Phthalate esters in clothing: A review

Environmental Toxicology and Pharmacology 108 (2024) 104457 Available online 25 April 2024 1382-6689/© 2024 The Author(s). Published by Elsevier B.V. This is an open access article under the CC BY-NC license (http://creativecommons.org/licenses/bync/4.0/). Phthalate esters in clothing: A review Natalia Aldegunde-Louzao a , Manuel Lolo-Aira b , Carlos Herrero-Latorre a , * a Research Institute on Chemical and Biological Analysis, Analytical Chemistry Nutrition and Bromatology Department, Faculty of Sciences, Universidade de Santiago de Compostela, Campus Terra, Lugo 27002, Spain b Applied Mass Spectrometry Laboratory (AMSlab), Avda. Benigno Rivera, 56, Lugo 27003, Spain ARTICLE INFO Edited by Ayse Basak Engin Keywords: Phthalates in clothing Review Legislation Analytical methods Exposure Health risks ABSTRACT Phthalate esters (PAEs) are widely used as plasticizers to enhance the flexibility and durability of different consumer products, including clothing. However, concerns have been raised about the potential adverse health effects associated with the presence of phthalates in textiles, such as endocrine disruption, reproductive toxicity and potential carcinogenicity. Based on examination of more than 120 published articles, this paper presents a comprehensive review of studies concerning the phthalate content in clothing and other textile products, with special emphasis on those conducted in the last decade (2014–2023). The types and role of PAEs as plasticizers, the relevant legislation in different countries (emphasizing the importance of monitoring PAE levels in clothing to protect consumer health) and the analytical methods used for PAE determination are critically evaluated. The review also discusses the models used to evaluate exposure to PAEs and the associated health risks. Finally, the study limitations and challenges related to determining the phthalate contents of textile products are considered. 1. Introduction Phthalates or phthalate esters are the ortho-di-esters of phthalic acid (PAE) with the general structure illustrated in Fig. 1A. Phthalates are used as plasticizers in the industrial fabrication of soft and flexible synthetic plastic polymers, especially polyvinyl chloride (PVC) (Tang et al., 2020). These compounds are also used in the preparation of adhesives, as well as in the manufacture of medical devices, personal care items, textile fabrics and other related textile plasticized products (Martínez et al., 2018). Consequently, PAEs appear in many plastic-based consumer and industrial products and, as they are not chemically bound to plastic polymers, they are widely released and disseminated in the environment (Gong et al., 2016). Human exposure to phthalates can take place through different routes, including ingestion, inhalation and dermal absorption. The main sources of PAE uptake are food, personal care products, air, home furnishings, nutritional supplements, pharmaceuticals, medical devices and phthalate-containing surfaces or clothes in contact with skin (Chang et al., 2021). These types of phthalate exposure have been related to adverse effects on human health. Specifically, PAEs are considered endocrine disruptors (Wang and Qian, 2021), and some of them are classified as possible carcinogens by the US Environmental Protection Agency (US-EPA, 2007). Other studies have observed a relationship between exposure to PAEs and reproductive toxicity, with premature delivery in pregnant women and ovarian failure amongst the effects reported (Radke et al., 2018; Zhao et al., 2022). Negative effects on insulin and an increased risk of diabetes have also been reported (Dirinck et al., 2015; Radke et al., 2019). In the light of these effects, many countries have established severe restrictions through legislation aimed at controlling the use of phthalates. Different research studies have investigated the levels of these contaminants in different foods (Giuliani et al., 2020; Liu et al., 2020a, 2020b) and environmental matrices (Net et al., 2015a), considering the main routes of entry such as inhalation and ingestion (dietary and non-dietary). However, few studies have evaluated exposure to different levels of phthalates in clothing and the contribution of the dermal route, which has recently also been considered significant (Rovira and Domingo, 2019; Bu et al., 2021; Yao et al., 2022). This paper presents a review of articles published in the last decade (2014–2023) concerning phthalate contents in textile materials. The objective of the review, which included more than 120 articles, was to provide an overview of the literature concerning the phthalate contents of textile and related products. Thus, findings regarding types of phthalates, exposure routes, analytical methods of determination, PAE concentrations determined in clothing and textile products were critically evaluated in relation to human exposure and potential impact on * Corresponding author. E-mail addresses: [email protected] (N. Aldegunde-Louzao), [email protected] (M. Lolo-Aira), [email protected] (C. Herrero-Latorre). Contents lists available at ScienceDirect Environmental Toxicology and Pharmacology journal homepage: www.elsevier.com/locate/etap https://doi.org/10.1016/j.etap.2024.104457 Received 4 October 2023; Accepted 21 April 2024 Environmental Toxicology and Pharmacology 108 (2024) 104457 2 human health. 1.1. Synthesis, classification and use of phthalates in the textile industry 1.1.1. Synthesis of phthalates Phthalates can be obtained through the acid-catalyzed esterification of phthalic anhydride, with two molar equivalents of an alcohol or an alcohol blend. PVC first became commercially available in 1931, and the subsequent synthesis of DEHP in 1933 led to the rapid growth of the flexible PVC industry (Graham, 1973). Since then, PAEs have been widely used in PVC, and they are now present in a huge variety of consumer products. The total world production of PAEs has recently been estimated to be 6–8 million tons a year (Seyoum and Pradhan, 2019), with 80 % of this being used as plasticizers (Yang et al., 2015). The global market continues to rise, and DEHP is the PAE consumed in greatest quantity (3.07 million tons according to data from 2017) (Wang and Qian, 2021). The PAEs used in the textile industry are obtained by synthesis. However, since 1967, more than 50 phthalate ester derivatives have been reported to be produced naturally by living systems, including algae, bacteria, fungi, actinomycetes and higher plants and animals (Roy, 2020). The naturally occurring PAEs share a structural similarity with their synthetic counterparts, although distinctions have been observed concerning 14 C abundance and bonding structure (Chen, 2004). Moreover, several studies exploring the natural synthesis of PAEs and the optimal conditions for their production from living organisms (Roy et al., 2006; Roy and Sen, 2013; Tian et al., 2016; Ortíz and Sansinenea, 2018) confirm the organic origin and unveil potential and promising avenues for production of environmentally friendly PAEs. 1.1.2. Classification of phthalates PAEs are classified according to their molecular weight, into two different groups. The low-molecular-mass PAEs (LMM-PAEs) include phthalate esters in which the side chains are composed of 1–4 carbons, e. g. dimethyl phthalate (DMP), diethyl phthalate (DEP) and dibutyl phthalate (DBP) (See Fig. 1B). High-molecular mass PAEs (HMM-PAEs) are esters with side chains containing 5 or more atoms of carbon and include butyl benzyl phthalate (BBP), di(2-n-ethylhexyl) phthalate (DEHP), diisononyl phthalate (DINP), diisodecyl phthalate (DIDP) and di(n-octyl) phthalate (DNOP), among others (Fig. 1C). A non-exhaustive list of most common phthalates and their characteristics is presented in Table 1. Although any alcohol can be used to produce a PAE, only the esters of the C1 to C13 alcohols are useful as plasticizers; the performance of each PAE is thus greatly influenced by the number of carbon atoms in the radicals. HMM phthalates obtained from alcohols with 7–13 carbon atoms are the best plasticizers, suitable for flexibilization of most PVCs. These compounds are known as primary plasticizers because they improve the elongation and softness of PVC-based plastics (Sastri, 2014). Phthalates with more than 13 carbon atoms in the radicals have poor compatibility for PVC applications; diisotridecyl phthalate (DITP) is the highest molecular weight (13 carbon atoms) phthalate used as a plasticizer. By contrast, plasticizers produced from alcohols with 4–6 carbon atoms are too volatile to be used by themselves. However, they can be combined with other compounds as secondary plasticizers because they are capable of enhancing flexibility at low temperatures. Among high molecular weight PAEs, DEHP is the most widely used as a plasticizer in PVC. DEHP is therefore present in all consumer products that are made from or include soft PVC parts in their composition. Other HMM-PAEs that are also used in the plastics industry for this task are DINP and DIDP (Larsen and Butte, 2017). Moreover, the addition of these HMM-PAEs greatly improves the durability of PVC. Thus, these compounds are used in the manufacture of industrial PVC-based products such as wires, cables, flooring, tarpaulin, wall coverings, synthetic leather and PVC patches (for clothing), roofing membranes and automotive materials (Chang et al., 2021). Finally, LMM phthalates from alcohols with fewer than 4 carbon atoms are not used as plasticizers because undesirable vaporization occurs at the PVC processing temperatures (180–210 ◦C) (Krauskopf, 2009). Low molecular mass PAEs (such as DMP, DEP and DBP) are present in some solvents or carriers in perfumes, varnishes and coatings (Gong et al., 2014). They are also included in adhesives and epoxy resins, waxes, inks, pharmaceutical products, insecticide formulations and several types of cosmetics. Moreover, DMP and DEP yield slower evaporation of perfumes, increasing the duration of fragrances. Small amounts of DBP also enhance the anti-chipping properties of nail polish (Net et al., 2015a). Fig. 1. A) General structure of ortho-phthalate esters; B) Examples of low-molecular-mass ortho-phthalate esters; C) Examples of high-molecular-mass orthophthalate esters. N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 3 1.1.3. Use of phthalates in the textile industry Since the 1930 s, PAEs have been widely used as additives in plastics manufacturing, mainly as plasticizers in polyvinyl chloride (PVC), but also in polyethylene terephthalate (PET), rubber and styrene production. PAEs act by increasing the flexibility, transparency and durability of these polymers. Phthalates have different applications in the textile industry, most of which are related to screen printing and coating of fabrics (Mohapatra and Gaonkar, 2021). The PVC used in these tasks must be softened before use, and phthalates are used for this purpose. Plasticizers are widely used in screen printing and coating fabrics, often making up around 30–60% of the total composition of the plastic part (Giuliani et al., 2020). In the specific case of the textile industry, three main sources of phthalates have been identified: (i) Processing chemicals containing phthalates. Dyes and textile inks, as well as printing plastisols, are common examples of materials containing phthalates employed in textile production; (ii) Raw materials and pre-production products treated with phthalates during their manufacture, such as plasticized fabrics, textile waterproof coatings and raincoats, synthetic leather, and rubber, among others; and (iii) The potential contamination of textile products by chemical impurities or additives contained in processing chemicals or water. In addition, the storage and/or packaging of textile products in contact with other materials that release phthalates (plastic packaging films, plastic containers and plastic labels) are other possible relevant sources (Rewe Group, 2018). 1.2. Phthalates as plasticizers High molecular-mass phthalate esters are functional additives used as plasticizers in PVC production (Hahladakis et al., 2018). Approximately 90% of all plasticizers produced globally are used to provide flexibility to this synthetic polymer (Jamarani et al., 2018). This can be explained by the fact that PVC is highly compatible with numerous plasticizer structures at a relatively wide range of concentrations and different temperatures. Amongst plastic polymers, PVC is a particularly good candidate for plasticization because of the polarity, helical structure and balance of amorphous and crystalline regions in its molecular geometry. Plastic polymers are interfused with plasticizers to increase their flexibility, transparency, workability and durability. This is achieved by lowering the glass transition temperature (T g ) to below the temperature at which the resulting flexible PVC will be used (Graham, 1973). The T g can be lowered internally or externally. Internal plasticizers lower the glass transition temperature by “grafting” or covalently bonding softer monomers units to the polymer chain. By contrast, external plasticizers (such as DEHP) are simply mixed with the polymer at high temperature and do not establish covalent bonds (Jamarani et al., 2018). External plasticizers are the mostly commonly used because they are added after polymerization, which enables better adjustment of the final properties of the PVC obtained. In addition, as there is no reaction between the plasticizer and the polymer, the external process is more cost-effective. Chemically, the plasticizing capacity of PAEs has been explained by Herbert et al. (2017) (using the gel theory according to Doolittle, 1954) on the basis of attenuation of the Van der Waals forces between plastic polymers produced by the inclusion of PAEs between the polymer chains (See Fig. 2). The PVC produced under these conditions is a resistant but flexible elastomer, rather than a rigid solid. Because of the absence of covalent bonding between the plasticizers and PVC, PAEs can migrate from the plastic and other consumer products in which they are used. Phthalates can be released from soft PVC plastic by surface contact, especially when mechanical pressure is applied (e.g. chewing on a PVC teether), as they are not strongly bound (Duty et al., 2003). Moreover, as PAEs are abundant in everyday products (cosmetics, building materials, car parts, liners, medical and pharmaceutical devices, food packaging, clothing and others such as children’s toys, in which they are permitted to make up 1040% of the Table 1 Characteristics of the most common phthalates. Name Abbreviation Alcohol C number Mol. Mass (g mol ¡1 ) CAS No. Butyl benzyl phthalate BBP 4–7 312.36 85–68–7 Butyl cyclohexyl phthalate BCP 4–6 304.38 84–64–0 Butyl decyl phthalate BDP 4–10 362.5 89–19–0 Bis-4-methyl-2-penthyl phthalate BMPP 6 334.36 146–50–9 Diallyl phthalate DAP 3 246.26 131–17–9 Dibutoxy ethyl phthalate DBEP 6 366.45 117–83–9 Di-n-butyl phthalate DBP 4 278.34 84–74–2 Dicyclohexyl phthalate DCP 6 330.42 84–61–7 Di-2-ethylhexyl phthalate DEEP 4 310.34 605–54–9 Di(2-ethylhexyl) phthalate DEHP 8 390.56 117–81–7 Bis(2-ethylhexyl) terephthalate DEHT 8 390.60 6422–86–2 Diethyl phthalate DEP 2 222.24 84–66–2 Diisobutyl phthalate DIBP 4 278.34 84–69–5 Diisodecyl phthalate DIDP 10 446.66 26761–40–0 Diisoheptyl phthalate DIHPP 7 362.5 41451–28–9 Diisohexyl phthalate DIHXP 6 334.45 146–50–9, Diisononyl phthalate DINP 9 418.61 28553–12–0 Diisooctyl phthalate DIOP 8 390.56 27554–26–3 Diisotridecyl phthalate DITP 13 530.82 68515–47–9 Diisoundecyl phthalate DIUP 11 474.72 85507–79–5 Di-2-methoxyethyl phthalate DMEP 3 282.29 117–82–8 Dimethyl phthalate DMP 1 194.18 131–11–3 Di-n-hexyl phthalate DNHP 6 334.45 84–75–3 Di(n-octyl) phthalate DNOP 8 390.56 117–84–0 Di-n-nonyl phthalate DNOP 9 418.6 84–76–4 Di-n-pentyl phthalate DNPP 5 306.4 131–18–0 Di(2-propylheptyl) phthalate DPHP 10 446.66 53306–54–0 Di-n-propyl phthalate DPP 3 250.29 131–16–8 Ditridecyl phthalate DTDP 13 530.82 119–06–2 Diundecyl phthalate DUP 11 474.72 3648–20–2 n-Octyl n-decyl phthalate ODP 8–10 418.61 119–07–3 N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 4 total weight of the plastic), they are now widespread in all environmental compartments, in living organisms, especially humans (Xie et al., 2016). In short, PAEs are ubiquitous (Net et al., 2015b). 