of DINP under experimental conditions and total DINP content of toys has been found. Nonetheless, investigators have attempted to calculate daily intake based upon the leaching rates observed. Due to differing analytical methodologies and exposure estimates, subsequent recommendations varied.27 To address discrepancies, further study emphasizing the standardization of techniques and correlation with in vivo simulations is needed, as well as additional information on the mouthing behavior of infants and children.27 Although in theory, exposure of children to BBP through mouthing toys is a potential source of phthalate exposure, according to the American Chemistry Council, BBP is not used in the production of toys.37 Children are also exposed to phthalates via dietary exposure, most often through the consumption of infant formula. A study analyzing phthalates in 29 diet samples, 11 baby food samples, and 11 samples of infant formula found one or more plasticizers in all of the diet samples (0.09-0.19 mg DBP/kg; 0.017-0.019 mg BBP/kg; 0.11-0.18 mg DEHP/kg and 0.13-0.14 mg DEHA/kg), as well as one or more phthalates in 50% of both the baby foods and infant formula samples.38 An unquantified and unspecified isomer of DHP was identified in baby formulas from the UK (7 of 12 formulas), and an unspecified DHP isomer was detected, at a level below the limit of detection of 0.01 mg/kg, in breast and commercial milk, cream, nuts, and baby food.28 Estimated DBP exposure levels from food ranged from 2.3 μg/kg bw/day to 5.0 mg/kg bw/day in children 12-19 years old and 6 months to 4 years of age, respectively.17,29 Based on a 1996 survey, the MAFF estimated exposure levels of DBP for infants at 2.4 μg/kg bw/day at birth and 1.4 μg/kg bw/day at 6 months of age.29,39 In a follow-up survey two years later, lower exposure levels were estimated.29,40 In 1996, the US reported DBP levels ranging from <0.005-0.011 mg/kg, approximately 10-fold lower than concentrations measured in the UK.29,41 Other reports of DBP in baby food, breast milk, and formula in Germany and Japan were found to be within the range of that reported by the MAFF.29 The MAFF also evaluated exposure to BBP via baby formulae. Exposure of infants and children to BBP was estimated at 0.2 μg/

712

Health and Safety Issues with Plasticizers and Plasticized Materials

kg bw/day at birth and 0.1 μg/kg bw/day at 6 months.14,39 Sampling of infant formulas by the US Food and Drug Administration suggests that phthalates are present less frequently in the US than in Europe.26,41

17.4.3.2 Women Children are not the only subpopulation with specific vulnerability to phthalates. A study found that women of reproductive age had significantly higher levels of monobutyl phthalate, a toxic metabolite which is found in urine, than other sexes. The higher phthalate levels in women are believed to be due to the fact that dibutyl phthalate is used in many beauty products including perfume, lotion, and nail polish.42,43 Animal models have been useful in providing insight on the potential mechanisms and subsequent health effects of phthalate exposure to women. Investigators propose that, using DEHP as a prototype and a rodent model, the metabolite MEHP acts through a receptor-mediated signaling pathway to suppress estradiol production in the ovary, leading to anovulation.42 17.4.3.3 Occupational Exposure Exposure to phthalates in the workplace can occur through the dermal and inhalation routes.27 Based on rodent studies, absorption of phthalates through the skin is believed to be minimal,26,27 however phthalates are lipophilic therefore are likely efficiently absorbed through human skin.44 Occupational exposure to phthalate esters is minimized by the fact that the compounds are manufactured under negative pressure within a closed system. Nonetheless, some exposures may occur during the loading and unloading of railroad cars and trucks.14,26-29 During the production of polyvinyl chloride PVC products, higher occupational exposure may be expected since such production involves elevated temperatures and more open processes.26 Exposures during the production of phthalates and PVC have been reported as less than 1 mg/m3 and 2 mg/m3, respectively.26,37 Assuming an inhalation rate of 10 m3/day and a 70 kg body weight, exposure estimates for persons employed in phthalate manufacturing and PVC productions are 143 and 286 μg/kg bw/day, respectively.14 Levels of DBP in US plants range from below detection limits to 0.08 mg/m3, still below the OSHA established permissible exposure limit of 5 mg/m3.29 Workers may be exposed via inhalation to relatively high concentrations of DEHP when it is compounded with PVC resins.19 Studies in Europe and the former USSR estimate exposures of DEHP are <2-6600 μg/kg bw/day.19,21 To meet OSHA workplace standards, DEHP air concentrations should not exceed 700 μg/kg bw/day.19 Phthalates, frequently used in the manufacture of epoxy resins, plasticizers, adhesives and a wide variety of other materials, have been identified as an important irritant and immunogen of at least four occupational respiratory syndromes.45 17.4.3.4 Medical Exposure Specialized equipment used for medical procedures often contains PVC components. Most often, the PVC used contains DEHP because this plasticizer imparts desirable flexibility and strength. For example, in PVC blood bags, DEHP may constitute as much as 40 percent of the plastic material.46 Other phthalates are not readily used in medical devices. In 1997, the IPCS reported a DBP level of 5 mg/g in plastic tubing used for oral/nasal feeding.16,29 There are no known uses of DnHP nor DNOP-containing medical devices.25,28 BBP is not approved by the FDA for use in medical devices.14

17.4 Public health implications of phthalates

713

DEHP is considered a suitable plasticizer to use in medical devices and procedures because of physico-chemical properties which allow for its use at a wide range of temperatures and makes it suitable for various sterilization processes. Also, medical equipment containing flexible DEHP as a plasticizer resists kinks, maintains optical clarity and barrier capability, and can be welded, centrifuged, and bonded.19 Some medical procedures that utilize devices and equipment which contains PVC include hemodialysis; storage and transfusion of whole blood and blood products; extracorporeal oxygenation; cardiopulmonary bypass; IV fluid administration; respiratory therapy; and enteral and parenteral feedings.19 Because of its lipophilicity, and because it is not chemically bound to PVC, DEHP will readily dissolve in lipid-containing fluids and solutions such as blood, plasma, and drug solutions.13,19,47 Newborn infants who undergo extracorporeal membrane oxygenation, ECMO, therapy are at a higher risk for exposure to high concentrations of DEHP because of the large surface area of tubing in the ECMO circuit.48 A prospective study of infants who had received ECMO therapy demonstrated that both DEHP levels and hemolysis were significantly associated with the degree of cholestasis, a condition which occurs frequently in neonates who have received ECMO therapy.49 Karle, et al.48 found that when the circuit surfaces are coated with heparin, DEHP does not leach and the amount is actually decreased. Hemodialysis patients are regularly exposed to high concentrations of DEHP, therefore such patients in chronic renal failure who undergo maintenance hemodialysis are at risk for toxicity due to chronic exposure.50 The occurrence of hepatotoxicity after use of PVC tubing in patients undergoing renal dialysis has been attributed to the release of DEHP from the tubing.51 Exposure of children to DEHP as a result of medical procedures such as blood transfusions or hemodialysis during which DEHP may leach from plastic equipment into biological fluids may be particularly significant to neonatal children and effects may be more adverse in neonatal males.3 The migration of DEHP from PVC blood bags is influenced by the duration of storage, and the temperature and composition of the blood or blood products. Consequently, newborn infants who receive blood transfusions are exposed to considerable quantities of DEHP and its primary metabolite, mono-(2-ethylhexyl) phthalate, MEHP.46 Exposure to DEHP in medical procedures varies greatly, and ranges from acute exposures (i.e., single blood transfusion) to chronic exposures in patients with long term therapy (i.e., maintenance dialysis patients).19 Very high exposures occur in newborn infants who have blood transfusions, cardiac surgery, and ECMO.19 Because of less efficient metabolism, preterm infants who have had mechanical ventilation utilizing PVC respiratory tubing systems may have subsequent toxic damage of the lungs due to inhalation exposure to DEHP.52

17.4.4 HEALTH EFFECTS OF PHTHALATE EXPOSURE Data of health effects resulting from exposure of humans to phthalates is limited. PVC production and exposure to PVC flooring has been associated with adverse health effects. A single German study found that the presence of PVC flooring was associated with a significant increase in the risk of bronchial obstruction during the first 2 years of life.14,53 An evaluation of young Puerto Rican girls with premature thelarche (breast development)9,54,55 demonstrated significantly high levels of phthalates in blood serum9 compared to girls who did not demonstrate premature thelarche.

714

Health and Safety Issues with Plasticizers and Plasticized Materials

In some cases occupational exposures to phthalates have been associated with adverse health effects. A case-control study among a population of Danish workers employed in PVC production for 5 years or greater demonstrated a significant increase in the risk of multiple myeloma.14,56 Phthalate mixtures containing BBP have been associated with respiratory or neurological effects and cancer,14,57 increased incidence of menstrual disorders, and spontaneous abortions.14 A study of college students with no known exposure to phthalates found a negative correlation between DBP concentration and sperm density or total number of sperm.29,58 However, this study did not consider confounders nor did it provide ample evidence for a causal relationship of sperm characteristics to DBP levels.29 There were no human health data available for DEHP, DIDP, DINP, DnHP, nor DnOP. The majority of data available to evaluate potential human health effects of phthalates comes from studies of experimental animals. Adverse health effects in animal models are well characterized,14,25-29 particularly following oral exposure to DEHP.19 The NTP Expert Panel reviewed health effects data in terms of general, developmental, and reproductive toxicity. Adequate information was available to identify DEHP and its metabolites as a developmental toxicant by the oral route. However, when using animal data to make inferences about human exposure, one must consider interspecies differences. For example, there are differences in the rate of absorption of metabolites following oral exposure to DEHP across different species. Specifically, rodents have higher levels of intestinal lipases necessary for the production of the metabolite MEHP. As a result, higher absorption of DEHP occurs in rodents than in primates, a species with lower levels of intestinal lipases. Similarly, adults are more efficient at metabolizing DEHP than children under 6 months of age.19 To illustrate this point, DEHP exposure in mice and rats produces effects in the liver and testes, while similar exposure in cynomolgus monkeys and marmosets do not produce such effects. Most hepatic effects observed in rodents occur as a result of peroxisome proliferation. This mechanism does not occur in primates or only occurs to limited extent subsequent to exposure to high doses.19 Nonetheless, species differences in the sensitivity to peroxisome proliferation are not believed to be relevant to extrapolation of developmental toxicity to humans.19 Another important difference in metabolism is that the glucuronidation pathways in humans, which do not mature until infants are 3 months old, are necessary to excrete MEHP as a glucuronide. This is also true in primates, but not in rodents who do not metabolize MEHP further. The developmental toxicity database used by the Expert Panel to evaluate DEHP consists of oral exposure studies of rats and mice exposed during gestation. Effects were noted during prenatal physical development of pups. Effects included the following: tail malformations, axial and appendicular skeletal abnormalities, cardiovascular malformations, and neural tube closure defects, developmental delays, and intrauterine death. Oral exposure of DEHP can affect reproductive processes of male and female rats and mice. In females, effects on estradiol synthesis and ovulation in rats were observed, but no structural changes were seen in the uterus or vagina.19,59 In males, the Sertoli cell is a cellular target for neonatal, pubertal, and adult male exposures and testicular pathology and reduced sperm numbers have been observed. The reproductive database used by the Expert Panel to evaluate DEHP consisted of adverse effects in rats, mice, guinea pigs, and

715

17.4 Public health implications of phthalates

ferrets. Most studies focused on the pubertal and adult male rodent and identified structural changes in the testis, reduced fertility and altered sperm measures.

