Environmental contamination status with common ingredients of household and personal care products

doi:10.21203/rs.3.rs-1625019/v1

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Environmental contamination status with common ingredients of household and personal care products

https://doi.org/10.21203/rs.3.rs-1625019/v1

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The continuous use of household and personal care products (HPCPs) produces an immense amount of chemicals, such as parabens, bisphenols, UV filters, and alkylphenol ethoxylates, which are of great concern due to their well-known endocrine-disrupting properties. These chemicals easily enter the environment through man-made activities, thus contaminating the biota, including soil, water, plants and animals. Thus, on top of the direct exposure on account of their presence in HPCPs, humans are also susceptible to secondary indirect exposure attributed to the ubiquitous environmental contamination. The aim of this review is therefore to examine the sources and occurrence of these noteworthy contaminants (i.e., parabens, bisphenols, UV filters, alkylphenol ethoxylates), to summarize the available research on their environmental presence and to highlight the potential exposure pathways.

parabens

bisphenols

UV filters

alkylphenol ethoxylates

environmental contamination

human exposure

The continuous use of household and personal care products (HPCPs) produces an immense amount of chemicals, such as parabens, bisphenols, UV filters, and alkylphenol ethoxylates, which are of great concern due to their well-known endocrine-disrupting properties (Klančič et al., 2022). These chemicals easily enter the environment through man-made activities, thus contaminating the biota, including soil, water, plants and animals. Thus, on top of the direct exposure on account of their presence in HPCPs, humans are also susceptible to secondary indirect exposure attributed to the ubiquitous environmental contamination.

The first group of interest here are the parabens, which are esters of para-hydroxybenzoic acid that carry a linear/branched alkyl or aryl chain. They are used as preservatives in cosmetic products, foods and pharmaceuticals (Fransway et al., 2019). Commonly used parabens include methylparaben (MeP), ethylparaben (EtP), propylparaben (PrP) and butylparaben (BuP) (Fransway et al., 2019; Soni et al., 2005). The bisphenols are another well-known class of pollutants, which are composed of two phenol moieties bridged by short linkers, such as methyl or sulfonyl. Bisphenols are used as monomers in the production of polycarbonate plastic, epoxy resins, food containers and thermal paper. The most frequently used bisphenol derivative is bisphenol A (BPA), which according to Michałowicz (2014), is one of the most produced chemicals in the world. The global production of BPA was estimated at 7.7 tonnes in 2015, while it is predicted that this figure will reach 10.6 tonnes by 2022 (Bisphenol-A - A Global Market Overview - Research and Markets). Due to the identified toxic effects of BPA, some restrictions on its production and use have been imposed in North America and the European Union (EU), which has led to increased use of other bisphenols (D. Chen et al., 2016). Moreover, the use of organic UV filters in sunscreens has increased greatly more recently. Organic UV filters are a chemically diverse group of substances that primarily absorb UV radiation. There are currently 28 organic filters approved for use in cosmetic products in the EU, USA, China and Japan, at concentrations ranging from 2–20%. The most widespread chemical classes of organic UV filters are benzophenones, cinnamates, octocrylene, camphor derivatives, salycilates, dibenzoylmethanes and para-aminobenzoates (Huang et al., 2021). Finally, the alkylphenols and their ethoxylated forms (APEOs) represent another major environmental concern. Alkylphenols are used as precursors in the production of nonionic surfactant APEOs, which are then used as or in various products, including lubricating oil additives, phenol resins and antioxidants for rubber and plastics. Chemically, APEOs are simple phenol derivatives with an ethoxylated hydroxyl group and a branched alkyl group at the para position. They include nonylphenol ethoxylates (NPEOs), which account for 85% of all APEOs, and octylphenol ethoxylates (OPEOs), as 15%, which have been widely used since 1940 (Acir & Guenther, 2018; Rijkswaterstaat, 2001). All these chemicals are a concern for both humans and wildlife, due to their persistency in the environment, tendency for bioaccumulation, and potential multi-organ toxicity.