1.3. Exposure routes for PAEs in humans Due to the ubiquity of these compounds, exposure of humans to PAEs takes place through multiple routes such as parenteral administration, inhalation, oral ingestion and dermal absorption. In addition, the main route can vary for each different PAE or PAE type (Wang et al., 2020; Chang et al., 2021). Most plastic medical devices used for intravenous infusions, including blood bags and infusion tubes, as well as parenteral nutrition feed bags and nasogastric tubes, are manufactured using PVC containing PAEs. Thus, exposure through the parenteral route can take place in patients during blood infusions or administration of parenteral nutrition (Panneel et al., 2021; Saab et al., 2022). As explained above, PAEs are semi-volatile compounds that are not covalently bound to polymeric plastic matrices. Therefore, inhalation exposure to PAEs occurs when phthalates are volatilized from PVC or other plastic polymers, nail polish, personal care products, etc. Due to the low vapour pressure of PAEs, high temperatures in ambient air favour the release of these compounds from PVC and plastic products to the atmosphere and the environment, and their subsequent adsorption onto indoor airborne particulate matter and surfaces. In a recent study, Wang et al. (2020) reported that dust inhalation is an important route of indoor exposure to PAEs in pre-school children in Beijing. Bu et al. (2016) studied the indoor phthalate concentration in residential apartments in Chongqing (China) and the implications for exposure in preschool children. Risk assessment revealed that exposure to DBP and DEHP surpassed child-specific levels for more than 90% of preschool children in this area. In general, the oral route is considered the most important for most PAEs used in the industry as plasticizers, such as HMM-PAEs and particularly DEHP (Guo et al., 2012; Martínez et al., 2018; Bølling et al., 2020). Dietary exposure is an obvious consequence of the consumption of food and drinks contaminated with PAEs that have migrated from plastics or other plasticized materials used in the packaging (Alp and Yerlikaya, 2020; Perestrelo et al., 2021; Edwards et al., 2022). However, high concentrations of PAEs have been found in vegetables, crops and fruits produced in contaminated soils (Lü et al., 2018) and in greenhouses with plastic covers containing PAEs (Zhang et al., 2019). In the past few years, researchers have suggested the importance of non-dietary routes of exposure/absorption and the need for more research on this topic, as dermal absorption may make a considerable contribution to phthalate exposure (Bu et al., 2018). Specifically, dermal exposure has received little attention, even though absorption is expected to take place through the skin, due to the highly lipophilic nature of PAEs. Recent in vitro studies (Zeng et al., 2019) have shown that the high dermal accessibility to different phthalates (especially DBP) indicate that the dermal route is important. In 2011, Kissel (2011) explained why dermal absorption was underestimated in the different studies because the exposure was assessed exclusively in terms of fractional absorption. Other authors have also stated that the dermal route has been underestimated (Weschler and Nazaroff, 2012; Gong et al., 2016; Zhao et al., 2022), and several recent studies have considered the dermal route to be important in living environments (Bu et al., 2018, 2021; Yao et al., 2022). In addition, dermal exposure may reflect exposure from multiple sources, such as direct absorption from the gas phase, particle deposition and contact with contaminated surfaces such as clothing (Bu et al., 2018; Licina et al., 2019). Thus, dermal absorption of PAE through skin contact with clothing containing phthalates is an important route of exposure, with a particularly important impact on children (Rovira and Domingo, 2019). Therefore, this route should be studied in detail, with particular attention given to textile products in which phthalates have been used as plasticizers (e.g. clothing, fabrics and plastic parts, including buttons, coatings, stamped designs and plastisol prints). 1.4. Effects of PAEs on human health Phthalates are bound to plastic polymers through van der Waals forces and hydrogen bonds. The weak bonding leads to the release of phthalates in most environmental media. Thus, humans are exposed to these compounds during the production and use of different products made or containing PVC, PET and other polymers, such as polyurethane, polystyrene and polybutadiene (Fromme et al., 2007). Potentially toxic, poorly biodegradable and lipophilic, PAEs have the capacity to accumulate in adipose tissue (Salazar Beltr´ an, 2019). Humans exposure to PAEs mainly takes place through three pathways: inhalation, ingestion and dermal absorption (Zhao et al., 2022). Different studies have linked these types of exposure to potential adverse effects on human and animal health (Brassea-P´ erez et al., 2011; Yue et al., 2022). Specifically, exposure to several PAEs has been associated with disruption of the endocrine system (Wang and Qian, 2021) and reproductive toxicity (Domínguez, 2019; Xie et al., 2019; Virant-Klun, Fig. 2. Plasticization of PVC by DEHP. Adapted from Herbert et al. (2017). N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 5 2022). PAEs can interfere with hormonal homeostasis because they are able to bind to molecular targets in the human body, producing different types of dysfunctions in children and adults (Katsikantami et al., 2016; Radke et al., 2019; Dong et al., 2022). In a recent review, Chang et al. (2021) observed a consistent association, demonstrated by epidemiological studies, between PAE exposure (particularly for DEHP and its metabolites) and a decrease in sperm quality in males. In fact, other studies also proved that BBP and DINP affect testosterone and semen quality parameters (Radke et al., 2018). However, these authors found there was insufficient evidence and a lack of consistency regarding the association between PAE exposure and cardiovascular diseases and other reproductive system-related diseases. Other studies have shown that, even at low concentrations, PAEs can have teratogenic (Czubacka et al., 2021; Sedha et al., 2021), mutagenic (Szendi et al., 2018; Xue et al., 2021) and carcinogenic effects (Zuccarello et al., 2018; Mughees et al., 2022). High levels of DEHP urinary metabolites have been associated with a higher risk of thyroid cancer and benign thyroid nodules (Liu et al., 2020a, 2020b). Phthalates are already recognized as a risk for breast cancer (Srivastava et al., 2022). Recent studies have confirmed the positive correlation between phthalate exposure and increased breast cancer risk. More specifically, DBP and DEHP have been associated with increased risk of invasive breast cancer in Indian women, as well as with potential mutations in specific genes involved in the biosynthesis of steroid hormones, inflammation and chemical metabolism (Ahern et al., 2019; Crobeddu et al., 2019). Das et al. (2022) revealed a significant correlation between DBP and DEHP exposure and an elevated risk of invasive breast cancer in Indian women. These phthalates were linked to potential mutations in specific genes related to the biosynthesis of steroid hormones, inflammation and chemical metabolism. However, the researchers pointed out that the specific mechanisms through which phthalates induce mammary carcinogenesis are complex, remain unknown and should be the subject of further research. Recently, Trasende et al. (2022) examined the association between urinary phthalate levels with all-cause and cause-specific (cardiovascular disease and cancer) mortality. The study involved a nationally representative cohort of 5030 adults (≥20 years) in the USA (2001–2010). The findings showed evidence of increased cardiovascular mortality related to HMW-PAE metabolites, especially DEHP metabolites. Given the adverse metabolic effects also associated with DINP and DIDP, further research is essential to assess the impact of these emerging PAEs, which have replaced DEHP in the last decade. This is particularly important due to the latency period of cardiovascular disease. In all cases, researchers highlight the need to control phthalate levels in the environment as well as in consumer products, and they recommend the introduction of more restrictive measures for the use of PAEs. 2. Phthalates in textile materials Phthalates are extensively used in a variety of products and have potentially widespread and undesired effects on animal, human and environmental health. As a result, authorities and regulatory bodies have promulgated and approved appropriate legislative actions with the goal of reducing exposure to phthalates (particularly in groups who are especially sensitive to their effects, such as children). In addition, valuable and abundant research has been carried out in relation to the analytical methods used to measure phthalates and to determine the levels of these compounds in different biological, environmental and industrial matrices. However, very few articles have been published in recent years regarding the PAE contents in clothing and other textile products. The most notable aspects of the legislation on PAEs, as well as the analytical methodologies used to determine these compounds will be presented in this chapter. Moreover, a literature survey of recent studies in which phthalates have been determined in distinct textile products will be critically addressed in relation to the effects on human health. 2.1. Legislation on PAEs Phthalates are considered toxic, on the basis of evidence summarized in the previous section, and strong limitations concerning their use have been included in European Union (EU) legislation. The PAEs considered most problematical are regulated by Regulation (EU) no. 1907/2006 (European Parliament and the Council, 2006) and its amendments, Regulation (EU) no. 552/2009 (European Union Commission, 2009) and Regulation (EU) no. 2015/326 (European Union Commission, 2015). Restrictions on the concentrations of six phthalates (BBP, DBP, DEHP, DIDP, DINP and DNOP) in plasticized and PVC materials, especially those aimed at children, have been established in the EU. The use of BBP, DBP and DEHP in PVC and other plasticized materials, as well as in all toys and childcare articles, is thus restricted. The use of DIDP, DINP and DNOP in clothes, toys and childcare articles which children can place in their mouths is also limited. In all cases, PAEs should not be present at concentrations higher than 1000 µg g −1 (equal to 0.1% by weight). More recently (in 2018), new legislation outlined in Regulation (EU) no 2018/1513 (European Union Commission, 2018) included other phthalates and extended the restriction to all articles and goods of consumption (including fashion items) that contain plastic parts. The maximum level of 1000 µg g −1 is maintained (individually or in combination with other previously regulated phthalates). Comparable restrictions have been included in US legislation by the United States Consumer Product Safety Commission (US-CPSC). In 2008, section 108(a) of the Consumer Product Safety Improvement Act (CPSIA) permanently prohibited any children’s toy or childcare article that contains more than 0.1% of DEHP, DBP or BBP. In addition, section 108(b)(1) prohibited on an interim basis any children’s toy that can be placed in a child’s mouth or childcare article containing more than 0.1% of DINP, DIDP or DNOP (US-CPSC, 2008). In 2017, the United States Consumer Product Safety Commission definitively banned DINP at concentrations higher than 0.1% (1000 mg kg −1 ). Another four phthalates (DIBP, DPENP, DHEXP and DCHP) were also restricted for use in plastic component parts of toys and childcare articles (US-CPSC, 2017). Similar legislation with the same restrictions on PAEs is in force in Japan, Canada, Taiwan, Turkey, China and other countries. The different regulations concerning PAEs in different parts of the world are summarised in Table 2. There seems to be general consensus regarding the limits of concentrations: all legislation situates the maximum content of six phthalates (DEHP, DBP, BBP, DINP, DIDP and DNOP, alone or together) at 0.1% by weight. In addition, other phthalates, such as DIHP, DMEP, DIPP, DPP, DNHP, DCP, DIBP and DNPP, are also restricted for use in certain types of consumer products. The strict control of regulated phthalates by authorities and regulatory bodies is a critical task for the following reasons: (i) Although phthalates are strictly regulated in most countries, they continue to be present in products that are still in use and will continue to be used for a long time; (ii) In some other countries, PAEs remain poorly or unregulated, and they continue to be used in the fabrication of cosmetics, PVCbased toys, school supplies and in textile products (Wang et al., 2019); and (iii) Due to these legal restrictions, the prohibited phthalates have been replaced by other non-regulated phthalates, by the simple strategy of modifying the R and R` radicals (Dodson et al., 2012; Wang et al., 2018). As a consequence of this replacement, a decrease in the production of some PAEs (such as DBP and DEHP) has been observed (Wang et al., 2018). However, exposure to these new PAEs must be evaluated because the substitute phthalates may have similar toxic properties as the banned ones (Aldegunde-Louzao et al., 2023). Controlling these compounds in a wide variety of articles and consumer goods is an important task that must be carried out using accurate, precise analytical methodologies. The analytical methods available for the determination of PAEs in consumer products are described in the next section. N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 6 Table 2 Summary of the regulations concerning PAEs in different countries worldwide. Country Limit Notes Legal reference European Union 1000 mg kg −1 1000 ppm of DEHP, DBP and BBP in the plasticized material of toys and childcare articles; and 1000 ppm of DINP, DIDP and DNOP in the plasticized material of toys and childcare articles that children can put in their mouths. European Regulation 1907/2006/CE, in force since 2007; and its amendments 552/2009/CE, in force since 2009, and 2015/ 326/UE, in force since 2015. 