17.4.5 US NTP EXPERT PANEL CONCLUSIONS Expert Panel conclusions are summarized in Table 17.4.2. Since primates showed no testis effects when exposed to DEHP at oral doses similar to those which would cause testicular toxicity in juvenile rodents, the Expert Panel has a minimal concern that ambient human exposures adversely affect adult human reproduction. However, exposure of healthy infants and toddlers, whose reproductive systems are still developing, is more of concern. Thus, if an infant or toddler is exposed to levels higher than that expected with ambient adult exposure, the Expert Panel has concern that exposure may adversely affect male reproductive tract development. Because it is possible that exposures of critically ill male infants who are exposed via medical therapy can approach doses that are toxic in rodent models, the Expert Panel has serious concern that such exposure may adversely affect male reproductive tract development. Because oral exposure for humans is estimated at <30 μg/kg bw/day and toxic effects are observed in rodents at >3 mg/kg bw/day, the Expert Panel has concern that ambient oral DEHP exposures of pregnant or lactating women may adversely affect the development of their offspring. Table 17.4.2. NTP phthalate Expert Panel findings Phthalate

Level of concern

Toxicity

Population at risk

BBP

negligible

reproductive

adult (male)

DIDP

minimal minimal

reproductive developmental

adult/child fetus/child

DINP

minimal low minimal

developmental reproductive/developmental reproductive

fetus child adult

DBP

negligible minimal

reproductive reproductive/developmental

adult adult/child

minimal negligible

reproductive

fetus/child adult

minimal concern serious concern concern

reproductive reproductive/developmental reproductive/developmental developmental

adult child (healthy male) child (critically ill#) fetus

DnHP* DnOP DEHP

* #

Insufficient data available to ascertain risk. Infants with extensive parenteral medical exposure.

Based on data available, the Expert Panel concludes that oral exposure to BBP can cause developmental toxicity in rats and mice, and reproductive toxicity in rats.14 Due to the low exposures expected to occur in adults in the general population (2 g/kg bw/day) and the high dose identified as the NOAEL (182-185 mg/kg bw/day), the expert panel concluded that there is negligible concern for male reproductive effects from adult expo-

716

Health and Safety Issues with Plasticizers and Plasticized Materials

sure to BBP.14 Oral exposure to DBP, which shares a common monoester metabolite with BBP, causes structural malformations leading to developmental toxicity in rats and mice. Based on estimated exposure of the general population to DBP (2-10 g/kg bw/day) and a NOAEL (for oral exposure in the rat) of 50 mg/kg bw/day, the NTP expert panel has minimal concern about effects to human development and development of the reproductive system.29 The Expert Panel also indicates minimal concern that DIDP is a human reproductive toxicant because the NOAEL for reproductive toxicity ranges from 427-929 mg/ kg bw/day.26 The limited exposure and toxicological data available for DnOP suggest that it is not a potent developmental or reproductive toxicant in rodents therefore, the Expert Panel has assigned negligible concern for its effects on the adult reproductive system.25 Data available for DnHP suggest that it is a developmental and reproductive toxicant in rodents. However, the limited quantitative information prevented the Expert Panel from being able to determine the potential for risk to human reproduction.28 Finally, the Expert Panel has minimal concern for unborn children due to ambient maternal exposure to DINP and low concern for potential health effects in children due to sucking and chewing of toys and other objects.

17.4.6 PUBLIC HEALTH IMPLICATIONS When considering public health implications, it is necessary to consider populations who are most vulnerable. Indeed the measure of a successful public health program is the measure of the health of those most at risk. Susceptibility is an all important aspect of environmental health risk assessment, particularly when pregnant women and their fetuses, infants and children, the elderly, and the infirm are known or suspected to be more at risk.60 When considering the plethora of animal data and the limited human data on phthalate exposure, it is clear that the most vulnerable population consists of infants and children − particularly critically ill male infants exposed to phthalates subsequent to medical interventions. While data is not sufficient to assign risk to pregnant women and their fetuses, or to young children who chew and suck phthalate containing toys and objects, enough information exists to warrant concern for both populations. Persons on maintenance medical therapy who are exposed to high levels of phthalates via medical equipment and procedures are another population who might be considered at risk (i.e. chronic dialysis patients). However, the relationship between the magnitudes of exposure and subsequent adverse health effects needs further study. Based on current literature, the most adverse effects of phthalate exposure is believed to occur in infants undergoing intensive medical therapy. Such medical therapy exposes this still developing subpopulation to phthalates − primarily DEHP − which is the predominant plasticizer used in medical devices. Infants are most at risk because of their developing reproductive systems and their immature glucuronidation pathways which render them less able to excrete DEHP metabolites. The implications are greater in male neonates since research has demonstrated reproductive toxicity which targets the testis and Sertoli cells. Nonetheless, in such cases, the medical intervention may be necessary to preserve life. Therefore the risk seems warranted. Notwithstanding, it is incumbent upon medical providers to fully inform the parents/guardians of such critically ill infants of both the potential benefit of therapy as well as the potential inadvertent risk. Many critical data needs were identified by the Expert Panel in its series of reports.25-29 Human data is quite limited, and exposure data is not nearly as robust as is

References

717

needed to adequately characterize human exposure. From a public health perspective, the most urgent needs involve gathering additional data on the levels of specific phthalates in toys and better estimates on salivary extracts of phthalates commonly used in toys (i.e., DINP; DEHP) to further characterize the risk of exposure to children, particularly 3-12 month old children who exhibit a great deal of mouthing behavior. Also, additional studies are needed to elucidate potential developmental effects which may occur when pregnant or nursing women are exposed to ambient or greater concentrations of phthalates.

17.4.7 ADDENDUM BASED ON UPDATED LITERATURE The developments and industrial activities in the Middle East, especially Saudi Arabia, are expected to contribute to the natural, regional, and anthropogenic input sources of organic matter.61 City of Dhahran, Saudi Arabia had measurable concentration of plasticizers of 131+/-119 ng m-3.61 Plasticizers were one of the major components detected in air.61 Compared with the data above, the exposure in Dhahran does not exceed European permissible levels of exposure.61 Despite the fact that some phthalates are hydrophobic and have low water solubility (e.g. di-ethylhexyl-phthalate (DEHP) (log Kow = 7.60 and S = 0.27 mg/L), they have become persistent in environment and they have spread into the aquatic environment.62 DEHP is considered a priority substance in the EU.62 Its presence in drinking water is also regulated by EPA.62 DEHP was detected in cave streams in the US and in groundwater used as a drinking water source in Mexico City (19–232 ng/L).63 Elevated concentrations of DEHP (46 µg/L) and other phthalates, i.e., diethyl phthalate (DEP) (1.5 µg/L), and dimethyl phthalate (DMP) (380 ng/L), were present in groundwater from the Chalk (UK).64 DEHP and d-n-butyl phthalate (DBP) were occasionally above 1 µg/L in groundwater intended for drinking water in China.65 N-butyl benzene sulfonamide was one of the most ubiquitous and abundant (240 µg/L) plasticizers present in UK groundwater.64 Bis(2-methoxyethyl) phthalate was detected in all of the samples blood plasma samples (153 in total) analyzed for Hong Kong population.66 DEHP, DnOP, diisobutyl phthalate and dibutyl phthalate were detected in over 95% of the samples, whereas the concentrations of diisodecyl phthalate and dihexyl phthalate were below the limit of detection.66 The mean concentration of DEHP was the highest among the eight phthalates (11.13 ng/ml ±3.95), followed by DMEP (11.01 ng/ml ±6.57), DnOP (6.53 ng/ml ±3.95) and DiBP (5.84 ng/ml ±2.64).66 The mean concentration of total phthalates in all of the human samples was 41.63 ng/ml ±10.62, ranging from 20.19 ng/ml to 84.98 ng/ml.66 Compared to a study in Sweden,67 the plasma DEHP, DnOP, DBP and DEP levels of our samples were 1.9–7.2 times higher.66

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41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67