2.1. Regulatory status

Risk assessments of parabens have been carried out by several European expert panels, including the European Food Safety Authority (EFSA) and the Scientific Committee on Consumer Safety. The EFSA first established an acceptable daily intake of 0–10 mg/kg body weight for the sum of MeP, EtP and PrP. However, due to the effects of PrP on male reproductive organs in juvenile rats and the lack of a clear ‘no observed adverse effect level’, approval for PrP for use as a food additive in the European Union was removed in 2006 (Directive 2006/52/EC) (EFSA Advice on the Safety of Paraben Usage in Food). Thus, only MeP and EtP are allowed in food in the EU (maximum concentrations allowed: 2 mg/kg in food, 1 mg/L in drinks) (Lincho et al., 2021), while PrP and BuP are not allowed in food or in food-contact plastics. In the USA and other countries, this restriction does not apply (Hessel et al., 2019), as the US Food and Drug Administration (FDA) allows the use of parabens in food additives at up to 0.1% (Lincho et al., 2021). In cosmetics, MeP and EtP are currently authorised for ready-to-use cosmetic preparations at a maximum concentration of 0.4% (as acid) for a single paraben, or 0.8% (as acid) for a mixture of parabens. The use of PrP and BuP is allowed provided that the sum of their individual concentrations does not exceed 0.14% (as acid), whereby this value is based on a study in newborn rats where no adverse effects on the testes were observed. However, PrP and BuP must not be used in leave-on cosmetic products that are intended for application to the nappy area of children under three years of age, due to concerns about metabolic immaturity and the possibility of skin damage in this area that can lead to increased absorption (Annex V to Regulation (EC) No 1223/2009) (Agency, 2013). Although several other parabens are known, including iso-PrP (iPrP), iso-BuP (iBuP), benzylparaben (BzP), phenylparaben and pentylparaben, these have been banned from all cosmetic products in the EU due to lack of data (Consumers: Commission Improves Safety of Cosmetics, 2014). On the other hand, the FDA does not have specific regulations that apply only to preservatives in cosmetics. The law in the USA treats preservatives in cosmetics the same as other cosmetic ingredients (Parabens in Cosmetics | FDA). Similarly, in China, the use MeP, EtP, PrP, BuP, iPrP and iBuP is allowed. In Japan, MeP, EtP, PrP, iPrP and BuP are allowed for use in food additives and cosmetic products at concentrations up to 1% (Lincho et al., 2021).

2.2. Major sources

Parabens are found almost everywhere, and they are commonly used in the cosmetics industry and the food and pharmaceutical industries. In the USA, the individual productions of MeP, EtP and PrP were rated at 250 tonnes per year in 2002, while in 1990 the production of MeP was 5000 tonnes per year (Andersen & Larsen, 2013). Fransway et al. (2019) collected the availale data on paraben sources in PCPs, foods and medications. In eye care products, parabens were present in 31.3%, similar to facial care products, where they were present in 31.8%. Moreoever, 22.8% of topical medications and 21.6% of skin care products included parabens. For lip care and nail care products, they were present in 14.0% and 17.3%, respectively, while in hair care products and oral care products, parabens were present in only 9.5% and 9.4%, respectively. Overall, parabens were present in 21.6% of cosmetic products (Fransway et al., 2019). A survey of foods in the USA reported that 90% of foods contained parabens at a mean concentration of 9.67 ng/g, with MeP, EtP and PrP predominating. This resulted in an estimated mean daily dietary intake rate of 940, 879, 470, 273 and 307 ng/kg body weight per day for infants, toddlers, children, adolescents and adults, respectively (Wang et al., 2013). Similarly, a Chinese study reported that almost all foods (i.e., meat, seafood, eggs, dairy products, bean products, vegetables, fruit) contained at least one paraben, with a mean concentration of 39 ng/g. Notably, paraben concentrations showed a maximum of 2530 ng/g fresh foods (Liao et al., 2013). In contrast, a study tested samples from farmers’ markets in China, and only one sample was positive for MeP, at 81 ng/g (Song et al., 2017). Parabens are also used as preservatives in enteral and parenteral medications, and particularly in multidose vial antibiotics, local anaesthetics, corticosteroids, vitamins, diuretics, insulin, heparin, antihypertensives, chemotherapeutics and haloperidol (Fransway et al., 2019). A recent study reported that 97% of pharmaceutical products in China contained parabens, with MeP being the most common (Ma et al., 2016), as opposed to the USA, where only 20% of pharmaceuticals included parabens (Moreta et al., 2015). It has been estimated that the net exposure to parabens amounts to 76 mg/day, of which about two-thirds can be ascribed to cosmetics and PCPs, slightly less than one-third to pharmaceuticals, with only about 1 mg/day from food, with concentrations usually less than 1% (Fransway et al., 2019).