1000 mg kg −1 Footwear will not be marketed when any of the following phthalates (DIHP, DMPE, DIPP, DPP, DNHP, individually or in combination with other phthalates) are present at concentrations, measured in homogeneous material, equal to or greater than 1000 mg kg −1 . European Regulation 1513/2018. (Modification Annex XVII of REACH Regulation) 1000 mg kg −1 Substances or mixtures will not be used when the concentration of any of the following phthalates (DEHP, DBP, BBP, DIBP, individually or in combination with other phthalates) is equal to or greater than 0.1% by weight (1000 mg kg −1 ) of the plasticized material, in toys and childcare articles. European Regulation (EU) 2005/2018. (Modification Annex XVII of REACH Regulation) Switzerland 1000 mg kg −1 1000 ppm of DEHP, DBP and BBP in childcare articles and 1000 ppm of DINP, DIDP and DNOP in childcare articles that young children can put in their mouths. SR 817.023.41. Ordinance on Articles for the Human Contact, effective 2005 and its amendments RO 2006 5121, in force since 2007; RO 2010 4763, in force since 2010; and RO 2019 3405, in force since 2019. Canada 1000 mg kg −1 The vinyl in a toy or childcare article must not contain more than 1000 mg kg −1 of DEHP, DBP or BBP when tested in accordance with a method that conforms to good laboratory practices. The vinyl in any part of a toy or childcare article that children under four years of age can put in their mouths, must not contain more than 1000 mg kg −1 of DINP, DIDP or DNOP when tested in accordance with a method conforming to good laboratory practices. Canada Consumer Product Safety Act (S.C. 2010, c.21), in force from 2011 to March 2023 and Phthalates Regulations (SOR/2016–188), in force from 2016 to March 2023. Turkey 1000 mg kg −1 (a) DEHP, DBP and BBP are limited in plasticized materials used to produce footwear. The sum of these three phthalates must not exceed the limit value of 0.1% (1000 mg kg −1 ). (b) DINP, DIDP and DNOP are also limited in childcare articles that children can put in their mouths. The sum of these three phthalates must not exceed of 0.1% (1000 mg kg −1 ). Regulation on Restrictions for the Manufacture, Marketing and Use of Certain Dangerous Substances & Preparations (Official Gazette No 27092 of 26 December 2008), in force since 2008; and its amendments: Official Gazette No 27687 of 29 August 2010, in force since 2010; Official Gazette No 27880 of 20 March 2011, in force since 2011; and Official Gazette No 29182 of 21 November 2014, in force since 2014. Market Surveillance and Control of Hazardous Chemicals Content in Some Consumer Products (Official Gazette No 29236 of 14 January 2015), in force since 2015. Communiqu´ e on Import Controls for Certain Consumer Products, Product Safety and Inspection: 2015/18 (Official Gazette No 29222 of 31 December 2014), in force 2015. The Turkey Ministry of Trade has published an order in the Official Gazette No. 31297 on 7th November 2020 regarding the import inspections of products in the Customs to be into force on 01/01/21. USA 1000 mg kg −1 The production, sale, distribution in commerce and import of articles, or parts thereof, designed or intended for children under 12 years of age that contain concentrations greater than 0.1% (1000 mg kg −1 ) of DEHP, DBP, BBP, DINP, DIBP, DPENP, DCHP and DNHP is prohibited. This limit applies to each individual phthalate. Section 108 of the Consumer Product Safety Improvement Act (CPSIA) of 2008 (Public Law 110–314; 15 U.S.C.2057c) and modified by Section 5 of the CPSIA of 2011 (Public law 112–28) and the Final Rule of 2018 (16 CFR part 1307). China 1000 mg kg −1 1000 ppm of DEHP, DBP and BBP, and 1000 ppm of DINP, DIDP and DNOP in coated or printed textile products intended for children under 3 years of age. GB 31701–2015 Covering Infant and Children’s Textile Products. China has issued a new product standard FZ/T 08002–2022 (Textile products for newborn). In addition, a new national product standard, GB/T 41002–2022 (General specifications for children’s cases and bags). Both standards are in force from October 1, 2022. 1000. mg kg −1 (a) DEHP, DBP, and BBP are limited in footwear for children under 14 years of age. (b) DINP, DIDP and DNOP, are limited in baby shoes. The sum of phthalates from both (a) and (b) must not exceed the limit value of 1000 mg kg −1 . GB 30585–2014 Safety technical specifications for children’s footwear. Egypt 1000 mg kg −1 1000 mg kg −1 of BBP, DBP, DEHP, DNOP, DIDP and DINP in textile garments. 1000 mg kg −1 for the sum of phthalates in synthetic leather products and in textile and synthetic leather parts of leather footwear intended for children under 3 years of age. 1000 mg kg −1 for the sum of phthalates (DEHP, DBP and BBP) in the plasticized material of toys and childcare products intended for children and 1000 mg kg −1 for the sum of phthalates (DIDP, DINP and DNOP) in the plasticized material of toys and childcare products that children can put in their mouths. ES 7266/2011. Safety and Health Criteria and Labeling for Textile Products. Part 4: Clothes, in force since 2011. ES 7322/2011. Standards of Safety & Hygiene in Leather, Leather Products & Parts, in force since 2011 ES 7562/2013. Restriction on the Use of Phthalates and their Derivatives in Toys and Childcare Articles, in force since 2013. (continued on next page) N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 7 2.2. Analytical methodologies for the determination of PAEs In general, the determination of PAEs in all types of samples requires an extraction step to be carried out before the instrumental measurement for quantification. Until recently, the Soxhlet extraction method was traditionally used to extract PAEs from solid samples. However, in the last decade this method has been replaced by other approaches that are more respectful of the principles of green chemistry. In addition, the novel extraction strategies are faster and more efficient, with less risk of contamination (Net et al., 2015b). Liquid-liquid extraction (LLE) is one of the methods most commonly used. LLE consists of adding a small volume of an organic solvent (propanol, hexane or other) to the aqueous sample containing the PAEs, shaking the contents and collecting the PAEs in the organic phase after decantation. LLE is simple to apply, but time consuming and requires moderate amounts of organic solvents. Other alternatives to LLE have been developed to miniaturize the extraction system: liquid-liquid microextraction (LLME) (Farahani et al., 2008) and dispersive liquid-liquid microextraction (DLLME) (Xue et al., 2014). Solid phase extraction (SPE) is also used because it is easy to apply, saves time and solvents and can be automated (Salazar-Beltr´ an et al., 2017). Typical devices for separating and concentrating PAEs from liquid samples consist of cartridges (made from polyethylene, polypropylene or pure PTFE) containing different sorbents such as octadecylsilane (OCS) C18, HLB, LiChrolut RP18 and Lichrolut EN (Net et al., 2015b). Moreover, other variants of SPE, such as dispersive SPE (DSPE) and magnetic SPE (MSPE), have been used to determine phthalates with more sophisticated sorbents based on nanomaterials such as carbon nanotubes, graphene and others. Gonz´ alez-S´ alamo et al. (2018) have recently developed a microdispersive SPE procedure for the extraction of nine phthalic acid esters in which multiwalled carbon nanotubes are used as sorbents. In another example of the application of DSPE, Hu et al. (2022) measured 16 phthalates in marine sediments. The performance of the proposed approach compared favourably with that of traditional Soxhlet extraction. On the other hand, magnetic extraction has also been used for PAE extraction. A magnetic o-hydroxyazobenzene (M-HAzo) porous organic polymer was used as an example of new magnetic sorbent for the preconcentration of phthalate esters from juice contained in plastic bottles (Wu et al., 2021). Solid phase microextraction (SPME) and stir-bar sorptive extraction (SBSE) have also been used to extract the analytes in other PAE determination methods (Si et al., 2016). Microwave-assisted extraction (MWAE) (Llompart et al., 2019; Sinha et al., 2019) and ultrasonic extraction (USE) (Fern´ andez-Gonz´ alez et al., 2017; Hidalgo-Serrano et al., 2022) have also been used to extract PAEs following the principles of green analytical chemistry. Finally, thermodesorption (TD) is an extraction technique that does not need sample preparation as the sample is heated to desorb PAEs from the solid matrices. The TD technique allows direct analysis of the sample with good reproductivity and data quality (Net et al., 2015b). The technique is often used for quantifying PAEs in plastics used in food packaging, disposable cutlery, plates and cups, and for the construction of medical devices or plastic toys (Ouchi et al., 2019; Wang et al., 2022). Different techniques have been used to quantify phthalates extracted from different types of samples. Optical and electrochemical sensors based on nanomaterials and molecularly imprinted polymers (MIPs) Table 2 (continued) Country Limit Notes Legal reference India 1000 mg kg −1 1000 mg kg −1 of BBP, DBP, DEHP, DNOP, DIDP and DINP in products or materials designed or clearly intended, whether or not exclusively, for use in play by children under 14 years of age. IS 9873 (Part 9): 2017. Certain phthalates esters in toys and children’s products. Taiwan 1000 mg kg −1 General requirements for the safety of children’s products. Limits: 1000 mg kg −1 for the sum of phthalates (DEP, DMP, DEHP, DBP, BBP, DINP, DIDP and DNOP) in the plastic parts of products intended for children under 14 years of age, such as clothing, footwear, towels, bedding and backpacks, among others. Plastic Shoes. Limits: 1000 mg kg −1 for the sum of phthalates (DEHP, DBP, BBP, DINP, DIDP and DNOP) in plastic shoes with rubber or plastic soles. Leather Casual Shoes. Limits: 1000 mg kg −1 for the sum of phthalates (DEHP, DBP, BBP, DINP, DIDP and DNOP) in the plasticized material of leather shoes with rubber or plastic soles. Leather Shoes. Limits: 1000 mg kg −1 for the sum of phthalates (DEHP, DBP, BBP, DINP, DIDP and DNOP) in the plasticized material of leather shoes with leather soles. CNS 15503. General Requirements for Safety of Children’s Products, in force since 2011. CNS 3478. Plastic shoes, in force since 2009. Two voluntary footwear standards, CNS 8634 (Leather Casual Shoes) and CNS 10632 (Leather Shoes), indicate the basic chemical and quality requirements of leather shoes in the Taiwan market. CNS 8634. Leather Casual Shoes, in force since 2011 CNS 10632. Leather Shoes, in force since 2011. Japan 1000 mg kg −1 (1) DBP, DEHP or BBP shall not be contained at concentrations of greater than 0.1% by mass (1000 mg kg −1 ) of the plasticized material in toys. (2) DIDP, DINP or DNOP shall not be contained at concentrations of greater than 0.1% by mass (1000 mg kg −1 ) of the plasticized material in toys as specified in Article 78, Paragraph 1 of the regulation of the Food Sanitation Law (this requirement applies only to their parts that are intended to be placed in the mouth). (3) DINP shall not be used as toy materials in which the major component is PVC-based resin (except the parts that are intended to come in contact with infant’s mouth, as defined in 2) as specified in Article 78, paragraph 1 of the enforcement regulations of the Food Sanitation Law. Notification N◦336 of Food Sanitation Law. United Arab Emirates, Bahrain, Saudi Arabia, Oman, Qatar, Kuwait and Yemen 1000 mg kg −1 1000 mg kg −1 for the sum of phthalates (DINP, DNOP, DEHP, DIDP, BBP and DBP) in textile and leather products intended for children younger than 2 years. GSO 1956/2009 Harmful substances used in textile products, in force since 2009. N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 8 (Zhang et al., 2023a), and aptasensors (Chen et al., 2021) have recently been used to detect PAEs in environmental samples. Raman spectrometry can be used to measure DBP in foods (Rong et al., 2021). However, high pressure liquid chromatography (HPLC) and (above all) gas chromatography-mass spectrometry (GC-MS) are the techniques most commonly used to determine these compounds in diverse matrices (Wenzl, 2009), such as foods (Panio et al.; 2020), water (Amritha et al., 2022), environmental samples (Hidalgo-Serrano et al., 2022), consumer goods (Khan and Jahangir, 2020) and other types of samples, including textiles (Aldegunde-Louzao et al., 2023). Micellar electrokinetic chromatography (MEKC) has been used as an alternative to HPLC and GC. In this case, different surfactants, e.g. sodium deoxycholate (SDC), are added to the separation buffer as a pseudostationary phase. Under these conditions, several PAEs have been determined in perfumes (P´ erez-Fern´ andez et al., 2013). However, several authors highlight the poor resolution and lack of sensitivity of this technique relative to traditional HPLC and GC-based techniques. In the specific case of the determination of PAEs in textile samples, slightly different analytical approaches are used depending on whether the analysis is for the control of phthalates in the textile industry or for research purposes. Although the methodology used in both cases is generally based on GC-MS, in control laboratories the extraction and quantification methodologies used are those specified in the USA and European Union regulations: CPSC-CH-C1001–09.4 (US-CPSC, 2018) and EN ISO14389:2022 (ISO, 2022), respectively, or in-house adaptations of these methods. In addition to those described in the official regulations, other extraction approaches or different chromatographic conditions are also used in research studies. In the case of PAE control, the plasticized parts of the textile products possibly containing phthalates must be cut into small pieces of approximately 1 mm 2 to eliminate the potential error due to unequal distribution of pollutants in the sample. The hard parts (such as plastic or plasticized buttons, or other similar parts as sequins, flocks, glitter, pearls and other plastic ornaments) are milled/ground into a representative powder prior to analysis. For plastic coated-metal or coated-wood parts, the surface is scratched, and the coating or paint is assayed. After the sample preparation step, extraction is based on the solubility of PVC in tetrahydrofuran (THF). The textile specimen sample (accurately weighed) is treated with THF in an ultrasonic bath. Once the plastic polymer is completely dissolved, it is precipitated in the appropriate solvent (n-hexane, acetonitrile or methanol). The solution thus obtained is centrifuged and concentrated, and the extract is diluted to a certain volume. Finally, in the quantification step, PAEs in the extract are determined by GC-MS, under the following conditions established in both