719

Kingdom, Ministry of Agriculture, Fisheries, and Food, 1998. Department of Health and Human Services UPHS. Phthalates in infant formula-assignment summary,1996. Lovekamp-Swan T, Davis BJ, Environ Health Perspect, 111, 139-145 (2003). Blount BC, Milgram KE, Silva MJ, Malek NA, Reidy JA, Needham LL, Brock JW, Anal Chem, 72, 4127-4134 (2000). Jobling S, Reynolds T, White R, Parker MG, Sumpter JP, Environ Health Perspect, 103, 582-587 (1995). Bardana EJ, Jr., Andrach RH, Eur J Respir Dis, 64, 241-251 (1983). Sjoberg PO, Bondesson UG, Sedin EG, Gustafsson JP, Transfusion, 25, 424-428 (1985). Rock G, Labow RS, Tocchi M, Environ Health Perspect, 65, 309-316 (1986). Karle VA, Short BL, Martin GR, Bulas DI, Getson PR, Luban NLC, O'Brien AM, Rubin RJ, Crit Care Med, 25, 696-703 (1997). A prospective analysis of cholestasis in infants supported with extracorporeal membrane oxygenation, J Pediatric Gastroenterology Nutrition, 13, 285-289 (1991). Faouzi MA, Dine T, Gressier B, Kambia K, Luyckx M, Pagniez D, Brunet C, Cazin M, Belabed A, Cazin JC, Intl J Pharmaceutics, 180, 113-121 (1999). Faouzi MA, Dine T, Luyckx M, Gressier B, Goudaliez F, Mallevais ML, Brunet C, Cazin M, Cazin JC, Intl J Pharmaceutics, 105, 89-93 (1994). Roth B, Herkenrath P, Lehmann HJ, Ohles HD, Homig HJ, Benz-Bohm G, Kreuder J, Younossi-Hartenstein A, Eur J Pediatr, 147, 41-46 (1988). Jaakkola JJ, Oie L, Nafstad P, Botten G, Samuelsen SO, Magnus P, Am J Public Health, 89, 188-192 (1999). Perez A, Lancet, 1, 1299-1300 (1982). Bongiovanni AM, J Pediatr, 103, 245-246 (1983). Butte W, Heinzow B, Rev Environ Contam Toxicol, 175, 1-46 (2002). IPCS. Concise international chemical assessment document 17- butyl benzyl phthalate. Geneva, Switzerland, World Health Organization,1999. Murature DA, Tang SY, Steinhardt G, Dougherty RC, Biomed Environ Mass Spectrom, 14, 473-477 (1987). Davis BJ, Maronpot RR, Heindel JJ, Toxicol Appl Pharmacol, 128, 216-223 (1994). Sexton K, Environ Toxicol Pharmacol, 4, 261-269 (1997). Rushdi, A I; El-Mubarak, A H; Lijotra, L; Al-Otaibi, M T; Qurban, M A; Al-Mutlaq, K F; Simoneit, B R T, Arabian J. Chem., 2014 in press. Postigo, C; Barcelo, D, Sci. Total Environ., 503-504, 32-47, 2015. Félix-Cañedo, T E; Durán-Álvarez, J C; Jiménez-Cisneros, B, Sci. Total Environ., 454-455, 109-18, 2013. Stuart, M; Lapworth, D; Thomas, J; Edwards, L, Sci. Total Environ., 468-469, 564-77, 2014. Liu, X; Shi, J; Bo, T; Zhang, H; Wu, W; Chen, Q, Environ. Pollut., 184, 262-70, 2014. Wan, H T; Leung, P Y; Zhao, Y G; Wei, X; Wong, M H; Wong, C K C, J. Hazardous Mater., 261, 763-9, 2013. Hogberg, J; Hanberg, A; Berglund, M; Skerfving, S; Remberger, M; Calafat, A M; Filipsson, A F; Jansson, B; Johansson, N; Appelgren, M; Hakansson, H, Environ. Health Perspect., 116, 334-9, 2008.

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Health and Safety Issues with Plasticizers and Plasticized Materials

17.5 PLASTICIZERS IN THE INDOOR ENVIRONMENT Werner Butte School of Mathematics and Natural Sciences, Department of Pure and Applied Chemistry, University of Oldenburg, P.O. Box 2503, D 26111 Oldenburg, Germany

17.5.1 INTRODUCTION On average, people in moderate climates are supposed to spend up to 95% of their time indoors.1 Residents of the Federal Republic of Germany, depending on the season and vocational activity, stay between 80% and 90% of their time in an indoor environment and most of this time is spent in their homes.2 The National Human Activity Pattern Survey, NHAPS, of the USA recorded adults to spend an average of 87% of their time in enclosed buildings and about 6% in enclosed vehicles.3 Changes in building design intended to improve energy efficiency have meant that modern homes are frequently more airtight than older structures.4 As a result there has been increasing concern over the effects of indoor contamination on health during the last two decades. But still we know much less about the health risks from indoor air pollution than we do about those attributable to the contamination of outdoor air.4 Several studies have shown that for inhabitants, especially children and other vulnerable subgroups, the home environment may be a dominant source of exposure to toxicants.5 Thus indoor pollution has been ranked by the United States Environmental Protection Agency Advisory Board and the Center for Disease Control (CDC) as a high environmental risk.6 A classification of organic indoor contaminants, according to their volatility, was given by a WHO working group on organic indoor air pollutants.7 This group initiated the common practice of classification of organic chemicals according to their boiling points (Table 17.5.1). Table 17.5.1. Classification of organic indoor pollutants [after the WHO7] Abbreviation

Description

Boiling point range, ºC

Examples

VVOC

very volatile (gaseous) < 0 to 50-100 organic compounds

carbon monoxide, carbon dioxide, formaldehyde

VOC

volatile organic compounds

50-100 to 240-260

solvents (aliphatic, aromatic), terpenes

SVOC

semi-volatile organic compounds

240-260 to 380-400

pesticides (chlorpyrifos, pentachlorophenol, etc.), plasticizers (phthalates, such as, DEP, DBP; organophosphates, such as, TEP, TCEP, TBP)

POM

particulate organic matter

> 380

pesticides (pyrethroids), polycyclic aromatic hydrocarbons, phthalates and organophosphates having high boiling points (e.g., TCP, DEHP)

721

17.5 Plasticizers in the indoor environment

VVOC (very volatile organic compounds) and VOC (volatile organic compounds) are transitory and predominantly found in air; organic pollutants with low volatility or high polarity are SVOC (semi-volatile organic compounds) or POM (particulate organic matter). Phthalate or organophosphate plasticizers may either be semi-volatile or “nonvolatile”. For example boiling points of DEP: 298°C and DBP: 340°C8 or TCEP: 330°C9 classify these plasticizers as semi-volatile. SVOC are supposed to partition between air and house dust, whereas plasticizers with boiling points >380°C like DEHP: 385°C9 or TCP: 410°C10 are POM and are found predominantly in house dust. Exposure of residents to phthalate and organophosphate plasticizers mainly results from indoor air and house dust; compounds and their abbreviations considered in this article are listed in the addendum.

17.5.2 SOURCES OF INDOOR PLASTICIZERS Plasticizers are mixed into polymers to increase flexibility and workability. The phosphoric acid esters, i.e., organophosphates, are also remarkable flame retardants, and for this reason they are extensively used in plastics.11 At present, some 300 plasticizers are manufactured, of which at least 100 are of commercial importance.12 Most reports on plasticizers known to be present in the indoor environment are either on phthalate esters (phthalates) or on phosphate esters (organophosphates) (many of the so-called alternative plasticizers reach high production rates).109 Their production capacities are high, they may reach from some thousand tons per year (e.g., TCEP) up to some million tons (DEHP) (Table 17.5.2). Table 17.5.2. Production of some plasticizers

Compound

Annual production, t

References

Adipates DEHA

10,000-100,000 (in 2014)

ECHA107 EPA108

DIDA

1,000-10,000 in 2014 in EU

ECHA107

10.000-100,000 in 2014

ECHA107 EPA108

10,000-100,000 in 2014 in EU

ECHA107

Benzoates DPGDB Citrates ATBC

Cyclohexane dicarboxylic acid 200,000 in 2013

Bui109

DEHP

3,000,000-4,000,000 (in 1985) ~ 8,000,000 (no statement of year)

Wams13 Blount14

DEP

30,000 (in 1995)

Rippen9

DEHT

10,000-100,000 2011-2014

Bui109

20,000 (in 1994)

Rippen9

DINCH Phthalates

Organophosphates TBP

722

Health and Safety Issues with Plasticizers and Plasticized Materials

Table 17.5.2. Production of some plasticizers

Compound

Annual production, t

References

TCEP

~4,000 (in 1997) > 3,000 (in 1999)

IPCS11 Rippen9

TCP

30,000 (in 1994); 1,000-10,000 in EU in 2014

Rippen9, Bui109

TCPP

> 40,000 (in 1997)

IPCS11

TDCPP

8,000 (in 1997)

IPCS11

TPP

40,000 (in 1994)

Rippen9

1,000-10,000 in EU in 2014

Bui109

10,000-100,000 in EU in 2014

ECHA107

Sebacates DOS Trimellitates TOTM

Vegetable oil derivatives COMGHA

1,000-10,000 in EU in 2014

ECHA107

ESBO

10,000-100,000 in EU in 2014

ECHA107

More than 1% for DEHP9 and about 5% for DEP and DBP9 are estimated to be dispersed into the environment. For organophosphates, rates are even higher. They reach from about 50% for TBP and TCP9 and up to 90% for TCEP,9 respectively. Dialkyl and alkyl aryl esters of phthalic acid, i.e. phthalates, are ubiquitous industrial chemicals with a wide range of applications. Phthalates are primarily used as plasticizers in polyvinylchloride, PVC, products. DEHP, DiNP, DiDP are the general purpose plasticizers for PVC in most applications. For wire and cable, DiDP is preferred.15 87% of the phthalates produced are used for formulating flexible PVC, which is consumed for manufacturing the following goods: wire and cable: 25%, film and sheeting: 23%, flooring: 15%, plastisol spread coatings: 11%, profiles and tubing: 10%, other plastisols: 8%, miscellaneous (shoe soles, blood bags, gloves): 8%.15 DBP and BBP are fast-fusing plasticizers for PVC. They are mostly used in combination with DEHP. BBP is further present in cosmetics, such as hair sprays containing 0.1 to 1% BBP.16 C1 to C4 phthalates are mainly used as plasticizers for cellulose resins and some vinyl ester resins. C4 phthalates are also appropriate plasticizers for nitrocellulose lacquers. Further non-polymeric uses of phthalates are fixatives, detergents, lubricating oils and solvents in products such as cosmetics and wood finishes17 as well as additives in insect repellents.8 Phthalates are also reported to be present in textiles, such as cotton diapers, bathrobes, T-shirts, upholstery fabric, and carpeted floor at concentrations of a few mg/kg.18 Plasticizers in general are further additives in modern electronic goods such as TV sets, computers, copying machines, etc.19 Organophosphates, i.e., trihaloalkyl-, trialkyl- and triaryl phosphates, have a variety of uses as flame retardant plasticizers. Trialkyl phosphates, such as, TEP are serving as flame retardants and plasticizers of polyurethane foam. Following the Montreal protocol, chlorofluorocarbons (CFC) have been phased out. The only blowing agents presently available are flammable compounds namely n-pentane, iso-pentane and cyclopentane.