2.3. Level of environmental contamination

Due to the frequent use of parabens in everyday life, their presence in the environment is unavoidable. Parabens are present in raw sewage water, influent water to wastewater treatment plants (WWTPs), man-made water bodies (e.g., swimming pools), surface waters (e.g., rivers, lakes, seas), and importantly, in biotic samples (see Table 1).

As expected, the highest concentrations of parabens are seen for WWTPs. In most cases, the most frequently used parabens MeP and PrP are predominant, with these preservatives reaching up to 1.88 and 12.96 µg/L, respectively (Canosa et al., 2006; Česen et al., 2019; Terasaki et al., 2012). However, Regueiro et al. (2009) did not detect MeP in influent or effluent wastewaters, while the concentrations of PrP in influent water reached 2.6 µg/L. In river water samples, the detected concentrations of the permitted parabens were one or two orders of magnitude lower (up to 180 ng/L) than those reported for samples from WWTPs (Česen et al., 2019; Serra-Roig et al., 2016; Terasaki et al., 2012, 2015). Notably, Slovenian and Croatian river samples also contained BzP at high concentrations (175, 457 ng/L, respectively), even though BzP had been banned for use in PCPs since 2014 according to the European Chemicals Agency (ECHA) (Česen et al., 2019). In coastal waters, the concentrations of individual parabens were in the low nanogram per litre range while in sediments, the parabens reached a maximum of 12.4 ng/g (Xue et al., 2017). MeP, PrP and BuP were also detected in source and drinking water in Egypt, with the highest concentration seen for MeP as 54 ng/L (Radwan et al., 2019). The concentrations of parabens in swimming pools vary from country to country. In Spain, the individual parabens were as high as 77.5 ng/L (López-Darias et al., 2010; Regueiro et al., 2009), while in Japan, they reached 270 ng/L (Terasaki et al., 2015) and in China, 872 ng/L (Li et al., 2015).