EU and USA standards: Column, DB-5MS (commonly of 30 m x 0.25 mm ID x 0.25 μ m); carrier gas, He (or H 2 ) at 1 mL min −1 ; injection volume, 1 μ L; inlet temperature, 290 ◦C; oven program, initially 150◦C for 1 min, ramp 30 ◦C min −1 to 280 ºC and ramp 15◦C min −1 to 310 ◦C; final hold time 3 minutes or longer. Samples are analyzed both in full scan mode and in selective ion monitoring (SIM) mode. Since 2019, interlaboratory studies have been conducted annually to test the proficiency for the measurement of phthalates in textile samples by control laboratories working for the textile industry. In the last edition, in 2022 (Starink, 2022), 71 laboratories in 27 different countries were required to measure sixteen individual PAEs in cotton samples. Half of the participants used the ISO14389 test method (the official methodology in the European Union: ISO, 2022). Another 25% used the CPSC-CH-C1001–09.4 test procedure (the official method in the USA: US-CPSC, 2018). The remaining 25% of participants used in-house developed test methods consisting of adaptations of the previous two tests. Moreover, almost all participants used the ultrasonic approach to PAE extraction, with THF or a THF mixture as extraction solvent. The analytical characteristics of the extraction and determination procedures used to measure PAEs in textile materials in the last decade are summarized in Table 3. Regarding extraction, traditional Soxhlet extraction with dichloromethane was used in several cases. However, this approach uses large volumes of solvents and is time-consuming, taking between 4 and 6 hours (Gong et al., 2014, 2016; Li et al., 2015; Negev et al., 2018). Other strategies that are more respectful of the green chemistry principles are thus used: ultrasound-assisted extraction (UAE) and accelerated solvent extraction (ASE). UAE consists of applying ultrasonic energy (in the form of waves) through a liquid solvent in which the solid textile sample is located. The sonic energy of ultrasound waves is transformed into mechanical pressure equivalent to several thousand atmospheres, facilitating extraction of the desired component. UAE is an attractive alternative because it is cheaper and faster, extracting PAEs in less than 30 min. Common extraction solvents include hexane, dichloromethane, acetone, ethyl acetate and mixtures of these (Boor et al., 2015; Li et al., 2015; Tang et al., 2020; Xie et al., 2022; Zhang et al., 2023b). ASE is a sample preparation technique that uses organic solvents at high temperature and pressure to remove undesirable matrix components from the sample. In the case at hand, Saini et al. (2016) successfully used ASE to extract PAEs from different textile fabrics. The solvent used was a mixture of hexane, dichloromethane and acetone (2:1:1, v/v), and the extractor was operated at 70 ºC and 1500 psi. Finally, other researchers (Xie et al., 2016 and Aldegunde-Louzao et al., 2023) used an extraction procedure based on the solubilization of PVC in THF according to the official US-CPSC test method. Regarding the quantification procedure, GC-MS is the analytical technique of choice in all cases, and capillary non-polar columns (5%) diphenyl-(95%) dimethyl polysiloxane bonded and cross-linked phase are also used. The chromatographic conditions and those of the MS detector were the same as in ISO14389:2022 (ISO, 2022) and CPSC-CH-C1001–09.4 (US-CPSC, 2018), described above. The conditions reported and the MS scanning mode used for the quantification in each study are summarized in Table 3. The only substantial difference between laboratories appears to be the limit of detection (LOD). In several legal control studies, a LOD in the range of µg g −1 (ppm) is sufficient, as the limit values of PAEs is 0.1% (equivalent to 1000 µg g −1 ). By contrast, in research studies involving potential exposure to phthalates, especially in children, the LOD was in the range of ng g −1 levels (ppb). After these considerations concerning the extraction and quantification procedures, two conclusions can be reached: (i) Despite several slight differences in sample amounts, sample preparation and GC-MS conditions, in practice most laboratories and research groups use similar analytical procedures; and (ii) In most control and research laboratories, no problems regarding the determination of PAEs at the concentrations presents in textile samples have been noted. Thus, the sample treatment procedures and the analytical tools available enable the determination of phthalates with appropriate analytical figures of merit in textile samples. In accordance with these findings and considering the homogeneity in the sample treatments and analytical quantification methods reported in the different papers, it should be possible to compare the results provided in textile samples by different researchers in different studies and in different countries. 2.3. Phthalate content of textile products Studies concerning phthalate contents determined in textile products (fabrics, clothing and other related textiles) in the last ten years are summarized in Table 4. 2.3.1. Textile and related products intended for children Most of the papers reviewed involve the determination of phthalates in textiles and related materials that are intended for use by children, because legal restrictions are strongest due to the high potential impact of these compounds on infants. For example, Li et al. (2019) studied the contents of six phthalate esters (DMP, DEP, DIBP, DBP, BBP and DEHP) in twenty-four items of infant cotton clothing produced in Harbin (China). The final objective was to estimate daily PAE intakes through via dermal absorption. Due to the well-known ubiquity of PAEs, as well N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 9 Table 3 Characteristics of the different methods used for the determination of PAEs in textile materials. Sample Type Phthalates analyzed (n) Sensitivity Extraction procedure Quantification technique Reference Firefighter personal protective clothing (gloves, hood, and one coat wristlet) BBP, DEHP, DEP, DBP, DNOP (5) Detection limit in the range 0.2–0.9 µg g −1 depending on the PAE and clothing type For semi-volatile organic compounds by GC-MS. Extraction and cleanup guidance is provided EPA 3600 and related methods (sample treatment and PAE extraction not detailed). GC-MS Details not included. Alexander and Baxter, (2014) Skin wipes DMP, DEP, DIBP, DBP, BBP, DEHP (6) Not available Each sample was extracted with dichloromethane in a Soxhlet extractor (6 h). The extracts were concentrated to 25 mL, filtered, concentrated to a final volume of about 0.3 mL and subsequently subjected to determination by GC-MS GC-MS A fused silica capillary column, HP-5MS (30 m x 0.25 mm ID x 0.25 µm) was used, and He was employed as carrier gas at a flow rate of 1.0 mL min −1 . Injector and ion source Tª 280 and 250 ºC. The temperature program from the initial oven temperature of 100 ºC for 2 minutes included a ramp from 10 ºC min-1 up to 300 ºC, which was maintained for at least 5 minutes. (Scan and Selective ion monitoring modes) Gong et al., (2014) Infant crib mattress covers and foam DEHP, iso-DEHP, DINP, DINCH, DEHA, DMP, DBP, BBP (8) Detection limit 1.0 mg g −1 in covers 0.1 mg g −1 in foam Extraction via ultrasonication with hexane (30 min). The extract was concentrated to 1 mL and subsequently subjected to determination by GC-MS GC-MS A fused silica capillary column, DB-5MS (30 m x 0.25 mm) was used. The temperature program from the initial oven temperature of 80 ºC for 0.5 minutes included a ramp 20 ºC min −1 for 8.5 min and then ramp 30 ºC min −1 for 2 min and hold 9 min. (Full scan mode) Boor et al., (2015) Baby waterproof fabrics, decorated waterproof tarpaulins, and printed textiles DIBP, DBP, BBP, DEHP, DNOP, DINP, DIDP (7) Detection limit 7µg L −1 DIBP 26 µg L −1 DBP 14 µg L −1 BBP 35 µg L −1 DEHP 43 µg L −1 DNOP 82 µg L −1 DINP 82 µg L −1 DIDP BS EN 15777–2009. Extraction in Soxhlet (4 h) with 50 mL of nhexane. The extract was dried and concentrated to 1 mL and subsequently subjected to determination by GC-MS GC-MS HP-5MS column (30 m × 0.25 mm ×0.25 µm). Nitrogen was the carrier gas at a flow rate of 2.0 mL min −1 . 1 µL sample was injected using the split mode at 250 ºC. The temperature of the GC-MS interface was 280 ºC. The oven temperature program was initiated at 150 ºC for 1 minute, increased to 250 ºC at 8 ºC min −1 , increased to 290 ºC at 3 ºC min −1 , and maintained for six minutes at 290 ºC. (Full scan mode) Li et al., (2015) Jeans DMP, DEP, DIBP, DBP, DEHP (5) Detection limit in the range from 0.10 to 0.52 ng mL −1 for their metabolites Samples were extracted in a Soxhlet extractor (4 h.) with 120 mL dichloromethane. Extracts were concentrated to a GC-MS A fused silica capillary column, HP-5MS (30 m x 0.25 mm ID x 0.25 µm) was used, and He was Gong et al., (2016) (continued on next page) N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 16 were lower (although still high) than those on uncovered locations. DEHP, DBP and DIBP were the most abundant phthalates detected in 80% body locations covered by clothing (DEHP was present at much higher concentrations than other two). The findings demonstrated that clothing positively influences dermal uptake, and one of the final conclusions was that PAE exposure will be underestimated by a factor of 2–5 if absorption through body locations covered by clothing is not considered. DEHP was also the most abundant phthalate in clothing and was present at much higher concentrations than DBP and DIBP. Phthalate levels in jeans and on legs increased with the time that the jeans were worn, as expected, and the increase was highest for DEHP, DBP and DIBP (the phthalates with lower vapour pressures). However, the levels in the jeans and on the legs were not correlated. The authors suggest that this may indicate that other pathways (e.g. contact with bedding or bedclothes) may contribute to increasing the phthalate concentrations on legs. Finally, evaluation of the efficiency of laundering regarding removal of phthalates from jeans showed that the concentrations of DMP, DEP, DIBP and DBP decreased with laundering, while those of DEHP increased in some cases. This finding was attributed by the study authors to contact between the jeans and sources of DEHP during washing and/or drying (such as detergent plastic containers, plastic surfaces of the washing machines, plastic drying racks, as well as the air itself). The efficiency of removal of PAEs from jeans by washing increased with decreasing K OW ; median values ranged from very low (<5%) for DEHP to very high (~75%) for DMP. In this case, the levels detected in the jeans analyzed (both on the calf and thigh areas) in the different experiments were in the ppm range. DEHP was the PAE detected at highest concentrations, of between 0.74 and 690 µg g −1 . More recently, Aldegunde-Louzao et al. (2023) reported the results of a long-term screening study of 1437 samples of adult clothing as well as 1884 fabric samples of diverse origin in southern Europe and northern Africa. Phthalate concentrations were found to be much higher in fabrics than in adult clothing. DEHP was the PAE most commonly detected, in 14.7% of fabric samples (range 10–660 µg g −1 ) and only in 1.32% of clothing samples (range 10–830 µg g −1 ). DINP was the second most common phthalate, with a rate of detection of 7.02% in fabrics (range 10–920 µg g −1 ) and only 0.97% in clothing (range 5–720 µg g −1 ). The other PAEs assayed appeared in only a few cases. Phthalate concentrations were also determined in work clothes with special protective characteristics. In order to evaluate the health risks associated with exposure to phthalate diesters, Alexander and Baxter (2014) analyzed personal protective clothing used by firefighters. Six PAEs were measured according to the EPA 8270E method (US-EPA, 2014) for semi-volatile organic compounds based on GC-MS determination after extraction with methylene chloride. DEHP was detected in all samples of hoods, coats and gloves analyzed. The concentrations ranged from 0.5 µg g −1 in unused cuff gloves to 1400 µg g −1 in the outer surface of used gloves. Another four phthalates (BBP, DEP, DBP and DNOP) were detected in some other samples, at concentrations between two and three orders of magnitude lower. The study authors stated the levels of human exposure cannot be deduced directly from these measurements. However, they suggest that the dermal route may be the major route of exposure to DEHP in firefighters (depending on the length of time the items are worn, skin and swatch temperatures and frequency of cleaning). 2.3.3. Other textile products Phthalates have also been measured in other textiles in contact with the skin, such as skin wipes, face towels and face masks, especially since the COVID-19 pandemic. Gong et al. (2014) determined the levels of six phthalates (DMP, DEP, DIBP, DBP, BBP and DEHP) in skin wipes (8 cm×8 cm gauze pads) passed over the skin on the surface of the forehead, forearm, back-of-hand and palm in 20 study participants (10 male and 10 female). Influencing factors such as body location, time of sampling and handwashing were also considered in order to estimate dermal absorption based on the concentrations measured. The phthalates most frequently detected were DEHP, DIBP and DBP; the concentrations of the first were much higher than those of the other two (in the range 200–1445 µg m −2 on the forehead; 316–2110 µg m −2 on forearms; and 566–12900 µg m −2 on hands). The concentrations were clearly influenced by the body location (palm >back-of-hand >forearm ≥forehead). The estimated total dermal absorption from the skin surface was roughly 10–20% of the total uptake reported for Chinese adults (Guo et al., 2011). This suggests that dermal absorption contributes significantly to the uptake of these phthalates. Washing hands with soap and water removed more than 50% of the phthalates and can thus contribute to decreasing phthalate exposure. The levels reported in this work are expressed relative to the surface area of the gauze pad, and it is therefore difficult to compare the values with others reported in the literature. In a later study, Saini et al. (2016) reported the concentrations of phthalates per surface area to be one order of magnitude lower in cotton and polyester than in other fabrics. Considering the widespread use of face masks since 2020, due to the COVID-19 pandemic, Xie et al. (2022) studied the use of the masks as a potential source of exposure to phthalates in humans. A total of 12 phthalates were measured in 56 face masks of different types produced in several countries (China, Europe, Japan, South Korea, and USA). PAEs were extracted from samples through double ultrasonic-assisted extraction with dichloromethane and ethyl acetate (1:1) and subsequent quantification by GC-MS. Eleven of the twelve phthalates were detected in all the samples analyzed, and DEHP was the most frequently detected PAE in all masks, followed by DBP and DIBP. The total concentration of phthalates ranged from 0.115 to 37.7 µg g −1 . No statistically significant differences were detected between countries, nor between masks designed for adults and for toddlers. After estimation of phthalate exposure and the risk assessment for the use of masks, the authors concluded a hazard index of phthalate exposure from all masks to be an acceptable level of risk. Zhang et al. (2023b) analyzed human exposure to phthalates via the use of face towels. Cut samples were subjected to ultrasonic extraction with n-hexane/acetone, concentrated to almost dryness and made up to approximately 1 mL with n-hexane. The concentrations of 5 PAEs (DMP, DEP, DIBP, DBP and DEHP) in the extracts were then determined by GC-MS. Face towels were found to play a significant role in dermal exposure to phthalates, as PAEs from personal care products remained on the towels. The raw material clearly influences the phthalate content, and coral velvet towels retained significantly higher concentrations of PAEs than cotton towels. Washing with water removed some PAEs; however, when the towels were washed with detergent, the concentrations of PAEs in face towels increased. The levels measured (ranging from up to 11.3 µg g −1 , with DBP the PAE present at highest concentration in towels) were not associated with health risks after long term exposure. 