17.5 Plasticizers in the indoor environment

723

Thus flame retardants, such as organophosphates, must be introduced to improve flame retarding and mechanical properties.20 Aryl phosphate plasticizers are utilized in PVC-based products. The principle advantage of phosphate esters, such as TCP, as plasticizers for PVC is their low volatility and the ability to impart fire-retardant properties to a PVC formulation. Also TEHP shows good compatibility with PVC and it imparts good low-temperature performance in addition to a good fire retarding properties. DPEHP has widespread use due to its combination of plasticizing efficiency, low-temperature properties, migration resistance, and fire retardancy. Miscellaneous applications of aryl phosphates are as pigment dispersants and peroxide carriers, and as additives in adhesives, lacquer coatings and wood preservatives.21 Halogenated phosphorus flame retardants, i.e., trihaloalkyl phosphates, combine the flame-retarding properties of both the halogen and the phosphorus group. One of the largest selling members of this group, TCPP is used in polyurethane foam. TCEP is utilized in the manufacture of polyester resins, polyacrylates, polyurethanes and cellulose derivatives. The most widely used bromine- and phosphorus-containing flame retarding plasticizer used to be tris(2,3-dibromopropyl) phosphate, but it was withdrawn from use in many countries due to the carcinogenic properties in animals.21 Soft foams, paints and wallpapers mainly contain TCEP, insulation sealant foams mainly TCPP.22 TCEP is further present in coatings of sound insulation panels23 leading to high concentrations in indoor air and dust.24 To summarize: organophosphates are present in adhesives, cellulose acetate, coatings, lacquers, latexes, lubricants, polymers like PVC, polymeric resins (phenolic and phenylene-oxide-based), as well as rigid and flexible polyurethane foam.11,25-27 Materials listed above are used in manufacture of electrical and automobile components and for goods used in the indoor environment, such as electronic devices (television sets, video recorders, computers etc.), furniture, and upholstery. These goods are the major source for organophosphates indoors. PVC utilizes 85% of the total production of phthalate plasticizers with DEHP being the most important.13 PVC is used in production of furniture, flooring and wall covering, cables, building and construction parts but also shower curtains, footwear, plastic bags, food-packing materials, toys, etc. The DEHP content of PVC varies, depending on the application of the plastic, but in general it is between 20 and 40 wt%.13 Indoor contamination with plasticizers results mainly from leaching. However, it is difficult to estimate the loss from plastics in use. The rate of migration depends on the characteristics of the particular plastic material and on the medium with which the plastic is in contact. If DEHP is formulated in paints, 15% DEHP is known to evaporate into the atmosphere.13 Phthalates (DBP, DEHP) and flame retarding organophosphates (TCEP) emitted from television sets and video recorders may amount to some µg per hour.29,30 Emission into the surroundings (in this case the indoor environment) is intensifying with temperature increasing (e.g., sunlight directly shining on a black television set). Polyurethane foam samples for building and indoor use have shown specific emission rates of the degradation products of TEP, TCEP TCPP and TCDPP, i.e., chloroethane, dichloroethane, and chloropropanol, of about a hundred µg/(m2 h); the specific emission rate for TEP was of the same order of magnitude.31 But an even higher likelihood of plasticizer escape into the

724

Health and Safety Issues with Plasticizers and Plasticized Materials

indoor environment results from the non-plasticizer use phthalates, e.g., as pesticide carriers, in cosmetics, fragrances, oils, and insect repellents. The results of testing material samples in an indoor environment demonstrate that polyurethane soft foams, insulation foams, mattresses, paints and finishes are the major indoor sources for TCEP, TCPP and TDCPP. Concentrations of organophosphate plasticizers in an indoor environment have been reported by Ingerowski et al.,22 Pardemann et al.,33 and Rippen.9 The quantities emitted are of the order of a few milligrams per kilogram of material, such as coated wood, carpets, or wallpaper coated with PVC. A review of primary sources of TCEP and TCPP in the indoor environment was given by Ingerowski et al.22 Maximum concentrations are compiled in Table 17.5.3. Table 17.5.3. Maximum concentrations of TCEP and TCPP which were emitted into the indoor environment from different materials [Data from Ingerowski et al.22] Material Wood preservation coatings Mattresses (polyurethane) Wallpaper (glass fiber) Carpet backing (polyurethane)

TCEP, mg/kg

TCPP, mg/kg

10,000

150

890

1,500

2,400

1,100

-

13,100

Polyurethane soft foam

19,800

-

Foam fillers (polyurethane)

32,000

180,000

-

220

68,000

-

Floor sealing material Acoustic ceilings (coating)

TCEP production in Germany has been discontinued in 1997.9 TCEP is no longer used in soft foams for mattresses and upholstery.34

17.5.3 OCCURRENCE OF PLASTICIZERS INDOORS 17.5.3.1 Indoor Air The analysis of plasticizers in indoor air is most frequently done using either adsorbent tubes (without a front filter) or sample trains. Sample trains consist of a filter followed by an adsorbent. They are intended to separate particles bound from gaseous plasticizers. Filters are made either of quartz35-39 or glass fiber.39-41 Adsorbents to trap the vapor phase plasticizers may be XAD 2,35,38 C18 extraction disks,37 polyurethane foam,22,24,35,36,39-41 or charcoal.42,43 After air sampling, the filters and sorbents are subsequently extracted and the plasticizers are analyzed using capillary gas chromatography. Either a flame photometric detector,37,39,43 or a mass spectrometer24,35,36,38,39,43 are used for quantification. Depending on the sampled volume, the air detection limits for phthalates (e.g., BBP and DBP)38 in indoor air may reach 0.04 ng/m3 or 1 ng/m3 for organophosphates (e.g., TCEP),22,24 respectively. But, the analytical methods used to determine plasticizers in air also have problems, mainly regarding blank samples. Prevention of contamination during sampling and sample processing is necessary.44 Analytical reagents may be contaminated with traces of TCP because of its widespread use.25 TCPP and TDCP were found in coconut shell-based, acti-

725

17.5 Plasticizers in the indoor environment

vated charcoal sorbent tubes, which resulted in high field blanks, especially for TCPP.45 High blanks were also reported for DEHP analyses in indoor air.42 Concentrations of plasticizers in indoor air are normally in the ng/m3 range. Plasticizers having lower boiling points and higher vapor pressures (e.g., DEP, DBP, TBP or TCEP) are expected to show higher concentrations than plasticizers having higher boiling points and lower vapor pressures (e.g., DEHP or TCP). For example, the maximum concentration of DOP in air at 25°C was only 10 µg/m3.12 Typical concentrations for phthalates in indoor air measured in different countries are complied in Table 17.5.4. Table 17.5.4. Concentrations of phthalates in indoor air in ng/m3 Compound Median/Mean BBP

DBP

DiBP

Range

Circumstances

Reference

<100 (median) <100-750

Private homes in Northern Ger- Hostrup & Butte46 many (2000) with high concentrations of phthalates in house dust, n = 24

-

<1.2-100

Pilot study in 6 contemporary Japanese houses

72 (median)

10-172

One office, 5 homes in Massachu- Rudel et al.35 setts, n = 6

100 (mean)

11.6-581

Child care centers in North Caro- Wilson et al.38 lina (USA), n = 10

829 (median)

572-1346

Office, classroom and room in a Clausen et al.47 day-care center, Denmark, n = 12

-

92-308

Results of a field trial of measur- Hino & ing phthalates in air (one house, Nakayama42 one year; Japan)

630 (median)

130-2160

Private homes in Northern Ger- Hostrup & Butte46 many (2000) with high concentrations of phthalates in house dust, n = 24

-

110-600

Pilot study in 6 contemporary Jap- Otake et al.43 anese houses

251 (median)

101-431

One office, 5 homes in Massachu- Rudel et al.35 setts, n = 6

420 (median)

1300 (90. percentile)

Californian homes, n = 125

239 (mean)

108-404

Child care centers in North Caro- Wilson et al.38 lina (USA), n = 10

0.37-4.0

Japanese homes

390 (median)

120-1580

Private homes in Northern Ger- Hostrup & Butte46 many (2000) with high concentrations of phthalates in house dust, n = 24

49 (mean)

11-108

One office, 5 homes in Massachu- Rudel et al.35 setts, n = 6

Otake et al.43

Sheldon et al.48

Takeuchi112

726

Health and Safety Issues with Plasticizers and Plasticized Materials

Table 17.5.4. Concentrations of phthalates in indoor air in ng/m3 Compound Median/Mean

Range

Circumstances

Reference

DCHP

-

<1.2-170

Pilot study in 6 contemporary Jap- Otake et al.43 anese houses

DEHP

258 (median)

111-1053

Office, classroom and room in a Clausen et al.47 day-care center, Denmark, n = 12

200 (median)

90-1040

Private homes in Northern Ger- Hostrup & Butte46 many (2000) with high concentrations of phthalates in house dust, n = 24

-

40-230

Pilot study in 6 contemporary Jap- Otake et al.43 anese houses

61 (mean)

20-114

One office, 5 homes in Massachu- Rudel et al.35 setts, n = 6

110 (median)

240 (90. percentile)

Californian homes, n = 125

-

50-190

Pilot study in 6 contemporary Jap- Otake et al.43 anese houses

793 (mean)