2.4. Evaluation of exposure risks In several of the papers reviewed, the concentrations of the different phthalates in diverse textile products were used to estimate the health risks associated with exposure via dermal absorption of the compounds. Estimation of the dermal pathway is of interest in assessing PAE exposure, as phthalates in contact with skin can cross the viable stratum corneum/epidermis and enter the dermal capillaries via surface lipids (Gong et al., 2014). There seems to be some consensus regarding the models used to estimate risk through this route. The estimated daily uptake (EDI, expressed in µg/kg-bw/day) of the different phthalates from clothing through dermal absorption can be divided into two parts: through the clothing in direct contact with the skin surface (EDI da-cont ), and through the gap between clothing and the skin surface (EDI da-ag ) (Li et al., 2019). EDIda =EDIda−cont +EDIda−ag (1) N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 17 Both EDI da-cont , and EDI da-ag can be calculated as follows: EDIda−cont =Ccl⋅Kssl−g⋅Kp−l⋅SA⋅EF Kcl−g⋅BW ⋅f(2) and EDIda−ag =Ccl⋅Kp−g⋅SA⋅EF Kcl−g⋅BW ⋅(1−f)(3) where C cl is the phthalate concentrations in clothing, K p-l is the skin permeability coefficient from skin surface lipids to dermal capillaries, SA is skin contact area via clothing; EF is the fraction of exposure duration; BW is body weight, and f is the fraction of the skin contact area. K is the partition coefficient between two phases and the subscripts cl-ssl, ssl-g and cl-g represent the phases between clothing and skin surface lipids, between skin surface lipids and gas phase (air), and between clothing and gas phase (air), respectively; and K p-g is the indoor air transdermal permeability coefficient. The values and the assumptions used in the calculation can be consulted in detail in the Supporting Information of the paper by Li et al. (2019). In several cases, the EDI via ingestion (EDI ing ) for infant from infant clothing was also considered, and calculated as follows: EDIing =Ccl⋅MA⋅EF⋅ERsal BW (4) where MA is the mouthing area, and ER sal is the eluted rates of phthalates in cotton clothing into saliva per day. Therefore, the total EDI is calculated as follows: EDItotal =EDIda +EDIing (5) For evaluation of health risks, the cumulative risk assessment was generally estimated from the hazard index (HI), which is the sum of individual hazard quotient (HQ) for each phthalate considered (US-EPA, 2004). Therefore, for each phthalate i determined, HQi=(EDIdai Reference limit valuedai +EDIingi Reference limit valueingi)⋅10−3(6) and HI, when n phthalates have been determined, can be estimated as follows: HI =∑n i=1HQi(7) Finally, in several cases, the carcinogenic risk for children from clothing was calculated according to the following equation (US-EPA, 2004, 2011): CR =(EDIda⋅SFda⋅EDIing⋅SFing)⋅ASF⋅10−6(8) where CR is the probability of developing cancer as a result of phthalate exposure, and SF is the slope factor for dermal absorption and ingestion (which can be calculated from the slope factor for oral exposure, US-EPA, 2004; Li et al., 2018; Li et al., 2019), and ASF is an age sensitivity factor. The results for exposure risk, in terms of estimated daily intake (EDI) and cumulative risk (CR and HI), reported in the papers in which these indicators were evaluated are summarized in Table 5. Chronologically, the first attempt to assess exposure was made by Alexander and Baxter (2014), who considered the levels of several phthalates in the personal protective equipment of firefighters. These researchers concluded that human exposure cannot be inferred from the PAE concentrations in personal protective equipment garments, because of the involvement of other interacting factors that can modify the effective internal dose in target organs. These factors include skin temperature and washing frequency, time the garments are worn, garment cleaning and heart rates. The researchers finally concluded that dermal absorption may also be a major route of exposure in firefighters. However, the exposure risk should be the objective of further, more detailed studies. The first study in which the PAE transport between skin surface and skin capillaries was estimated is the pioneering research conducted by (Gong et al., 2014). In this study, the amount of PAEs adsorbed transdermally from the skin surface lipids (DA) was estimated through the flux from the skin surface lipids to the dermal capillaries (J): J=Kpl⋅Cl(9) DA =J⋅texp⋅A BW (10) where Kpl is the permeability coefficient from skin surface lipids to blood and Cl is the concentration of phthalate in contact with skin surface, texp is the time of exposure, A is the exposed skin area, and BW is the body weight. According to this model, levels of dermal absorption of DIBP, DBP and DEHP were 0.48, 0.68 and 0.66 µg/kg-bw/day, respectively. These levels are similar, but slightly higher, than those reported for dermal absorption from other sources (Guo and Kannan, 2011; Wang et al., 2014). However, the authors recognized several limitations in this study, including the small sample size, the limited number of body locations evaluated and various calculations that would underestimate the dermal absorption of phthalates. Li et al. (2019) applied a more sophisticated model based on Eqs. (1) to (7) to estimate the daily intake and the health risk derived from infant exposure to PAEs through cotton clothing. The total EDI of phthalates, considering both dermal absorption and ingestion, ranged from to 9810 ng/kg-bw/day, with a median value of 776 ng/kg-bw/day. As can be seen from the individual median values for each PAE in Table 5, DBP contributed most to exposure in infants. Some interesting results were obtained on comparing the two exposure routes for dermal absorption. The EDIda−cont values of phthalates through direct contact with the skin surface were approximately 6 times higher than those corresponding to the gap between skin surface and clothing EDIda−ag. The contribution of ingestion EDIing was negligible relative to the other two routes, reaching median values ranging from 1.48 ×10 −5 to 3.82 ×10 −3 ng/kg-bw/day, i. e., 4–6 orders of magnitude lower. These results were also used to calculate the cumulative risk for exposure to DIBP, DBP and DEHP. Median HI values (values below 1 were considered safe, according to Benson, 2009) based on three different reference limit values ranging from 0.060 to 0.0091, suggesting that PAE exposure from cotton clothing for infants is within acceptable levels. The carcinogenic risk was also evaluated according to Eq. (8) for BBP and DEHP. Calculation of the total CR for these two PAEs in infant clothing yielded values in the range 1.57 ×10 −6 to 7.02 ×10 −5 , with a median value of 5.60 ×10 −6 . DEHP contributed the most (CR values 3 orders of magnitude greater). These values exceeded the safe threshold (potentially adverse effects must be considered when CR values are higher than 1 ×10 −6 , according to US-EPA, 2018) and indicate that exposure to DEHP could have serious effects on children’s health. Tang et al. (2020) studied dermal exposure to 15 phthalates in relation to children’s clothing manufactured in several countries in Asia and using a similar model. DMP, DEP, DBP, BBP, DEHP and DNOP accounted for most of the total concentration of all of the 15 PAEs determined. Thus, dermal exposure to these six phthalates was considered. The results of EDI for the joint exposure through different items of clothing (the median value for each clothing item was considered) in the two scenarios selected according to the type and number of items studied are summarized in Table 5. The results were similar to those reported by Li et al. (2018). Moreover, the importance of the exposure route was also confirmed in this study, as EDIing reached values 3–5 orders of magnitude lower than those associated with dermal uptake. The summed HQ of DBP and DEHP for reproductive risk exceeded the acceptable level (1.0). However, the HQ of DEHP for carcinogenic risk indicated a low cancer risk. Face masks were recently evaluated as a potential source of N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 18 Table 5 Phthalate exposure and health risk evaluation from clothing. Sample Type (nº of samples) Phthalates considered (n) Estimated daily intake (EDI) Cumulative risk (HI and CR) Notes Reference Firefighter personal protective clothing (6) BBP, DEHP, DEP, DBP, DNOP (5) Not determined Not determined Authors stated that human exposure cannot be inferred from the PAE concentrations because other interacting factors (such as skin temperature, time of wear, heart rate, frequency of cleaning and skin washing) can modify the effective internal dose in target organs. Alexander and Baxter, (2014) Skin wipes (20) DMP, DEP, DIBP, DBP, BBP, DEHP (6) (Dermal absorption) DIBP 0.48 μ g/kgbw/day DBP 0.68 μ g/kgbw/day DEHP 0.66 μ g/kgbw/day Not determined Calculation of dermal absorption (DA) has been made according to Eq. (9) Gong et al., (2014) Infant’s cotton clothing (24) DMP, DEP, DIBP, DBP, BBP, DEHP (6) (Median EDI) DMP 29 ng/kg-bw/ day DEP 20 ng/kg-bw/ day DIBP 256 ng/kgbw/day DBP 296 ng/kgbw/day BBP 0.41 ng/kgbw/day DEHP 66.5 ng/kgbw/day Median HI (DIBP, DBP and DEHP) ranged from 0.0066 to 0.060 according to three different reference limits (EFSA, EPA, and R f D AA) CR (BBP and DEHP) ranged from 1.57×10 −6 to 7.02×10 −5 (Median: 5.60×10 −6 ) Calculation of estimated daily intake (EDI) and cumulative risk as hazard index (HI) and carcinogenic risk (CR) has been made according to the model described in Eqs. (1) to (8). Li et al., (2019) Preschool children’s clothing (67) DMP, DEP, DIBP, DBP, DMEP, BMPP, DEEP, DPP, DHXP, BBP, DBEP, DEHP, DIDP, DNOP, DNP (15) (Median EDI) DMP 25.0–41.6 ng/kgbw/day DEP 61.2–101 ng/ kg-bw/day DIBP 249–428 ng/ kg-bw/day DBP 123–260 ng/ kg-bw/day BBP 1.17–0.991 ng/kgbw/day DEHP 80.3–119 ng/kgbw/day HQ (DBP and DEHP) by item type were in the range of 0.331–7.89 and 0.001–0.076, respectively. CR (DEHP) in the two scenarios ranged from 0.054 to 0.080. Estimated daily intake (EDI) and cumulative risk as hazard index (HI) and carcinogenic risk (CR) were calculated according to the model described in Eqs. (1) to (8). Tang et al., (2020) Face masks (56) DMP, DEP, DIBP, DBP, DMEP, DPP, DHXP, DCHP, DEHP, DPHP, DNOP, DNP (12) (Median EDI) DMP 0.245 ng/kgbw/day DEP 0.165 ng/kgbw/day DIBP 5.29 ng/kgbw/day DBP 5.20 ng/kgbw/day DMEP 0.0 ng/kgbw/day DPP 0.580 ng/kgbw/day DHXP 0.0 ng/kgbw/day DEHP 9.66 ng/kgbw/day DPHP 1.19 ng/kgbw/day DNOP 0.045 ng/ kg-bw/day DNP 0.485 ng/kgbw/day HI in the range 1.33 ×10 −4 to 3.17 ×10 −2 (Median: 9.5 ×10 −4 ) CR in the range 6.29 ×10 −8 to 4.26 ×10 −5 (Median: 2.77 ×10 −6 ) Estimated daily intake (EDI) and cumulative risk as hazard index (HI) and carcinogenic risk (CR) were calculated according to the model described in Eqs. (1) to (8). Xie et al., (2022) Face towels (31) DMP, DEP, DIBP, DBP, DEHP (5) DMP<1 ×10 −4 ng/ kg-bw/day DEP<1 ×10 −4 ng/ kg-bw/day DIBP up to 2 ×10 −4 ng/kg-bw/day DBP up to 3 ×10 −4 ng/kg-bw/day Median HI =7.33 ×10 −9 (Maximum 6.74 ×10 −8 ) DCR (DEHP) =2.26 ×10 −12 (Maximum 1.22 ×10 −11 ) Estimated daily intake (EDI) and cumulative risk as hazard index (HI) and carcinogenic risk (CR) were calculated according to the model described in Eqs. (1) to (8). Zhang et al., (2023b) (continued on next page) N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 19 phthalate exposure (Xie et al., 2022). Twelve PAEs were considered in this study, and the EDIs of phthalates for 56 masks of different types ranged from 3.71 to 639 ng/kg-bw/day, with a median value of 33.9 ng/kg-bw/day. DEHP was the most abundant PAE regarding phthalate exposure (mean 39.1%), followed by DIBP (19.2%) and DBP (16.2%). Considering the maximum EDI values for the five most abundant PAEs, these accounted for respectively 27.9%, 1.10%, 13.6%, 61.9% and 81.9% of the total phthalate exposure. Thus, face masks can be considered a non-negligible source of phthalate exposure. In fact, estimation of the risk of phthalate exposure showed that the median hazard index (HI) values for adults and toddlers were respectively 7.08 ×10 −4 and 3.92 ×10 −3 , well below acceptable levels in terms of carcinogenic risks. However, the cumulative CR values of phthalate exposure for adults and toddlers ranged from 5.30 ×10 −7 to 1.45 10 −5 and from 6.29 ×10 −8 to 4.26 ×10 −5 . Cumulative CR values for 50 mask samples (accounting for 89.3% of the mask samples) were higher than 1 ×10 −6 (the reference maximum level). Thus, the potential carcinogenic effects of phthalates from masks should not be ignored. Finally, Zhang et al. (2023b) investigated PAE exposure in 24 volunteers in relation to the use of face towels, made of cotton or coral velvet, by calculating EDI, HQ, HI and dermal cancer risk (DCR). The EDIs were in all cases below 8.0 ×10 −4 ng/kg-bw/day, and the values were affected by diverse factors, including the gender of volunteers (male or female), weight and towel material. EDI values for several PAEs were significantly (p<0.05) higher in females. This was attributed by the authors to a greater variety and more frequent use of personal care products in females. In addition, the lowest mean EDI according to the weight was observed in volunteers weighing more than 65 kg, which was explained by considering the greater dilution effect on the absorption of phthalates in heavier people (See US-EPA, 2011b). As expected, the material the towels are made from was also an influencing factor, and the EDIs of PAEs were found to be significantly higher (p<0.05) in coral velvet than in cotton towels. In fact, EDI values for coral velvet ranged from 4to 14-fold higher than those for cotton, because PAEs adhere more strongly to coral velvet than to cotton. The hazard index method was used to estimate simultaneous exposure to PAEs via face towels. Logically, HI and HQ values present the same influencing factors as EDIs. The median HI for PAEs, 7.33 ×10 −9 (maximum 6.74 ×10 −8 ), was much lower than the safe value of HI=1. The DCR values are in the range 10 −11 -10 −12 , also much lower than the value of 10 −6 regarded as safe. The study authors thus established that the carcinogenic risk associated with the use of face towels was low. 3. Concluding remarks Phthalates are a family of chemical compounds commonly used as plasticizers in the manufacture of some textiles and clothing. They are known to have potentially adverse effects on human health (including endocrine disruption and reproductive toxicity). Evaluation and control of phthalates in clothing is thus of interest to assess the potential risks. This type of analysis also helps to identify specific clothing items that may contain high concentrations of phthalates and encourages manufacturers to reduce or eliminate these phthalates from their products by promoting production of safer, more sustainable textiles. This review paper, based on the findings reported in more than 120 papers, considers the use of phthalates as plasticizers in the textile industry and the potential adverse effects of these compounds on health. The legislation concerning phthalates, the analytical methodologies used to extract and determine PAEs, and analysis of studies evaluating the PAE contents in clothing -and their associated risksare also critically discussed. According to the information reported in papers published in the last decade, the following general conclusions can be reached: (i) Only some tens of studies have determined phthalates in clothing in the last decade. However, in the same period (see Section 1.3) the dermal route has been shown to be an important route of exposure to PAEs. Further research in this field is obviously required to demonstrate the potential effects on human health of exposure to phthalates through this pathway. (ii) Variable numbers and types of PAEs were analyzed in the different papers reviewed. In most studies, all or some of 6 phthalates that have been regulated in the different countries (DBP, BBP, DEHP, DNOP, DINP and DIDP) were analyzed (Alexander and Baxter, 2014; Gong et al., 2014, 2016; Li et al., 2015; Saini et al., 2016; Negev et al., 2018; Li et al., 2019; Aldegunde-Louzao et al., 2023; Zhang et al., 2023b, among others). In other cases, larger sets of PAEs were assayed when the characteristics of the tested material indicated that this was advisable. Thus, Boor et al. (2015) measured other phthalates (DINCH, DEHA, DMP and DBP), in addition to the those already mentioned, in infant crib mattress covers and foam. Xie et al. (2016) determined ATBC (tributyl o-acetylcitrate), DEHA, DEHT and DINCH, in addition to DBP, DIBP, DEHP and DINP, in different parts of children’s backpacks. Xie et al. (2022) determined the levels of 12 PAEs in face masks (DMP, DEP, DIBP, DBP, DMEP, DPP, DHXP, DCHP, DEHP, DPHP, DNOP and DNP). Finally, Tang et al., (2020) determined the levels of 15 different phthalates (DMP, DEP, DIBP, DBP, DMEP, BMPP, DEEP, DPP, DHXP, BBP, DBEP, DEHP, DIDP, DNOP and DNP) in diverse items of preschool children’s clothing produced in several Asian countries. A large set of PAEs should be established for future screening of clothing, as due to the legal control of several phthalates in most countries, textile manufacturers are replacing the prohibited phthalates with other non-regulated phthalates in which the R and R` radicals are modified. Moreover, in dermal exposure, phthalates pass the lipid barrier of the skin surface and thus enter the dermal capillaries (Gong et al., 2014). Flux from the clothing through the gap between clothing and skin to the lipids in the skin surface is directly related to the K OW (n-octanol/water partition coefficient) of each PAE. As K OW is a measure of the relationship between lipophilicity and hydrophilicity of a substance (Ellington, 1999), absorption of phthalates with high values of this partition coefficient (high lipid solubility) is greater because these compounds can cross the viable stratum corneum/epidermis composite and reach the dermal capillaries more easily. On the other hand, the high K OW phthalates are also more difficult to remove by washing, given their low solubility in water. Thus, K OW can also be a useful tool for selecting which PAEs should be measured. Thus, future screening studies should include an extended range of PAEs, including at least those considered by Tang et al. (2020). (iii) The concentrations of PAEs detected in the different studies are very variable. In general, in most of the papers reviewed, the concentrations of almost all PAEs reported were in the µg g −1 range (ppm) (e.g. in Alexander and Baxter, 2014; Gong et al., 2016; Li et al., 2019; Aldegunde-Louzao et al., 2023 and Zhang Table 5 (continued) Sample Type (nº of samples) Phthalates considered (n) Estimated daily intake (EDI) Cumulative risk (HI and CR) Notes Reference DEHP up to 8 ×10 −4 ng/kg-bw/ day N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 20 et al., 2023b). However, in several studies, the concentrations detected were three orders of magnitude higher, in the mg g −1 range, e.g. in the studies by Boor et al. (2015) and Negev et al. (2018) concerning foam and mattress covers for children’s cribs, and other childcare items. The waterproof characteristics of these products may explain the high levels of PAEs detected. Similar results were obtained (in the same mg g −1 range) by Xie et al. (2016) for different parts of children’s backpacks, also made from waterproof textile materials. Xie et al. (2022) reported lower levels in face masks, with concentrations of several PAEs in the range of ng g −1 (ppb). In addition, in studies analyzing both children’s and adult clothing simultaneously (Aldegunde-Louzao et al., 2023), the detection rates for the different phthalates were lower in children’s clothing than in clothing intended for adults, and in other textile products such as fabrics, or other related materials. (iv) The number of samples analyzed in the studies is generally very small. Only 70 % of the papers reviewed analyzed more than 50 samples, and in several papers fewer than 10 samples were analyzed. This can be explained by the special characteristics of some samples (e.g. fireproof clothing, foam and mattress covers), which reduces the availability of samples. However, the small number of samples impedes generalization of the results. Finally, as the geographic origin of clothing samples can influence the phthalate content (due to the different legislation), studies should be carried out in several different countries. However, almost half of the studies on clothing are carried out on garments originating in China and/or various countries in Southeast Asia. Another 30% are carried out on samples from the USA, and the others on samples from Israel and some European countries. Consequently, larger studies (both in terms of number of samples and their temporal and spatial distribution) should be conducted separately for the different types of samples, for children’s and adult clothing, as well as fabrics and other materials. (v) DEHP was the PAE detected at the highest concentration in most studies (see Table 4) This finding is consistent with data showing that DEHP is the PAE consumed in greatest quantities by the industry in general, and by the textile industry in particular (Wang and Qian, 2021). In addition, DEHP is also the PAE associated with the highest EDIs values and health risks (Gong et al., 2014; Li et al., 2019; Tang et al., 2020; Xie et al., 2022; Zhang et al., 2023b). However, the gradual replacement of DEHP by other phthalates such as DINP, DIDP and DPHP or DEHT, noted by other authors (Kataria et al., 2017), can be observed in the fact that DINP was present at higher concentrations than DEHP in several studies involving crib mattresses (Boor et al., 2015), printed textiles (Li et al., 2015), textile fabrics (Aldegunde-Louzao et al., 2023) and childcare items (Negev et al., 2018). DEHT was also present at higher concentrations than DEHP in children’s backpacks (Xie et al., 2016) confirming the replacement of DEHP by other phthalates or other plasticizers (Qadeer et al., 2022). This indicates the need to develop analytical methods capable of detecting all phthalates, irrespective of the type of R and R ′ chains included. For this purpose, GC-MS/MS and LC-MS/MS procedures based on precursor ion scanning may be suitable options. This operational mode enables identification of all of the precursor ions that can generate a given ion product, and all phthalates would thus be detected in the corresponding analysis, regardless of the R and R ′ chains they contain. Use of these analytical methodologies and implementation of the appropriate legislative regulations may contribute to better control of phthalates in textile products in the future. (vi) Very few of the studies used the phthalate concentrations to estimate the health risks associated with PAEs in clothing (see Table 5). After the study by Gong et al. (2014), in which the dermal absorption was first estimated, Li et al. (2019) proposed an exposure model based on estimated daily intake (EDI) and cumulative risk assessment by hazard index (HI) and hazard quotient (HQ). This model has been applied by other authors to other textile products (Xie et al., 2022; Zhang et al., 2023b). It thus seems that here too there is some consensus regarding methods of estimating exposure to PAEs and quantifying the associated risks. However, this model assumes that phthalate partitioning between clothing/gas-phase air and between skin surface lipids/gas-phase air reaches a state of equilibrium. This can influence the EDIs obtained (mainly for higher molecular weight phthalates, with longer equilibrium times). Furthermore, due to the lack of available data, the values of the different coefficients (such as Kp−l,Kcl−g,and Kssl−g) were estimated under several assumptions or derived from other research, which may result in overor underestimation of the EDIs. On the other hand, comparison of the results of the different studies is not possible as EDIs and cumulative risks were calculated for different numbers and types of PAEs. Further research is required to establish values for these partition coefficients. Moreover, further research involving estimation of exposure and health risks due to the presence of PAEs in clothing is also required. However, such research should be conducted according to an agreed protocol, i. e. by determining and considering the same group of PAEs, in similar types of clothing and in different countries, in order to produce results that can be validly compared. (vii) The ubiquitous use of synthetic PAEs has created major concerns about the risks of these compounds for both human and environmental health. The replacement of synthetic phthalates with natural compounds or different natural-based plasticizers, less dangerous and obtained from living organisms, is a promising line of research that could contribute to reducing the unwanted effects (Adeodato Vieira et al., 2011; Roy, 2020). Information on funding sources This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sectors. CRediT authorship contribution statement Natalia Aldegunde-Louzao: Conceptualization, Writing – original draft, Writing – review & editing, Investigation, Methodology. Manuel Lolo-Aira: Writing – original draft, Writing – review & editing, Resources. Carlos Herrero-Latorre: Conceptualization, Writing – original draft, Writing – review & editing, Formal analysis, Project administration, Supervision. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Data Availability This is a review paper, all data in this paper are available from references used References Adeodato Vieira, M.G., Altenhofen da Silva, M., Oliveira dos Santos, L., Masumi Beppu, M., 2011. Natural-based plasticizers and biopolymer films: A review. Eur. Polym. J. 47, 254–263. https://doi.org/10.1016/j.eurpolymj.2010.12.011. Ahern, T.P., Broe, A., Lash, T.L., Cronin-Fenton, D.P., Ulrichsen, S.P., Christiansen, P.M., Cole, B.F., Tamimi, R.M., Sørensen, H.T., Damkier, P., 2019. Phthalate exposure and breast cancer incidence: a Danish nationwide cohort study. J. Clin. Oncol. 37, 1800–1809. https://doi.org/10.1200/JCO.18.02202. N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 21 Aldegunde-Louzao, N., P´ erez L´ opez, P., Lolo Aira, M., Herrero-Latorre, C., 2023. Sevenyear-long screening of phthalate esters in clothing and textile products from a quality control laboratory. Text. Res. J. 93, 1670–1685. https://doi.org/10.1177/ 00405175221135619. Alexander, B.M., Baxter, C.S., 2014. Plasticizer contamination of firefighter personal protective clothing – A potential factor in increased health risks in firefighters. J. Occup. Environ. Hyg. 11, D43–D48. https://doi.org/10.1080/ 15459624.2013.877142. Alp, A.C., Yerlikaya, P., 2020. Phthalate ester migration into food: effect of packaging material and time. Eur. Food Res. Technol. 246 (3), 425–435. https://doi.org/ 10.1007/s00217-019-03412-y. Amritha, P.S., Vinod, V., Harathi, P.B., 2022. A critical review on extraction and analytical methods of phthalates in water and beverages. J. Chromatogr. A 1675, 463175. https://doi.org/10.1016/j.chroma.2022.463175. Attina, T.M., Trasande, L., 2015. Association of exposure to di-2-Ethylhexylphthalate replacements with increased insulin resistance in adolescents from NHANES 2009–2012. J. Clin. Endocrinol. Metab. 100, 2640–2650. https://doi.org/10.1210/ jc.2015-1686. Benson, R., 2009. Hazard to the developing male reproductive system from cumulative exposure to phthalate esters-dibutyl phthalate, diisobutyl phthalate, butylbenzyl phthalate, diethylhexyl phthalate, dipentyl phthalate, and diisononyl phthalate. Regul. Toxicol. Pharmacol. 53, 90–101. https://doi.org/10.1016/j. yrtph.2008.11.005. Boberg, J., Christiansen, S., Axelstad, M., Kledal, T.S., Vinggaard, A.M., Dalgaard, M., Nellemann, C., Hass, U., 2011. Reproductive and behavioral effects of diisononyl phthalate (DINP) in perinatally exposed rats. Reprod. Toxicol. 31, 200–209. https:// doi.org/10.1016/j.reprotox.2010.11.001. Bølling, A.K., Sripada, K., Becher, R., Bek¨ o, G., 2020. Phthalate exposure and allergic diseases: Review of epidemiological and experimental evidence. Environ. Int. 139, 105706 https://doi.org/10.1016/j.envint.2020.105706. Boor, B.E., Liang, Y., Crain, N.E., Ja¨rnstro¨m, H., Novoselac, A., Xu, Y., 2015. Identification of phthalate and alternative plasticizers, flame retardants, and unreacted isocyanates in infant crib mattress covers and foam. Environ. Sci. Technol. Lett. 2, 89–94. https://doi.org/10.1021/acs.estlett.5b00039. Brassea-P´ erez, E., Hern´ andez-Camacho, C.J., Labrada-Martag´ on, V., V´ azquez-Medina, J. P., Gaxiola-Robles, R., Zenteno-Savín, T., 2011. Oxidative stress induced by phthalates in mammals: State of the art and potential biomarkers. Environ. Res. 206, 112636 https://doi.org/10.1016/j.envres.2021.112636. Bu, Z., Dong, C., Mmereki, D., Ye, Y., Cheng, Z., 2021. Modeled exposure to phthalates via inhalation and dermal pathway in children’s sleeping environment: A preliminary study and its implications. Build. Simul. 