236-1290

One office, 5 homes in Massachu- Rudel et al.35 setts, n = 6

DEP

TCP

150

car dust

Sheldon et al.48

Brommer et al.110

Regarding DEHP, concentrations can reach 1 mg/m3 (= 1,000,000 ng/m3) in the air inside cars and up to 150-260 µg/m3 (= 150,000 to 260,000 ng/m3) in rooms with a new floor and/or wall covering.49 In a Norwegian study,50 particulate matter suspended in indoor air in six dwellings in Oslo has been sampled and the phthalate content in micrograms phthalate per 100 mg of suspended matter was reported (BBP, DBP, DiBP, DEHP, DEP). However, calculations of concentrations of phthalates in air on the basis of suspended matter gave concentrations of only some ng/m3 with maximum values hardly more than 20 times this value. Data from Germany, Japan and Sweden are available for indoor air concentrations of organophosphates. But, there is no data based on large collectives, thus it is not possible to give medians and ranges of concentrations to the data included in Table 17.5.4 for phthalates in indoor air. Hansen et al.24 analyzed TCEP, TBEP TPP and p-TCP (para tricresyl phosphate) in indoor air. They only found TCEP (detection limit: 1 ng/m3). In rooms supposed to have no sources for organophosphates, TCEP concentrations were about 20 ng/ m3 (= 0.02 µg/m3), in rooms containing materials with TCEP, its concentrations between 300 and 3900 ng/m3 were observed. Ingerowski et al.22 found TCEP concentrations of up to 6000 ng/m3 in indoor air (median: 10 ng/m3, n = 50), but gave no description of the houses. Sagunski et al.36 reported that about 3/4 of TCEP in indoor air was gaseous, and 1/ 4 of it was particle bound (total concentration 39 ng/m3). Air samples from a Swedish office building, a day care center and three school buildings analyzed for TBEP, TBP, TCEP, TCPP, TPP, TEHP showed concentrations between 0.4 and about 30 ng/m3 with TCEP having the highest concentration measured (250 ng/m3).39 Saito et al.37 reported concentrations of TBEP, TBP, TCEP, TCPP, TEP, and TPP in 44 rooms of 22 houses and

17.5 Plasticizers in the indoor environment

727

22 offices of 11 buildings compared to 17 locations outdoors. The compound showing the highest medians in air of houses was TBP with 12.0 ng/m3, while in offices TCEPP had the highest concentration at 19.3 ng/m3. The maximum concentration was obtained for TCPP in houses, it was as high as 14,000 ng/m3 (115 times higher than in offices). The median concentration ratio of indoor to outdoor of these six organophosphates varied from 2.1 to 18.3.37 Yoshida and Matsunaga51 monitored organophosphates in the air of 10 residences in Osaka. TBP, TCEP, TCP and TPP were detected indoors at the levels of 0.5130.6, 0.6-18.9, 0.4-4.9, and 0.1-1.8 ng/m3, respectively. In a pilot study conducted in six houses, TBP and TCEP was detected in two houses only at concentrations between 10 and 100 ng/m3, whereas TBEP and TPP were present in one house only, each with a concentration of 10 ng/m3.43 Outdoor air concentrations for phthalates and organophosphates tend to be about a factor of 10 lower than their concentrations in the indoor air, they normally do not exceed a few ng/m3. DEHP concentrations in the atmosphere above the North Pacific were 0.32.7 ng/m3 but its concentration in the outdoor air of Denmark was 22 ng/m3.13 A few ng/ m3 (medians between 0.37 and 2.0 ng/m3) were reported for DBP and DEHP in the atmosphere of Sweden and the United States.41 Outdoor air concentrations of phthalates in Japan were from <23 to 121 ng/m3 for TBP42 and 290 ng/m3 in average in Yamaguchi in 4-day measurements, respectively.52 Umemoto et al.52 also reported concentrations for DEP and DEHP in outdoor air, i.e., 26 and 29 ng/m3. Quantitative analyses of organophosphates in outdoor air samples showed concentrations at the low nanogram per cubic meter level for TBP, TCP, TCEP TCPP, TDCPP, TEP, and TPP. TBP in air collected near paper manufacturing plants in Japan was 13.4 ng/m3.27 TCP reached concentrations up to 70 ng/m3 in Japan, but at a maximum of only 2 ng/m3 in the production site in the USA.25 For TCPP and TDCP concentrations of 5.3 and 4.7 ng/ m3 were measured in the ambient air of Kitakyushu, Japan,11 and the maximum level of TPP was 23.2 ng/m3.26 Less than 1 ng/m3 quantities were measured for each of the individual concentrations of TBEP, TBP, TCEP, TCPP, TPP, TEHP in Sweden.39 Highest concentrations of contaminants including plasticizers in air were reported for test chambers built to identify and quantify emissions from products used indoors. In test chamber studies evaluating the emission of phthalates from PVC coated wall coverings maximum concentrations of 5.1 µg/m3 (= 5,100 ng/m3) for DBP, of 0.50 µg/m3 for DiBP and of 0.94 µg/m3 for DEHP were observed in a 14 day test period (Uhde et al.53). Organophosphates (TBP, TBEP, TCEP, TCPP, TDCPP, TEHP, TPP) used for car interior components and evaluated in emission test chambers led to the highest concentrations for TPP of about 2 µg/m3 (= 2,000 ng/m3).33

17.5.3.2 House Dust House dust is a “long-term accumulative sample” trapping, accumulating and preserving contaminants.54 It has been regarded as an “indoor-pollution-archive”.55 It serves as a sink and reservoir for semi-volatile and non-volatile substances, thus collecting them like a passive sampler. Biodegradation, that occurs readily under aerobic and moist conditions (e.g.: t1/2 = 2 to 4 weeks for DEHP)13 does not occur in the indoor environment13 because temperature extremes, sunlight, and moisture are absent. Thus dust may provide a useful record of post indoor chemical use and exposure.56 Accumulation of plasticizers will not

728

Health and Safety Issues with Plasticizers and Plasticized Materials

only occur in house dust as a large sorbent reservoir, but plasticizers are further present in or adsorbed to other sorbant surfaces in the indoor environment. Due to its varying sources, house dust is very heterogeneous and consists of a variety of inorganic and organic particles as well as fibers of different sizes.57,58 Average particle size distribution data showed about 65% of particles to be >300 µm, 30% to be 75-300 µm, and 5% to be <75 µm.58 Microscopic analysis revealed that house dust contains organic particles, such as, cellulose fibers (cotton, paper, wood), animal fibers (hair, wool), gums, resins, fats, oils, and rubber particles as well as inorganic materials including soil minerals (quartz, feldspar), and particles of soil or building materials (gypsum, limestone, and dolomite).58 Quantity and composition of house dust varies greatly with seasonal and environmental factors. According to Butte and Walker,55 the portion of organic matter in house dust samples may be between <5% and >95%. Fergusson and Kim59 reported the organic content of house dust to vary from 25.7% to 56.5%, and floor dust from Danish offices had a mean organic fraction of 33%.60 Even though there are standard protocols for sampling of house dust61,62 a great variety in sampling techniques has been noted. These were reviewed by Butte and Heinzow,57 Lioy et al.,5 Macher,63 and Millette and Few.64 As deposited dust (mainly floor dust) is most often collected by vacuuming, “house dust” is often regarded to be the content of vacuum cleaner bags. After sampling the crude dust, the vacuum cleaner bag has to be processed to obtain a sub-sample appropriate for analysis. Obtaining a more homogeneous sub-samples may be achieved by discarding larger particles, such as hair and feathers, atypical objects, and coarse material (e.g., clips, small toys, knobs, small stones, etc.) from the gross dust, or by choosing just “pieces of fluff” (dust bunnies) for analysis, as well as by sieving the dust.57 On sieving, coarse material is removed and a separation of particles from fibers is performed. It has to be noted that concentrations of contaminants, including plasticizers in house dust may greatly vary with the sampling and the sample preparation technique. Detection limits for phthalates (e.g., BBP and DBP) in floor dust may reach 0.001 mg/kg,38 and about 0.1 mg/kg-0.5 mg/kg for organophosphates.22,32,65 First, but very few, results on the occurrence of phthalates indoors were collected by Wams.13 The first report on the occurrence of a flame retardant plasticizer in house dust was on TDCPP. Sellström and Jansson66 mentioned it to be present in one out of two Swedish dust samples (from vacuum cleaner bags) but gave no concentration. In the last

729

17.5 Plasticizers in the indoor environment

years however, quite a few results on phthalates and organophosphate plasticizers have been published. Some of them are compiled in Tables 17.5.5 and 17.5.6. Table 17.5.5. Concentrations of phthalates in house dust in mg/kg = µg/g Compound

Median

95. Percentile

Range

N

Reference

14.7

207

?-745

199

Becker et al.67 (1)

30.5

320

0.3-1,400

286

Butte et al.68 (2)

19

230

?-700

65

Kersten & Reich34 (3)

24

270

<0.7-510

272

Pöhner et al.69 (4)

117 (mean)

-

12.1-524

6

Rudel et al.35 (5)

67.7 (mean)

-

15.1-175

10

Wilson et al.38 (6)

BMEP

2

8

1-17

65

Kersten & Reich34 (3)

DBP

41.5

160

?-502

199

Becker et al.67 (1)

49

240

3.5-500

286

Butte et al.68 (2)

47

180

?-600

65

Kersten & Reich34 (3)

87

370

<0.7-1,200

272

Pöhner et al.69 (4)

27.4 (mean)

-

11.1-59.4

6

Rudel et al.35 (5)

18.4 (mean)

?