14, 1785–1794. https://doi. org/10.1007/s12273-021-0769-8. Bu, Z., Wang, J., Yu, W., Li, B., 2018. Dermal exposure to phthalates in home environment: Handwipes, influencing factors and implications. Build. Environ.133 1–7. https://doi.org/10.1016/j.buildenv.2018.02.013. Bu, Z., Zhang, Y., Mmereki, D., Yu, W., Li, B., 2016. Indoor phthalate concentration in residential apartments in Chongqing, China: Implications for preschool children’s exposure and risk assessment. Atmos. Environ. 127, 34–45. https://doi.org/ 10.1016/j.atmosenv.2015.12.010. Chang, W.H., Herianto, S., Lee, C.C., Hung, H., Chen, H.L., 2021. The effects of phthalate ester exposure on human health: A review. Sci. Total Environ. 786, 147371 https:// doi.org/10.1016/j.scitotenv.2021.147371. Chen, C.Y., 2004. Biosynthesis of di-(2-ethylhexyl) phthalate (DEHP) and di-n-butyl phthalate (DBP) from red alga-Bangia atropurpurea. Water Res 38, 1014–1018. https://doi.org/10.1016/j.watres.2003.11.029. Chen, Y., Wang, Z., Liu, S., Zhao, G., 2021. A highly sensitive and group-targeting aptasensor for total phthalate determination in the environment. J. Hazard. Mater. 412, 125174 https://doi.org/10.1016/j.jhazmat.2021.125174. Crobeddu, B., Ferraris, E., Kolasa, E., Plante, I., 2019. Di(2-ethylhexyl) phthalate (DEHP) increases proliferation of epithelial breast cancer cells through progesterone receptor dysregulation. Environ. Res. 173, 165173 https://doi.org/10.1016/j. envres.2019.03.037. Czubacka, E., Czerczak, S., Kupczewska-Dobecka, M., 2021. The overview of current evidence on the reproductive toxicity of dibutyl phthalate. Int. J. Occup. Med. Environ. Health 34, 15–37. https://doi.org/10.13075/ijomeh.1896.01658. Das, A.K., Gogia, A., Garg, M., Elaiyaraja, A., Arambam, P., Mathur, S., BabuRajendran, R., Deo, S.V.S., Kumar, L., Das, B.C., Janardhanan, R., 2022. Urinary concentration of endocrine-disrupting phthalates and breast cancer risk in Indian women: A case-control study with a focus on mutations in phthalate-responsive genes. Cancer Epidemiol. 79, 102188 https://doi.org/10.1016/j. canep.2022.102188. Dirinck, E., Dirtu, A.C., Geens, T., Covaci, A., Van Gaal, L., Jorens, P.G., 2015. Urinary phthalate metabolites are associated with insulin resistance in obese subjects. Environ. Res. 137, 419–423. https://doi.org/10.1016/j.envres.2015.01.010. Dodson, R.E., Nishioka, M., Standley, L.J., Perovich, L.J., Brody, J.G., Rudel, R.A., 2012. Endocrine disruptors and asthma-associated chemicals in consumer products. Environ. Health Perspect. 120, 935–943. https://doi.org/10.1289/ehp.1104052. Domínguez, F., 2019. Phthalates and other endocrine-disrupting chemicals: The 21st century’s plague for reproductive health. Fertil. Steril. 111, 885–886. https://doi. org/10.1016/j.fertnstert.2019.01.029. Dong, Y., Chen, L., Gao, D., Li, Y., Chen, M., Ma, T., Ma, Y., Liu, J., Zhang, Y., Ma, Q., Wang, X., Song, Y., Zou, Z., Ma, J., 2022. Endogenous sex hormones homeostasis disruption combined with exogenous phthalates exposure increase the risks of childhood high blood pressure: A cohort study in China. Environ. Int. 168, 107462 https://doi.org/10.1016/j.envint.2022.107462. Doolittle, K.A., 1954. The Technology of Solvents and Plasticizers. John Wiley & Sons, New York. Duty, S.M., Silva, J., Barr, D., Brock, J.W., Ryan, L., Chen, Z., Herrick, R.F., Christiani, D. C., Hauser, R., 2003. Phthalate exposure and human semen parameters. Epidemiol 14, 269–277. 〈https://www.jstor.org/stable/3703846〉. Edwards, L., McCray, N.L., Van Noy, B.N., Yau, A., Geller, R.J., Adamkiewicz, G., Zota, A. R., 2022. Phthalate and novel plasticizer concentrations in food items from U.S. fast food chains: A preliminary analysis. J. Expo. Sci. Environ. Epidemiol. 32, 366–373. https://doi.org/10.1038/s41370-021-00392-8. Ellington, J.J., 1999. Octanol/Water partition coefficients and water solubilities of phthalate esters. J. Chem. Eng. Data 44, 1414–1418. https://doi.org/10.1021/ je990149u. European Parliament and the Council., 2006. Regulation (EC) No 1907/2006 of the European Parliament and of the Council concerning the Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH), establishing a European Chemicals Agency, amending Directive 1999/45/EC and repealing Council Regulation (EEC) No 793/93 and Commission Regulation (EC) No 1488/94 as well as Council Directive 76/769/EEC and Commission Directives 91/155/EEC, 93/67/ EEC, 93/105/EC and 2000/21/EC. Off. J. Eur. Union L396: 30.12.2006. 〈https://eur -lex.europa.eu/legal-content/EN/TXT/?uri=OJ:L:2006:396:TOC〉. European Union Commission., 2009. Regulation EU No 552/2009 of 22 June 2009 amending Regulation (EC) No 1907/2006 of the European Parliament and of the Council on the Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) as regards Annex XVII. Off. J. Eur. Union L164/7: 26.6.2009. 〈https://eur -lex.europa.eu/legal-content/EN/TXT/?uri=OJ:L:2009:164:TOC〉. European Union Commission, 2015. Regulation EU No 2015/326 of 2 March 2015 amending Annex XVII to Regulation (EC) No 1907/2006 of the European Parliament and of the Council on the Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) as regards polycyclic aromatic hydrocarbons and phthalates. Off. J. Eur. Union L58/43: 3.3.2015. 〈https://eur-lex.europa.eu/legal-content /EN/TXT/?uri=OJ:L:2015:058:TOC〉. European Union Commission, 2018. Regulation EU No 2018/1513 of 10 October 2018 amending Annex XVII to Regulation (EC) No 1907/2006 of the European Parliament and of the Council concerning the Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) as regards certain substances classified as carcinogenic, mutagenic or toxic for reproduction (CMR), category 1A or 1B. Off. J. Eur. Union L256/1: 12.10.2018. 〈https://eur-lex.europa.eu/legal-content/EN/TXT/? uri=OJ:L:2018:256:TOC〉. Farahani, H., Ganjali, M.R., Dinarvand, R., Norouzi, P., 2008. Screening method for phthalate esters in water using liquid-phase microextraction based on the solidification of a floating organic micro drop combined with gas chromatography– mass spectrometry. Talanta 76, 718–723. https://doi.org/10.1016/j. talanta.2008.03.002. Fern´ andez-Gonz´ alez, V., Moscoso-P´ erez, C., Muniategui-Lorenzo, S., L´ opez-Mahía, P., Prada-Rodríguez, D., 2017. Reliable, rapid and simple method for the analysis of phthalates in sediments by ultrasonic solvent extraction followed by head spacesolid phase microextraction gas chromatography mass spectrometry determination. Talanta 162, 648–653. https://doi.org/10.1016/j.talanta.2016.10.068. Fromme, H., Gruber, L., Schlummer, M., Wolz, G., Bohmer, S., Angerer, J., Mayer, R., Liebl, B., Bolte, G., 2007. Intake of phthalates and di(2-ethylhexyl)adipate: Results of the integrated Exposure Assessment Survey based on duplicate diet samples and biomonitoring data. Environ. Int. 33, 1012–1020. https://doi.org/10.1016/j. envint.2007.05.006. Giuliani, A., Zuccarini, M., Cichelli, A., Khan, H., Reale, M., 2020. Critical review on the presence of phthalates in food and evidence of their biological impact. Int. J. Environ. Res. Public Health 17, 5655. https://doi.org/10.3390/ijerph17165655. Gong, M., Weschler, C.J., Zhang, Y., 2016. Impact of clothing on dermal exposure to phthalates: Observations and insights from sampling both skin and clothing, 2016 Environ. Sci. Technol. 50, 4350–4357. https://doi.org/10.1021/acs.est.6b00113. Gong, M.Y., Zhang, Y.P., Weschler, C.J., 2014. Measurement of phthalates in skin wipes: Estimating exposure from dermal absorption. Environ. Sci. Technol. 48, 7428–7435. https://doi.org/10.1021/es501700u. Gonz´ alez-S´ alamo, J., Hern´ andez-Borges, J., Afonso, M.M., Rodríguez-Delgado, M.A., 2018. Determination of phthalates in beverages using multiwalled carbon nanotubes dispersive solid-phase extraction before HPLC-MS. J. Sep. Sci. 41, 2613–2622. https://doi.org/10.1002/jssc.201800192. Graham, P.R., 1973. Phthalate esters plasticizers-Why and how they are used. Environ. Health Perspect. 3, 3–12. 〈https://ehp.niehs.nih.gov/doi/pdf/10.1289/ehp.73033〉. Guo, Y., Kannan, K., 2011. Comparative assessment of human exposure to phthalate esters from house dust in China and the United States. Environ. Sci. Technol. 45, 3788–3794. https://doi.org/10.1021/es2002106. Guo, Y., Wu, Q., Kannan, K., 2011. Phthalate metabolites in urine from China, and implications for human exposures. Environ. Int. 37, 893–898. https://doi.org/ 10.1016/j.envint.2011.03.005. Guo, Y., Zhang, Z., Liu, L., Li, Y., Ren, N., Kannan, K., 2012. occurrence and profiles of phthalates in foodstuffs from China and their implications for human exposure. J. Agr. Food Chem. 60 (27), 6913–6919. https://doi.org/10.1021/jf3021128. Hahladakis, J.N., Velis, C.A., Weber, R., Iacovidou, E., Purnell, P., 2018. An overview of chemical additives present in plastics: Migration, release, fate and environmental impact during their use, disposal and recycling. J. Hazard. Mat. 344, 179–199. https://doi.org/10.1016/j.jhazmat.2017.10.014. Herbert, C., De Andrade Lima, L., Gonçalves, C., 2017. Alternative to phthalate plasticizer for PVC/NBR formulation used in automotive fuel system with biodiesel, SAE Technical Paper. SAE Int. https://doi.org/10.4271/2017-01-0482. N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 22 Hidalgo-Serrano, M., Borrull, F., Marc´ e, R.M., Pocurull, E., 2022. Phthalate esters in marine ecosystems: Analytical methods, occurrence and distribution. TrAC, Trends Anal. Chem. 151, 116598 https://doi.org/10.1016/j.trac.2022.116598. Hu, H., Li, T., Bao, J., Zhang, X., Sun, X., Xu, K., Liu, Q., Guo, Y., 2022. Determination of phthalates in marine sediments using ultrasonic extraction followed by dispersive solid-phase extraction and gas chromatography-mass spectrometry. J. Chromatogr. Sci. 60, 207–216. https://doi.org/10.1093/chromsci/bmab080. ISO., 2022. International Organization for Standardization. Textiles-Determination of the phthalate content-Tetrahydrofuran method (ISO Standard No. 14389). 〈htt ps://www.iso.org/standard/80463.html〉. Jamarani, R., Erythropel, H.C., Nicell, J.A., Leask, R.L., Mari´ c, M., 2018. How green is your plasticizer? Polymers 10, 834. https://doi.org/10.3390/polym10080834. Kataria, A., Levine, D., Wertenteil, S., Vento, S., Xue, J., Rajendiran, K., Kannan, K., Thurman, J.M., Morrison, D., Brody, D., Urbina, E., Attina, T., Trasande, T., Trachtman, H., 2017. Exposure to bisphenols and phthalates and association with oxidant stress, insulin resistance, and endothelial dysfunction in children. Pediatr. Res. 81, 857–864. https://doi.org/10.1038/pr.2017.16. Katsikantami, I., Sifakis, S., Tzatzarakis, M.N., Vakonaki, E., Kalantzi, O.I., Tsatsakis, A. M., Rizos, A.K., 2016. A global assessment of phthalates burden and related links to health effects. Environ. Int. 97, 212–236. https://doi.org/10.1016/j. envint.2016.09.013. Khan, U., Jahangir, M., 2020. Development of a new gas chromatographic-mass spectrometry method for the simultaneous analysis of all regulated phthalates in consumer goods. Int. J. Environ. Anal. Chem. 100, 1299–1308. https://doi.org/ 10.1080/03067319.2019.1651850. Kissel, J.C., 2011. The mismeasure of dermal absorption. J. Expo. Sci. Environ. Epidemiol. 21, 302–309. https://doi.org/10.1038/jes.2010.22. Krauskopf, L.G., 2009. Plasticizers, in von Hans Zweifel, H., Maier, R.D. Schiller M. (Eds.), Plastics additives handbook. Verlag, Munich, Germany. pp. 485-513. Larsen, S.T., Butte, W., 2017. Health and safety issues with plasticizers and plasticized materials. In: Wypych, G. (Ed.), Handbook of Plasticizers (Third edition). Chem Tec Publishing, Ontario, Canada, pp. 681–744. Li, H.L., Ma, W.L., Liu, L.Y., Zhang, Z., Sverko, E., Zhang, Z.F., Song, W.W., Sun, Y., Li, Y. F., 2019. Phthalates in infant cotton clothing: occurrence and implications for human exposure. Sci. Total Environ. 683, 109–115. https://doi.org/10.1016/j. scitotenv.2019.05.132. Li, X.Y., Yang, Y.C., Cui, X., Li, S., Zhu, X.L., Tang, S.K., 2015. Determination of phthalate esters in textiles by solid phase extraction and gras chromatography-mass spectrometry. Anal. Lett. 48, 2544–2552. https://doi.org/10.1080/ 00032719.2015.1043665. Li, J.F., Zhang, Z.Z., Ma, L.Y., Zhang, Y., Niu, Z.G., 2018. Implementation of USEPA RfD and SFO for improved risk assessment of organophosphate esters (organophosphate flame retardants and plasticizers). Environ. Int. 114, 21–26. https://doi.org/ 10.1016/j.envint.2018.02.027. Licina, D., Morrison, G.C., Bek¨ o, G., Weschler, C.J., Nazaroff, W.W., 2019. Clothingmediated exposures to chemicals and particles. Environ. Sci. Technol. 53, 5559–5575. https://doi.org/10.1021/acs.est.9b00272. Liu, C., Deng, Y.L., Zheng, T.Z., Yan, G.P., Jiang, X.Q., Liu, E.N., Miao, X.P., Wang, L.Q., Jiang, M., Zeng, Q., 2020a. Urinary biomarkers of phthalates exposure and risks of thyroid cancer and benign nodule. J. Hazard. Mater. 383, 121189 https://doi.org/ 10.1016/j.jhazmat.2019.121189. Liu, J.M., Li, C.Y., Yang, F.E., Zhao, N., Lv, S.W., Liu, J.C., Chen, L.J., He, Z., Zhang, Y., Wang, S., 2020b. Assessment of migration regularity of phthalates from food packaging materials. Food Sci. Nutr. 8, 5738–5747. https://doi.org/10.1002/ fsn3.1863. Llompart, M., Celeiro, M., Dagnac, T., 2019. Microwave-assisted extraction of pharmaceuticals, personal care products and industrial contaminants in the environment. TrAC, Trends Anal. Chem. 116, 136–150. https://doi.org/10.1016/j. trac.2019.04.029. Lü, H., Mo, C.H., Zhao, H.M., Xiang, L., Katsoyiannis, A., Li, Y.W., Cai, Q.Y., Wong, M.H., 2018. Soil contamination and sources of phthalates and its health risk in China: A review. Environ. Res. 164, 417-429. https://doi.org/10.1016/j.envres.2018.03.013. Martínez, M.A., Rovira, J., Sharma, R.P., Nadal, M., Schuhmacher, M., Kumar, V., 2018. Comparing dietary and non-dietary source contribution of BPA and DEHP to prenatal exposure: A Catalonia (Spain) case study. Environ. Res. 166, 25–34. https:// doi.org/10.1016/j.envres.2018.05.008. Mohapatra, P., Gaonkar, O., 2021. An overview of toxic chemicals in textiles. Toxics Link, New Delhi. India, p. 18. 〈https://www.researchgate.net/publication/35880 4949_AN_OVERVIEW_OF_TOXIC_CHEMICALS_IN_TEXTILES〉. Mughees, M., Chugh, H., Wajid, S., 2022. Mechanism of phthalate esters in the progression and development of breast cancer. Drug Chem. Toxicol. 45, 1021–1025. https://doi.org/10.1080/01480545.2020.1802480. Negev, M., Berman, T., Reicher, S., Sadeh, M., Ardi, R., 2018. Concentrations of trace metals, phthalates, bisphenol A and flame-retardants in toys and other children’s products in Israel. Chemosphere 192, 217–224. https://doi.org/10.1016/j. chemosphere.2017.10.132. Net, S., Semp´ ere, R., Delmont, A., Paluselli, A., Ouddan´ e, B., 2015a. Occurrence, fate, behavior and ecotoxicological state of phthalates in different environmental matrices. Environ. Sci. Technol. 49, 4019–4035. https://doi.org/10.1021/ es505233b. Net, S., Delmont, A., Semp´ ere, R., Paluselli, A., Ouddan´ e, B., 2015b. Reliable quantification of phthalates in environmental matrices (air, water, sludge, sediment and soil): A review. Sci. Total Environ. 515-516, 162–180. https://doi.org/10.1016/ j.scitotenv.2015.02.013. Ortíz, A., Sansinenea, E., 2018. Di-2-ethyl hexyl phthalate may be a natural product, rather than a pollutant. J. Chem. 1–7. https://doi.org/10.1155/2018/6040814. Ouchi, Y., Yanagisawa, H., Fujimaki, S., 2019. Evaluating phthalate contaminant migration using thermal desorption–gas chromatography–mass spectrometry (TD–GC–MS). Polymers 11, 683. https://doi.org/10.3390/polym11040683. Panio, A., Corsarini, S.F., Bruno, A., Lasagni, M., Labra, M., Saliu, F., 2020. Determination of phthalates in fish fillets by liquid chromatography tandem mass spectrometry (LC-MS/MS): A comparison of direct immersion solid phase microextraction (SPME) versus ultrasonic assisted solvent extraction (UASE). Chemosphere 255, 127034. https://doi.org/10.1016/j.chemosphere.2020.127034. Panneel, L., Malarvannan, G., Jorens, P.G., Covaci, A., Mulder, A., 2021. Plasticizers in the neonatal intensive care unit: A review on exposure sources and health hazards. Crit. Rev. Environ. Sci. Technol. 22, 39473972. https://doi.org/10.1080/ 10643389.2021.1970455. Perestrelo, R., Silva, C.L., Algarra, M., Cˆ amara, J.S., 2021. Evaluation of the occurrence of phthalates in plastic materials used in food packaging. Appl. Sci. 11, 2130. https://doi.org/10.3390/app11052130. P´ erez-Fern´ andez, V., Gonz´ alez, M.J., García, M.A., Marina, M.L., 2013. Separation of phthalates by cyclodextrin modified micellar electrokinetic chromatography: Quantitation in perfumes. Anal. Chim. Acta 782, 67–74. https://doi.org/10.1016/j. aca.2013.03.072. Qadeer, A., Kirsten, K.L., Ajmal, Z., Jiang, X., Zhao, X., 2022. Alternative plasticizers as emerging global environmental and health threat: another regrettable substitution? Environ. Sci. Technol. 56, 1482–1488. https://doi.org/10.1021/acs.est.1c08365. Radke, E.G., Braun, J.M., Meeker, J.D., Cooper, G.S., 2018. Phthalate exposure and male reproductive outcomes: a systematic review of the human epidemiological evidence. Environ. Int. 121, 764–793. https://doi.org/10.1016/j.envint.2018.07.029. Radke, E.G., Galizia, A., Thayer, K.A., Cooper, G.S., 2019. Phthalate exposure and metabolic effects: a systematic review of the human epidemiological evidence. Environ. Int. 132, 104768 https://doi.org/10.1016/j.envint.2019.04.040. Rewe Group, 2018. DETOX Program. Hazardous Substances Fact Sheet. Phthalates. 〈htt ps://www.rewe-group.com/content/uploads/2020/12/FS-Phthalates.pdf〉. Rong, Y., Ali, S., Ouyang, Q., Wang, L., Li, H., Chen, Q., 2021. Development of a bimodal sensor based on upconversion nanoparticles and surface-enhanced Raman for the sensitive determination of dibutyl phthalate in food. J. Food Compos. Anal. 100, 103929 https://doi.org/10.1016/j.jfca.2021.103929. Rovira, J., Domingo, J.L., 2019. Human health risks due to exposure to inorganic and organic chemicals from textiles: A review. Environ. Res. 168, 6269. https://doi.org/ 10.1016/j.envres.2018.09.027. Roy, R.N., 2020. Bioactive natural derivatives of phthalate ester. Crit. Rev. Biotechnol. 40, 1–17. https://doi.org/10.1080/07388551.2020.1789838. Roy, R.N., Laskar, S., Sen, S.K., 2006. Dibutyl phthalate, the bioactive compound produced by Streptomyces albidoflavus 321.2. Microbiol. Res. 161, 121–126. https:// doi.org/10.1016/j.micres.2005.06.007. Roy, R.N., Sen, S.K., 2013. Fermentation studies for the production of dibutyl phthalate, an ester bioactive compound from Streptomyces albidoflavus MTCC 3662 using lowpriced substrates. Jordan J. Biol. Sci. 6, 177–181. https://doi.org/10.1080/ 07388551.2020.1789838. Saab, Y., Oueis, E., Mehanna, S., Nakad, Z., Stephan, R., Khnayzer, R.S., 2022. Risk assessment of phthalates and their metabolites in hospitalized patients: A focus on diand mono-(2-ethylhexyl) phthalates exposure from intravenous plastic bags. Toxics 10, 357. https://doi.org/10.3390/toxics10070357. Saini, A., Thaysen, C., Jantunen, L., McQueen, R.H., Diamond, M.L., 2016. From clothing to laundry water: investigating the fate of phthalates, brominated flame retardants, and organophosphate esters. Environ. Sci. Technol. 50, 9289–9297. https://doi.org/ 10.1021/acs.est.6b02038. Salazar Beltr´ an, D., 2019. Determinaci´ on de ftalatos en PET, su grado de migraci´ on al agua y su degradaci´ on mediante fotocat´ alisis heterog´ enea (PhD Thesis). Universitat de les Illes Balears, p. 11. 〈http://eprints.uanl.mx/id/eprint/21359〉. Salazar-Beltr´ an, D., Hinojosa-Reyes, L., Ruíz-Ruíz, E., Hern´ andez-Ramírez, A., Guzm´ anMar, J.L., 2017. Determination of phthalates in bottled water by automated on-line solid phase extraction coupled to liquid chromatography with UV detection. Talanta 168, 291–297. https://doi.org/10.1016/j.talanta.2017.03.060. Sastri, V.R., 2014. Polymer additives used to enhance material properties for medical device applications. Plastics in Medical Devices. Elsevier, Amsterdam. https://doi. org/10.1016/B978-0-8155-2027-6.10005-4. Sedha, S., Lee, H., Singh, S., Kumar, S., Jain, S., Ahmad, A., Bin Jardan, Y.A., Sonwal, S., Shukla, S., Simal-G´ andara, J., Xiao, J., Huh, Y.S., Han, Y.K., Bajpai, V.K., 2021. Reproductive toxic potential of phthalate compounds - State of art review. Pharmacol. Res. 167, 105536 https://doi.org/10.1016/j.phrs.2021.105536. Seyoum, A., Pradhan, A., 2019. Effect of phthalates on development, reproduction, fat metabolism and lifespan in Daphnia magna. Sci. Total Environ. 654, 969–977. https://doi.org/10.1016/j.scitotenv.2018.11.158. Si, Q., Li, F., Gao, C., Wang, C., Wang, Z., Zhao, J., 2016. Detection of phthalate esters in seawater by stir bar sorptive extraction and gas chromatography–mass spectrometry. Mar. Pollut. Bull. 108, 163–170. https://doi.org/10.1016/j.marpolbul.2016.04.034. Sinha, P., Anshu, A.K., Tiwari, M., Ahmad, M.Y., 2019. Determination of PAEs in soil of municipal wards of Patna, India by microwave assisted extraction and LC-MS/MS. Soil Sediment Contam. 28, 162–170. https://doi.org/10.1080/ 15320383.2018.1553933. Srivastava, A.K., Kumar, D., Singh, D., Rajesh Singh, R.K., 2022. Environmental exposures as xenoestrogens (bisphenol A and phthalates) enhance risk for breast cancer. Xenobiotics in Chemical Carcinogenesis. Academic Press, London, UK, pp. 197–215. Chapter 11. Starink, R.J., 2022. Results of proficiency test. Phthalates in textile April 2022. Institute for Interlaboratory Studies, Spijkenisse, The Netherlands. 〈https://www.iisnl.com/p df/iis22T03%20Report.pdf〉. N. Aldegunde-Louzao et al. Environmental Toxicology and Pharmacology 108 (2024) 104457 23 Szendi, K., Gy¨ ongyi, Z., Kont´ ar, Z., Gerencs´ er, G., Ber´ enyi, K., Hanzel, A., Fekete, J., Kov´ acs, A., Varga, C., 2018. Mutagenicity and phthalate level of bottled water under different storage conditions. Expo. Health 10, 51–60. https://doi.org/10.1007/ s12403-017-0246-x. Tang, Z., Chai, M., Wang, Y., Cheng, J., 2020. Phthalates in preschool children’s clothing manufactured in seven Asian countries: Occurrence, profiles and potential health risks. J. Hazard. Mater. 387, 121681 https://doi.org/10.1016/j. jhazmat.2019.121681. Tian, C.K., Ni, J.R., Chang, F., Liu, S., Xu, N., Sun, W., Xie, Y., Guo, Y., Ma, Y., Yang, Z., Dang, C., Huang, Y., Tian, Z., Wang, Y., 2016. Bio-Source of di-n-butyl phthalate production by filamentous fungi. Sci. Rep. 6, 19791 https://doi.org/10.1038/ srep19791. Trasende, L., Liu, B., Bao, W., 2022. Phthalates and attributable mortality: A populationbased longitudinal cohort study and cost analysis. Environ. Pollut. 292 (Part A), 118021 https://doi.org/10.1016/j.envpol.2021.118021. US-CPSC., 2008. United States Consumer Product Safety Commission. Public Law 110314-Aug. 14, 2008. 〈https://www.cpsc.gov/s3fs-public/pdfs/blk_media_cpsia.pdf〉, US-CPSC, 2017. United States Consumer Product Safety Commission. Prohibition of children’s toys and child care articles containing specified phthalates: Determinations regarding certain plastics. 16 CFR Part 130. 〈https://www.govinfo. gov/content/pkg/FR-2017-08-30/pdf/2017-18387.pdf〉. US-CPSC, 2018. United States Consumer Product Safety Commission. Directorate for Laboratory Sciences. Division of Chemistry. Test Method: CPSC-CH-C1001-09.4 Standard operating procedure for determination of phthalates. 〈https://www.cpsc. gov/s3fs-public/CPSC-CH-C1001-09.4_Standard_Operating_Procedure_for_Deter mination_of_phthalates.pdf〉. US-EPA, 2004. Risk Assessment Guidance for Superfund Volume I: Human Health Evaluation Manual (Part E, Supplemental Guidance for Dermal Risk Assessment) (Final). U.S. Environmental Protection Agency, Washington, DC. 〈https://www.epa. gov/risk/risk-assessment-guidance-superfund-rags-part-e〉. US-EPA, 2007. Phthalates. TEACH chemical summary. Environmental Protection Agency, Washington, DC. 〈https://www.epa.gov/expobox/exposure-factors-han dbook-chapter-7〉. US-EPA, 2011. Exposure Factors Handbook 2011 Edition (Final Report). U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-09/052F. 〈https ://cfpub.epa.gov/si/si_public_file_download.cfm?p_download_id=522996&Lab =NCEA〉. US-EPA, 2011b. Exposure factors handbook. Chapter 7: Dermal exposure factors. 〈https://www.epa.gov/sites/default/files/2015-09/documents/efh-chapter07.pdf〉. US-EPA, 2014. Method 8270E (SW-846): Semivolatile Organic Compounds by Gas Chromatography/ Mass Spectrometry (GC-MS). Washington, DC. 〈https://www.epa. gov/esam/epa-method-8270e-sw-846-semivolatile-organic-compounds-gas-chromat ographymass-spectrometry-gc〉. US-EPA, 2018. Regional Screening Level (RSL) Summary Table. U.S. Environmental Protection Agency, Washington, DC. 〈https://www.epa.gov/risk/regional-screenin g-levelsrsls-generic-tables〉. Virant-Klun, I., Imamovic-Kumalic, S., Pinter, B., 2022. From oxidative stress to male infertility: review of the associations of endocrine-disrupting chemicals (Bisphenols, Phthalates, and Parabens) with human semen quality. Antioxidants 11, 1617. https://doi.org/10.3390/antiox11081617. Wang, W., Jiang, D., Jin, L., Wang, P., Xu, F., Li, H., Ding, C.F., 2022. Quantitative analysis of Phthalate Esters by in-situ thermal desorption atmospheric pressure photoionization mass spectrometry using a dopant as the internal standard. Int. J. Mass Spectrom. 475, 116819 https://doi.org/10.1016/j.ijms.2022.116819. Wang, W., Leung, A.O.W., Chu, L.H., Wong, M.H., 2018. Phthalates contamination in China: Status, trends and human exposure-with an emphasis on oral intake. Environ. Pollut. 238, 771–782. https://doi.org/10.1016/j.envpol.2018.02.088. Wang, Y., Qian, H., 2021. Phthalates and their impacts on human health. Healthcare 9, 603. https://doi.org/10.3390/healthcare9050603. Wang, X., Tao, W., Xu, Y., Feng, J., Wang, F., 2014. Indoor phthalate concentration and exposure in residential and office buildings in Xi’an, China. Atmos. Environ. 87, 146–152. https://doi.org/10.1016/j.atmosenv.2014.01.018. Wang, L., Wu, Z., Gong, M., Xu, Y., Zhang, Y., 2020. Non-dietary exposure to phthalates for pre-school children in kindergarten in Beijing, China. Build. Environ. 167, 106438 https://doi.org/10.1016/j.buildenv.2019.106438. Wang, Y., Zhu, H., Kannan, K., 2019. A review of biomonitoring of phthalate exposures. Toxics 7, 21. https://doi.org/10.3390/toxics7020021. Wenzl, T., 2009. Methods for the determination of phthalates in food. EUR 23682 EN. Luxembourg (Luxembourg): OPOCE, 2009. JRC49529. 〈https://doi.org/10.2787 /1406〉. Weschler, C.J., Nazaroff, W.W., 2012. SVOC exposure indoors: Fresh look at dermal pathways. Indoor Air 22, 356–377. https://doi.org/10.1111/j.16000668.2012.00772.x. Wu, Q., Song, Y., Wang, Q., Liu, W., Hao, L., Wang, Z., Wang, C., 2021. Combination of magnetic solid-phase extraction and HPLC-UV for simultaneous determination of four phthalate esters in plastic bottled juice. Food Chem. 339, 127855 https://doi. org/10.1016/j.foodchem.2020.127855. Xie, F., Chen, X., Weng, S., Xia, T., Sun, X., Luo, T., Li, P., 2019. Effects of two environmental endocrine disruptors di-n-butyl phthalate (DBP) and mono-n-butyl phthalate (MBP) on human sperm functions in vitro. Reprod. Toxicol. 83, 1–7. https://doi.org/10.1016/j.reprotox.2018.10.011. Xie, H., Han, W., Xie, Q., Xu, T., Zhu, M., Chen, J., 2022. Fase mask-A potential source of phthalate exposure for human. J. Hazard. Mater. 422, 126848 https://doi.org/ 10.1016/j.jhazmat.2021.126848. Xie, M., Wu, Y., Little, J.C., Marr, L.C., 2016. Phthalates and alternative plasticizers and potential for contact exposure from children’s backpacks and toys. J. Expo. Sci. Environ. Epidemiol. 26, 119–124. https://doi.org/10.1038/jes.2015.71. Xue, L., Zhang, D., Wang, T., Wang, X.M., Du, X., 2014. Dispersive liquid–liquid microextraction followed by high performance liquid chromatography for determination of phthalic esters in environmental water samples. Anal. Methods 6, 1121–1127. https://doi.org/10.1039/C3AY41996G. Xue, P., Zhao, Y., Zhao, D., Chi, M., Yin, Y., Xuan, Y., Wang, X., 2021. Mutagenicity, health risk, and disease burden of exposure to organic micropollutants in water from a drinking water treatment plant in the Yangtze River Delta, China. Ecotoxicol. Environ. Saf. 221, 112421 https://doi.org/10.1016/j.ecoenv.2021.112421. Yang, J., Li, Y., Wang, Y., Ruan, J., Zhang, J., Sun, C., 2015. Recent advances in analysis of phthalate esters in foods. TRAC, Trends Anal. Chem. 72, 10–26. https://doi.org/ 10.1016/j.trac.2015.03.018. Yao, J., Hu, M., Yuan, F., Ye, H., Xu, Z., Zhang, X., Qiu, G., Dong, C., Mmereki, D., Xu, Y., Zheng, Y., Bu, B., 2022. Exposure to phthalates in the sleeping microenvironment of university dormitories: A preliminary estimate based on skin wipe and dust sampling. Build. Environ. 218, 109135 https://doi.org/10.1016/j. buildenv.2022.109135. Yue, X., Wang, Y., Zhou, Q., Lyu, Y., He, Y., Tang, Z., 2022. Phthalates in soil and road dust from a large processing trade center of children’s clothing: Occurrence, profiles and potential health risks. Process Saf. Environ. 162, 291–300. https://doi.org/ 10.1016/j.psep.2022.04.031. Zeng, D., Kang, Y., Chen, J., Li, A., Chen, W., Li, Z., He, L., Zhang, Q., Luo, J., Zeng, L., 2019. Dermal bioaccessibility of plasticizers in indoor dust and clothing. Sci. Total Environ. 672, 798–805. https://doi.org/10.1016/j.scitotenv.2019.04.028. Zhang, Y., Huang, B., Sabel, C.E., Thomsen, M., Gao, X., Zhong, M., Chen, Z., Feng, P., 2019. Oral intake exposure to phthalates in vegetables produced in plastic greenhouses and its health burden in Shaanxi province, China. Sci. Total Environ. 696, 133921 https://doi.org/10.1016/j.scitotenv.2019.133921. Zhang, C., Zhou, Ji-, Ma, T., Guo, W., Wei, D., Tan, Y., Deng, Y., 2023a. Advances in application of sensors for determination of phthalate esters. Chin. Chem. Lett. 34, 107670 https://doi.org/10.1016/j.cclet.2022.07.013. Zhang, W., Zheng, N., Wang, S., Sun, S., An, Q., Li, X., Li, Z., Ji, Y., Li, Y., Pan, J., 2023b. Characteristics and health risks of population exposure to phthalates via the use of face towels. J. Environ. Sci. 130, 1–13. https://doi.org/10.1016/j.jes.2022.10.016. Zhao, A., Wang, L., Pang, X., Liu, F., 2022. Phthalates in skin wipes: Distribution, sources, and exposure via dermal absorption. Environ. Res. 204, 112041 https://doi. org/10.1016/j.envres.2021.112041. Zuccarello, P., Oliveri Conti, G., Cavallaro, F., Copat, C., Cristaldi, A., Fiore, M., Ferrante, M., 2018. Implication of dietary phthalates in breast cancer. A systematic review. Food Chem. Toxicol. 118, 667–674. https://doi.org/10.1016/j. fct.2018.06.011. N. Aldegunde-Louzao et al.