1.58-46.3

10

Wilson et al.38 (6)

22.4

130

?-192

199

Becker et al.67 (1)

34

130

1.1-330

286

Butte et al.68 (2)

33

78

?-470

65

Kersten & Reich34 (3)

1.32 (mean)

-

1.05-2.05

6

Rudel et al.35 (5)

1

5

1-80

62

Kersten & Reich34 (3)

1.86 (mean)

-

0.569-5.38

6

Rudel et al.35 (5)

DiDP

31

340

1-4,200

62

Kersten & Reich34 (3)

DEHP

416

1,190

?-7,530

199

Becker et al.67 (1)

735

2,600

62-12,000

286

Butte et al.68 (2)

600

1,600

?-2,700

65

Kersten & Reich34 (3)

450

2,000

<0.7-8,600

272

Pöhner et al.69 (4)

315 (mean)

-

69.4-524

6

Rudel et al.35 (5)

3.3

89.7

<0.1-1,233

199

Becker et al.67 (1)

5

350

1-570

65

Kersten & Reich34 (3)

3.1

96

<0.5-310

272

Pöhner et al.69 (4)

2.15 (mean)

-

1.01-3.58

6

Rudel et al.35 (5)

0.2

3.7

<0.1-75.8

199

Becker et al.67 (1)

1

20

1-64

65

Kersten & Reich34 (3)

0.6

4.1

<0.5-35

272

Pöhner et al.69 (4)

DiNP

72

540

1-1,000

61

Kersten & Reich34 (3)

DOP

4

73

1-160

65

Kersten & Reich34 (3)

BBP

DiBP

DCHP

DEP

DMP

730

(1) (2) (3) (4) (5) (6)

Health and Safety Issues with Plasticizers and Plasticized Materials

House dust from the German Environmental Survey, representative for Germany regarding age, gender, community size and place of residence. A subset of 200 randomly selected vacuum cleaner bags was taken, dust was sieved to <2 mm, House dust from vacuum cleaner bags collected in Northern Germany (1998/1999), sieved to <63 µm, House dust from vacuum cleaner bags collected in Hamburg (Germany, 1998-2000), sieved to <63 µm, Dust from the space between the two filters of double-layer vacuum cleaner bags, Germany (43% of this house dust was <63 µm, only 8% >160 µm), Dust samples taken from the surface of rugs, upholstery, wood floors, window sills, ceiling fans, and furniture from one office and 5 homes in Massachusetts (USA), sieved to <150 µm, Floor dust from child care centers in North Carolina (USA, 1997), collected with the high volume sampler HVS3 using ASTM standard method D5438-94 (ASTM 1997), sieved to <150 µm.

Table 17.5.6. Organophosphates in house dust in mg/kg = µg/g

Compound

Median

95. Percentile

Range

N

Reference

DPEHP

0.8

370

?-990

29

Nagorka & Ullrich65 (1)

TBEP

5.8

58

<0.1-854

199

Becker et al.67 (2)

107

-

0.7-1,310

11

Hansen et al.24 (3)

5.0

40

?-120

65

Kersten & Reich34 (4)

16.1

162

?-210

29

Nagorka & Ullrich65 (1)

0.4

1.5

?-5.7

65

Kersten & Reich34 (4)

2.5

34.4

?-49.3

29

Nagorka & Ullrich65 (1)

<0.1

1.0

<0.1-6.0

199

Becker et al.67 (2)

470

-

2.5-2,190

12

Hansen et al.24 (3)

TBP TCEP

6.9-16

TCP TCPP

TDCPP

TEHP

Langer et al.113

230-1800 (max)

0.6

8.4

?-330

1,569

Haumann & Thumulla32 (5)

0.66

9.4

<0.1-121

983

Ingerowski et al.22 (6)

1.6

6.2

?-9.5

65

Kersten & Reich34 (4)

2.5

6.33

?-6.84

29

Nagorka & Ullrich65 (1)

0.9

8.4

<0.1-94

59

Sagunski et al.36 (7)

< 0.1

0.4

<0.1-80.7

199

Becker et al.67 (2)

2.2

15

0.1-36

65

Kersten & Reich34 (4)

1.0

14

?-470

1,337

Haumann & Thumulla32 (5)

0.57

5.9

<0.1-375

436

Ingerowski et al.22 (6)

1.4

12

?-27

63

Kersten & Reich34 (4)

<0.5

2.3

?-120

503

Haumann & Thumulla32 (5)

1.2

6.8

0.1-35

62

Kersten & Reich34 (4)

1.69

12.4

?-18.3

29

Nagorka & Ullrich65 (1)

<0.1

1.6

<0.1-4.6

199

Becker et al.67 (2)

0.2

0.9

0.1-2.0

62

Kersten & Reich34 (4)

0.80

22.4

?-37.5

29

Nagorka & Ullrich65 (1)

731

17.5 Plasticizers in the indoor environment

Table 17.5.6. Organophosphates in house dust in mg/kg = µg/g

Compound TPP

(1) (2) (3) (4) (5) (6) (7)

Median

95. Percentile

Range

N

Reference

0.3

1.8

<0.1-7.2

199

Becker et al.67 (2)

1

-

<1-220

12

Hansen et al.24 (3)

2.9

16

?-56

65

Kersten & Reich34 (4)

6.51

19.5

?-22.9

29

Nagorka & Ullrich65 (1)

Passively deposited house dust (sedimentation time 3 months to 1 year) from German residences, House dust from the German Environmental Survey, representative for Germany regarding age, gender, community size and place of residence. A subset of 200 randomly selected vacuum cleaner bags was taken, dust was sieved to <2 mm, House dust from public buildings (Germany) with buildings parts supposed to contain organophosphates flame retarding plasticizers, House dust from vacuum cleaner bags collected in Hamburg (Germany, 1998-2000), sieved to <63 µm; House dust, exactly one week old, sampled with commercial vacuum cleaners (Germany); only the “fine dust fraction” (not specified exactly) was analyzed, House dust, exactly one week old, sampled with commercial vacuum cleaners (Germany), analysis of total dust, values given are the average of 3 laboratories participating in an inter-laboratory exposure study, House dust sampled with commercial vacuum cleaners (Germany), analysis of total dust.

Phthalates are present in house dust with medians of a few mg/kg (e.g., DMP or DEP) up to some hundred mg/kg (DEHP). Regarding the 95. percentiles, for example DEHP, phthalates may also be present in house dust in the gram per kilogram range. In general, the distribution of pollutants in the environment is not Gaussian but “lognormal” That is, when the logarithms of the observed concentrations are plotted as a frequency distribution, the resulting distribution is normal or Gaussian. A physical explanation of the lognormality of pollutant concentrations in environmental samples was given by Ott.70 Lognormal distributions have also been described for the phthalates (e.g., BBP, DBP, DiBP, and DEHP) in house dust.68 Results for other phthalates, such as, dipropyl, diallyl and diphenyl phthalates were given by Kersten and Reich.34 These compounds are present in less than 10% of house dust samples with concentrations not exceeding 5 mg/kg (in the <63 µm fraction of the dust). Concentrations of organophosphates tend to be much smaller compared to phthalates. Medians do seldom exceed a few milligrams per kilogram, even the 95. percentiles are seldom higher than 10 mg/kg. For the semi-volatiles (e.g., TCEP), a strong correlation was reported for concentrations in the indoor air and in the house dust on the basis of 9 objects showing concentrations of TCEP in indoor air between 0.02 and 3.0 µg/m3, in house dust between 2.5 and 2,190 mg/kg, respectively.24 Hostrup and Butte,46 analyzing BBP, DBP, DiBP, and DEHP in house dust and air of dwellings known to show high concentrations of these phthalates in house dust, observed a weak but significant correlation for DBP only, but not for BBP, DiBP or DEHP. Roinestadt et al.71 analyzing 23 biocides in indoor air and dust, reported that pesticides in air were always found in the corresponding dust with the exception of dichlorvos, o-phenylphenol, and chlordane. But the majority of household pesticides is preferably detected in the home environment by dust sampling. The data suggest that semi-volatile compounds might be sampled in the air, although they are adsorbed to house

732

Health and Safety Issues with Plasticizers and Plasticized Materials

dust as well. For non-volatile, i.e. particle-bound compounds, however, house dust is the material of choice to indicate an indoor contamination. Discontinued use or the replacement of one plasticizer by another may result in downward or upward trends. As the production of TCEP in Germany has ceased since 19979 a downward trend should to be expected. Further results for contaminants in house dust including other compounds are compiled in the reviews of Butte and Heinzow,57 Lioy et al.,5 and Santillo et al.72

17.5.4 IMPACT OF PLASTICIZERS IN THE INDOOR ENVIRONMENT 17.5.4.1 Indoor Plasticizers and Health Indoor contamination is one source of exposure to toxic pollutants, which has been classified as a high environmental risk.9,73 Regarding the principal exposure routes, such as, inhalation, dietary ingestion, dermal, and non-dietary ingestion, plasticizers in indoor air and bound to suspended particles may contribute to exposure via inhalation, house dust to dermal penetration and non-dietary ingestion, respectively. There is evidence that inhalation exposure to DEHP, the major plasticizer indoors, may increase the risk of inducing inflammation in the airways, which is a characteristic of asthma.50 Adverse health effects have also been discussed in connection with an indoor occurrence of organophosphates like TCEP.23,36 Small children are considered to be the population at highest risk since they spend most of their time indoors and much of this time is spent in contact with floors, engaging in mouthing of hands, toys and other objects. Regarding pesticides for example, recent findings of indoor exposure indicate that young children are at higher risks than had been previously estimated,74 but improvements in measurements of household exposures relating to health effects are still needed.75 Acute toxicity of all phthalates is low. Subacute and chronic effects of phthalates were summarized by Lorz et al.15 They depend on chain length and structure of the alcohol moiety. Thus, a differentiation between short-, medium- and long-chain esters seems necessary. After entering the human organism, one of the two ester bonds of phthalates is cleaved and the alcohol is released. Toxicological effects may result from the phthalic acid monoester and its sequel products and/or from the alcohol. Short-chain phthalates (DMP, DEP) revealed no significant toxicological effects, although dermal exposure to DMP in humans occurs widely due to its use in insect repellants. DMP and DEP used in cosmetics at concentrations of <10% may be dermally resorbed. They are considered to be non-irritants, non-sensitizers, and non-phototoxic agents. Developmental toxicity is regarded as insignificant. Medium-chain phthalate esters (DBP, DiBP, DCHP) exhibit degenerative testicular effects upon repeated oral administration in all species investigated. Furthermore, in rodents, but not in other species, hepatic alteration (peroxisome proliferation and hepatomegaly) is observed. The testicular effects are also obtained with DEHP to a similar degree (see below) but are not observed with the linear longer-chain like DOP or with short-chain phthalates. 1% DBP in the diet of mice caused embryotoxicity and teratogenicity. BBP at 2.5% and 5.0% in the feed also caused testicular effects, enlargement of liver and kidneys and decrease of bone marrow cellularity and thyroid function. DEHP is the long-chain phthalate ester examined most extensively, and hundreds of publications on toxicological profile, biochemistry, and metabolism exist. Far fewer data have been generated on other phthalates of this group. Branched-chain phthalates com-

17.5 Plasticizers in the indoor environment

733

monly share a potential to induce peroxisomes and hepatomegaly in livers of rats and mice at high doses. In long-term feeding experiments, DEHP and related phthalates also led to an increase of liver tumors in rodents. DEHP also caused embryotoxic and teratogenic effects in mice, DOP did neither exert testicular toxicity in vivo nor significant peroxisome proliferation in the liver. It was shown to be a liver tumor promoter in a rat feeding experiment. Reviews summarizing toxic effects of phthalates, i.e., BBP, DBP, DiDP, DEHP, DiNP, DOP were given in a series of publications by Kavlock et al.76-81 Toxicities of organophosphate plasticizers are compiled in the hazardous substances data bank of the National Library of Medicine's Toxicological Data Network.83 All of the organophosphates mentioned in this chapter show a low acute toxicity for mammals, but seem to have weak serum cholinesterase inhibiting properties. Some of them are skin or eye irritants (e.g., TEHP, TCEP) or sensitizers (e.g., DPEHP). Several subchronic studies with TBEP in laboratory animals have shown that the liver is the target organ. The long term toxicity and carcinogenicity of TBEP have not been studied. Bacterial and mammalian cell tests for gene mutation gave negative results. Teratogenicity was not observed, other aspects of reproductive toxicity have not been reported. TBP may cause irritation of the eyes, nose, and throat. It may also cause nausea and headache. From in vitro test results, it is not considered to be mutagenic. TCEP adversely affects the fertility of male rats and mice, but is not teratogenic. In vitro mutagenicity test results were inconsistent. In studies on rats and mice, TCEP showed neurotoxic properties.82 TEHP gave negative results in several in vivo and in vitro tests for mutagenicity, but there was some evidence of carcinogenicity based on an increased incidence of hepatocellular carcinomas in female mice and on the increased incidence of adrenal pheochromocytomas in male rats. Considering the low incidence of this tumor and its occurrence in only one sex of one species, the lack of evidence of genetic toxicity, and the low exposure of humans to TEHP, it is unlikely that TEHP poses a significant carcinogenic risk to humans.83 Besides the toxic effects, phthalates are compounds that have endocrine disrupting properties, i.e., they belong to the so-called xenoestrogens.84 Although the estrogenic potency of phthalates is rather small compared to the reference compound 17ß-estradiol, their endocrine disrupting properties may nevertheless be a problem as production rates are high. Regarding reproductive and developmental effects, phthalates vary in potency with DEHP being the most potent. DBP and BBP are roughly one order of magnitude less potent.17 An Expert Panel convened by the National Toxicology Program, NTP, Center for the Evaluation of Risks to Human Reproduction, CEHR, has announced that after intensive evaluation of seven phthalates (including BBP, DBP, DEHP, DiNP, and DOP) only one presents a serious concern regarding human reproduction or development: DEHP.85 Chemicals that give rise to toxic endpoints other than cancer and gene mutations are often referred to as “systemic toxicants” because of their effects on the function of various organ systems.86 Based on the understanding of homeostatic and adaptive mechanisms, systemic toxicity was treated by the US-EPA as if there is an identifiable exposure threshold (both for the individuals and for populations) below which there are no observable

734

Health and Safety Issues with Plasticizers and Plasticized Materials

adverse effects, i.e., the (chronic) reference dose, RfD. Principles of determining and using RfD values were summarized in a background document.86 Based on the NOAEL (no observable adverse effect level) for animal experiments TDI values (tolerable daily intake) were calculated for some phthalates by the CSTEE.87 RfD and TDI values for phthalates are summarized in Table 17.5.7. Neither RfD nor TDI values are available for organophosphates. Table 17.5.7. TDI (tolerable daily intake) and RfD (reference dose) of some phthalates in µg/kg bw/d* TDI = Tolerable daily intake87

RfD = Reference dose88

200

200

DBP

100

100

DiDP

250

-

DEHP

Compound BBP

37

20

DEP

-

800

DiNP

150

-

DOP

370

-

* µg/kg bw/d = microgram per kilogram body weight per day

Regarding carcinogenicity, only BBP and DEHP were classified by the International Agency for Research on Cancer (IARC). There was no evidence for carcinogenicity to humans, both phthalates were reported to be “not classifiable as to its carcinogenicity to humans (Group 3)”.89,90 Even though there was “sufficient evidence” in experimental animals for the carcinogenicity of DEHP,90 the mechanism by which DEHP increased the incidence of hepatocellular tumors in rats and mice were regarded as not relevant to humans. A similar classification of DEHP regarding its carcinogenicity was given Doull et al.91 They classified DEHP as clearly non genotoxic. Regarding carcinogenic effects of organophosphates only TCEP has been classified by the IARC. There is limited evidence for its carcinogenicity in experimental animals. But it was assessed as “not classifiable as to its carcinogenicity to humans (Group 3)”.92

17.5.4.2 Human Exposure Assessment for Plasticizers in the Indoor Environment The input of plasticizers into the indoor environment may either result from a direct application (e.g,. phthalates in sprays for fighting insects, in repellents or hair sprays), from plasticizers used in varnishes, colors, adhesives, etc., or from pieces of furniture, electronic goods, floor coverings and surface coated wallpaper produced with flame retarding organophosphates and/or phthalates. Gaseous plasticizers are evenly dispersed in the air. In the case of inhalation, the anatomy and physiology of the respiratory system diminishes the pesticide concentration in inspired air. As phthalates and organophosphates are lipid soluble, they are usually not removed in the upper airways but tend to deposit in the distal portion of the lung, the alveoli and may then be absorbed into the blood stream.93 Particulate or particle-bound plasticizers are either dispersed in the air or deposited/adsorbed by house dust. They may enter

17.5 Plasticizers in the indoor environment

735

the human body either by inhalation of suspended particulate matter or through oral intake, i.e. non-dietary ingestion of dust (infants, toddlers) and ingestion of particles adhering to food, to surfaces in the homes (e.g., toys), and to the skin. Plasticizers in surfaces or adsorbed to surfaces may further be passed directly through the skin by dermal contact. Dust suspended in air and thereby inhaled is deposited in different parts of the alveolar tract (nose, throat and lung) dependent upon its size (aerodynamic diameter). The efficacy of deposition in the alveolar tract increases with decreasing particle size. Particles smaller than 10 µm enter in the tracheo-bronchial area, reaching, depending on their size, trachea, bronchi or alveoli. Substances adsorbed onto the dust particles that enter the alveoli can be absorbed by epithelial cells of the lung or through macrophagial phagocytosis.57 But exposure to house dust does not exclusively and may not even occur predominantly via inhalation. For instance, ingestion of house dust particles adhering to food objects and the skin or direct absorption through the skin may be primary routes of exposure.94 This holds true especially for small children as they have the tendency to be very tactile and handle and place non-food objects into their mouth.95 Upon taking PVC toys or other items into the mouth, plasticizers may be leached out. The release of phthalate plasticizers from soft PVC toys and child care articles was determined by a working group of the TNO (Netherlands). PVC toys containing 24% of DEHP released a mean of 1.67 µg of DEHP per minute and 10 cm2.96 Oral uptake is greatly dependent on the frequency of mouthing and the duration the items are held. Regarding mouthing behavior, the daily frequency of both mouth and tongue contacts with hands, other body parts, surfaces, natural objects, and toys was evaluated by Tulve et al.97 for children from 11 to 60 months of age. A clear relationship was observed between mouthing frequency and age. Children less than 24 months of age exhibited a frequency of mouthing with an average of 81 events per hour, children more than 24 months of age, 42 events per hour, respectively. Estimations of the quantity of dust ingested daily vary greatly. Infants and toddlers, who crawl and mouth their hands and other objects, ingest a daily rate twice that of adults that is estimated to be 0.02-0.2 g.98 Toddlers, possibly eating non-food items, may consume as much as 10 g of soil or dust per day (“pica-behavior”).98 US-EPA assumptions on the indoor exposure among young children (i.e., 2.5 years old) were based on an ingestion of 100 mg of house dust per day during the winter months and of 50 mg during the warmer months.93 Oral intake of phthalates and organophosphates, inhalative intake of phthalates, respectively, are complied in Table 17.5.8. In the case of DEHP some micrograms per kilogram body weight and day may be ingested with house dust. But up to 21 µg/kg bw/d may be absorbed, if the 95. percentile of Butte et al.68 for the DEHP in house dust is reached. This quantity exceeds the RfD of the US-EPA.88

736

Health and Safety Issues with Plasticizers and Plasticized Materials

Table 17.5.8. Daily indoor intake of some plasticizers (phthalates/organophosphates) calculated on the basis of means or medians published for house dust (oral intake) and air (inhalative intake) Compound

Oral intake, µg/kg bw/d1,2

Inhalative intake, µg/kg bw/d1,3

Phthalates BBP

0.12-0.95

0.040-0.055

DBP

0.15-0.71

0.13-0.46

DEHP

2.7-6.0

0.034-0.14

DEP

0.017-0.041

0.44

DiNP

0.58

-

DOP

0.032

-

Organophosphates TBEP

0.04-0.87

-

TBP

0.0033-0.020

-

TCEP

<0.0001-3.8

-

TCP

<0.0001-0.018

-

TCPP

0.0046-0.011

-

TDCPP

<0.004-0.014

-

TEHP

<0.0001-0.0065

-

TPP

0.0024-0.053

-

1

µg/kg bw/d = micrograms per kilogram body weight per day, Assuming an oral intake of 100 mg of house dust per day and a weight of 12.3 kg for a toddler (age 1-2);93,94 concentrations in house dust are compiled in Tables 17.5.5 and 17.5.6, 3 Assuming an inhalation rate of 6.8 m3 per day, and a weight of 12.3 kg for a toddler (age 1-2);93 concentrations in indoor air are compiled in Table 17.5.4. 2

Kohn et al.99 reported that the “Center for the Evaluation of Risks to Human Reproduction” (USA) estimated a daily intake for DEHP of 3 to 30 µg/kg bw/d. Wilson et al.38 ranked the exposure pathways for phthalates (based on data for DBP and BBP) as follows: dietary ingestion > non-dietary ingestion > inhalation, but dust ingestion scenarios show that child exposures may exceed the reference dose (RfD) via non-dietary ingestion of contaminated house dust in case of DEHP. Thus the indoor environment seems to contribute significantly to human intake of plasticizers especially for DEHP. On the other hand, inhalative intake of phthalates and organophosphates as well as oral intake of organophosphate plasticizers are much smaller and seem to be negligible. An excellent review on dust as a metric for use in residential and building exposure assessment and source characterization was recently published by Lioy et al.5 The quantity of suspended dust inhaled and the intake of plasticizers adsorbed to it may be calculated rather exactly, and the intake from deposited dust via oral pathways may at least be roughly estimated. But data supporting the amount of chemicals absorbed

17.5 Plasticizers in the indoor environment

737

through dermal contact of contaminated house dust and through direct contact to contaminated surfaces are still lacking. Regarding concentrations of plasticizers in material of human origin (blood, serum, urine, body fat, etc.) there are several data for DEHP and some data for the other phthalates, but values for organophosphate plasticizers are rare. A review on toxicological aspects and concentrations of DEHP in material of human origin after using it in medical devices has recently been published by the US-FDA.100 Concentrations of metabolites, i.e., the monoesters, of BBP, DBP, DCHP, DEHP, DiNP and DOP in urine for the American population were measured by Blount et al.17 for BBP, DBP, DEP, DEHP and DOP in urine and for the German population by Koch et al.,101 respectively. Both studies showed monoethyl phthalate, the metabolite of DEP, to be the phthalate with the highest concentration. DEP was concluded to be the phthalate with the highest intake for the American population,17 DEHP for the German population, respectively.101 Phthalates as well as organophosphates seem to be accumulated in the human body as they have been detected in human body fat. Butyl benzyl phthalate amounted to 0.19 µg/g (= mg/kg) in human adipose tissue in the US.83 In a series of studies trialkyl-, trihaloalkyl-, and triaryl phosphates were analyzed in human adipose tissue.102-104 Although there was no history of exposure to organophosphates, TDCPP and TBEP were detected at levels of 0.5-110 µg/kg and 4.026.8 µg/kg, respectively. TCEP, however could not be detected.102 TBEP and TDCPP were also the only organophosphates detected in adipose tissue obtained from Scheme 17.5.1. Derivation of a guideline value for tri- human cadavers from 6 Ontario municipali(2-chloroethyl) phosphate, TCEP, in indoor air [After ties in the Great Lakes Basin area. TBEP levSagunski and Roßkamp, 2002]. LOAEL: lowest els ranged from 25 to 483 µg/kg; TDCPP observable adverse effect level; mg/kg bw/d: millifrom 1.5 to 22 µg/kg, respectively.104 grams per kilogram body weight per day.

738

Health and Safety Issues with Plasticizers and Plasticized Materials

17.5.4.3 Reference and Guideline Values of Plasticizers to Assess Indoor Quality For the interpretation of results obtained from monitoring plasticizers in the indoor environment in general two approaches might be used: 1. the comparison with reference values 2. the application of a risk assessment methodology related to the hazard of the compound leading to standards or guideline values. The former approach will help determine whether a measured value is “normal”, but has no health meaning or regulatory implication. Reference values are intended to characterize the upper margin of a current background contamination of a pollutant at any given time. Regarding plasticizers in the indoor environment reference values may be calculated applying the procedure as for environmental toxins in body fluids.105 This concept defines the upper margin (95th percentile) of the concentration levels as the reference value. To be valid for the general population, reference values must be derived from studies that are large enough to be representative for a studied population. Reference values may show trends, as contaminant levels may change with time. Of all the studies mentioned in this chapter only the work of Becker et al.67 and Butte et al.,68 both reporting concentrations of plasticizers in the house dust, may be taken to calculate the reference values, as these studies rely on data representative for their studied population. Other studies are valuable in getting an impression of “normal” concentrations, especially if they rely on many measurements, such as the work of Haumann and Thumalla,32 Ingerowski et al.,22 Kersten and Reich,34 and Pöhner et al.69 It must be emphasized that the reference values have no uniform character and relate to the reference population only. They are subject to regional/spatial and temporal variation and, in the case of the house dust, they are significantly governed by sampling technique and sample preparation. The reference values for plasticizers in indoor air are not available. Applying a risk assessment methodology related to the hazard of a pesticide leads to the standard or guideline values. For the levels or concentrations of plasticizers in the residential indoor environment, there are no standards, i.e., values which might be legally enforced. Guideline values, however, have been defined for TCEP in indoor air.23 They are based on the toxicity data on TCEP. The methods of defining guideline values for TCEP are shown in Scheme 17.5.1. The official German concept of guideline values for indoor air involves two levels. Guideline value II is a health-related value based on the current toxicological and epidemiological knowledge. If a concentration of a pollutant indoors reaches or exceeds this value, immediate action has to be taken to avoid health impairments. Guideline level I is the level at which a substance does not give rise to adverse health effects even under lifelong exposure. It may further be regarded as the level to be reached after a sanitation of rooms (diminishing the indoor contamination after exceeding guideline values II). A review on guideline values for indoor air reflecting various countries was given by Pluschke.106 because of toxicological concerns and restrictions of different dialkyl ortho-phthalates, other plasticizers have been increasingly used in recent years in Germany.111 Diisononyl cyclohexane-1,2-dicarboxylate (DINCH), di(2-ethylhexyl) terephthalate (DEHT), di(2-ethylhexyl) adipate (DEHA), acetyl tri-n-butyl citrate (ATBC), and trioctyl

References

739

trimellitate (TOTM) plasticizer levels in indoor air and dust samples from 63 day-care centers in Germany were measured.111 The urine samples of 208 children who attend 27 of these facilities were analyzed for the presence of four DINCH metabolites.111 DINCH, DEHT, and DEHA were present in indoor air with median values of 108 ng/ m3, 20 ng/m3, and 34 ng/m3, respectively.111 Median values of 302 mg/kg for DINCH, 49 mg/kg for DEHA, 40 mg/kg for DEHT, and 24 mg/kg ATBC were found in dust. In the urine samples, the three secondary metabolites of DINCH were observed with median values (95th percentiles) of 1.7 µg/l (10.0 µg/l) for OH-MINCH, 1.5 µg/l (8.0 µg/l) for oxoMINCH, and 1.1 µg/l (6.1 µg/l) for cx-MINCH.111 Overall, these metabolite levels are orders of magnitude lower than the current HBM I values set by the German Human Biomonitoring Commission.111 Compared with tolerable daily intake values, the total daily intake of DINCH reached only 1% of its maximum value to date; however, due to its increased use, higher exposure of DINCH is expected in the future.111

17.5.5 SUMMARY The use of indoor air and house dust to identify an indoor contamination with plasticizers, i.e., organophosphates and phthalates, and for characterizing a potential exposure are discussed in this chapter. Organophosphates and phthalates are either semi-volatile organic compounds, SVOC, or particulate organic matter, POM. If they are SVOC, a correlation of concentrations in indoor air and house dust might be observed. To obtain information on an indoor contamination, the consecutive measurements of indoor air or of house dust should be performed. Indoor air study provides information on an actual indoor contamination and the exposure via inhalation. House dust, that serves as a reservoir for plasticizers, may be regarded as a pollution-archive as it is trapping, accumulating, and preserving plasticizers, which were applied or brought in by treated items. Ingestion and dermal contact with house dust as well as with objects containing plasticizers may be the primary routes of exposure especially for toddlers and infants. Regarding concentrations of plasticizers in indoor air and house dust they may be assessed in comparison to reference values, as guideline values are only available for TCEP in indoor air (in Germany). The indoor air plasticizer concentrations, detected in Japanese study, were lower than concentrations typically associated with sick-building syndrome.112 At the same time, it is difficult to establish the concentration limits for indoor air chemicals with regard to human health because of a lack of knowledge regarding the pathogenesis of sick-building syndrome and/or chemical sensitivity.112

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Addendum

ADDENDUM LIST OF ABBREVIATIONS Abbreviation

Name

CAS No.

Phthalates BBP BMEP DBP DiBP DCHP DiDP DEP DEHP DMP DiNP DOP

benzylbutyl phthalate bis(2-methoxyethyl) phthalate dibutyl phthalate diisobutyl phthalate dicyclohexyl phthalate di(isodecyl) phthalate diethyl phthalate di-(2-ethylhexyl) phthalate dimethyl phthalate di(isononyl) phthalate di-n-octyl phthalate

85-68-7 117-82-8 84-74-2 84-69-5 84-61-7 26761-40-0 84-66-2 117-81-7 131-11-3 28553-12-0 117-84-0

diphenyl 2-ethylhexyl phosphate tri-(2-butoxyethyl) phosphate tributyl phosphate tri-(2-chloroethyl) phosphate tricresyl phosphate (mixture of isomers) trichloroisopropyl phosphate tridichloroisopropyl phosphate = tri-(1,3-dichloro-2-propyl) phosphate tri-(2-ethylhexyl) phosphate = trioctyl phosphate triethyl phosphate triphenyl phosphate

1241-94-7 78-51-3 126-73-8 115-96-8 1330-78-5 13674-84-5

Organophosphates DPEHP TBEP TBP TCEP TCP TCPP TDCPP

TEHP TEP TPP

13674-87-8 78-42-2 78-40-0 115-86-6