Table 1

Occurrence of parent parabens in waters, sediments and animals.
Sample/Location	Paraben	Concentration	Reference
Raw sewage water (Spain)	MeP	1.88 µg/L	(Canosa et al., 2006)
Raw sewage water (Spain)	PrP	0.67–12.96 µg/L	(Canosa et al., 2006)
Influent water to WWTP (Japan)	PHBA, MeP, EtP, PrP, iPrP, BuP, iBuP, BzP	up to 2.3 µg/L	(Terasaki et al., 2012)
Wastewater (Slovenia and Croatia)	MeP	11.8–1910 ng/L	(Česen et al., 2019)
Influent wastewater	MeP	n.d.	(Regueiro et al., 2009)
	EtP	198.6 ng/L
	PrP	2640 ng/L
	BuP	344 ng/L
Effluent wastewater	MeP	n.d.
	EtP	n.d.
	PrP	14.0
	BuP	n.d.
River (Japan)	PHBA, MeP, EtP, PrP, iPrP, BuP, iBuP BzP	up to 25 ng/L	(Terasaki et al., 2012)
River (Japan)	Sum of parent parabens	12 to 180 ng/L	(Terasaki et al., 2015)
River (Slovenia and Croatia)	MeP	14.8–139.5 ng/L	(Česen et al., 2019)
	EtP	4.8–67.2 ng/L
	PrP	0.8–23.4 ng/L
	BzP	175–457 ng/L
River (Spain)	MeP	102 ng/L	(Serra-Roig et al., 2016)
	PrP	73.3 ng/L
	BuP	n.d.
	BzP	n.d.
Coastal waters (Florida)	MeP	3.02–31.7 ng/L	(Xue et al., 2017)
Coastal waters (Florida)	EtP	0.52–17.2 ng/L
Sediments coastal waters (Florida)	MeP	0.85–9.00 ng/g wt
Sediments coastal waters (Florida)	EtP	2.15–12.4 ng/g wt
Source water (Egypt)	MeP	54.6 ng/L	(Radwan et al., 2019)
	PrP	21.3 ng/L
	BuP	7.2 ng/L
Drinking water (Egypt)	MeP	47.2 ng/L
	PrP	15.7 ng/L
	BuP	8.4 ng/L
Pool water (Japan)	Sum of parent parabens	up to 270 ng/L	(Terasaki et al., 2015)
Swimming pool water (China)	Parent parabens	up to 872 ng/L	(Li et al., 2015)
Swimming pool (Spain)	BuP	29.1 ng/L	(López-Darias et al., 2010)
	BzP	42.9 ng/L	(López-Darias et al., 2010)
	MeP	n.d.	(Regueiro et al., 2009)
	EtP	n.d.
	PrP	32.3 ng/L
	BuP	77.5 ng/L
Seagrass, mangrove, different fish species (U.S.)	MeP, EtP, PrP, BuP	up to 213 ng/g	(Xue et al., 2017)
Different fish species (Philipines)	MeP	up to 3600 ng/g	(Ramaswamy et al., 2011)
	EtP	up to 840 ng/g
	PrP	up to 1 100 ng/g
	BuP	up to to 70 ng/g
	MeP	605 ng/g–3450 ng/g lw	(Kim et al., 2011).
	EtP	46.6 ng/g–195 ng/g lw
	PrP	46 ng/g–1140 ng/g lw
	BuP	6.61 ng/g–37.3 ng/g lw
n.d. not detected

The highest paraben concentrations have been reported for raw sewage, as the influent to WWTPs, which in itself is not a reason for concern. However, high concentrations of parabens are also present in man-made water bodies. In addition, parabens are also present in surface waters, such as rivers, lakes and oceans, albeit at lower concentrations, although they can still pose a threat to the biota. Their presence in numerous biotic samples (e.g., seagrass, mangroves, various fish), and in particular for MeP as high as 3600 ng/g, supports this (Kim et al., 2011; Ramaswamy et al., 2011; Xue et al., 2017).

2.4. Fate in the environment

Parabens can bioaccumulate and potentially bioconcentrate in aquatic organisms due to their lipophilic nature. Humans have also been shown to accumulate parabens in the fatty tissues (Darbre et al., 2004). The short-chain parabens MeP and EtP (with clogP values of 1.58, 1.80, 1.57, respectively) have a rather low potential for bioconcentration, while long-chain parabens have intermediate (clogP (PrP) = 1.96)) or even high (clogP (BuP) = 2.35) potentials for bioconcentration (Haman et al., 2015). The persistence of parabens evidently increases with the hydrocarbon chain length.

Studies have shown that parabens are removed in WWTPs with an efficiency of > 90% (Andersen et al., 2007; Yu et al., 2011), unlike in sludge, where they can remain stable for extended periods of time (Albero et al., 2012). Nonetheless, parabens can be biodegraded in raw sewage, and their half-life is estimated to be from 9.6 h to 35.2 h. Also, branched chain parabens have been shown to be less biodegradable than their linear analogues (González-Mariño et al., 2011).

2.5. Human exposure levels

The main route of exposure to parabens from cosmetic products is transdermal. However, it is estimated that only 1–9% of intact parabens are available for absorption due to their metabolism in the skin. Exposure via dermal absorption accounts for two-thirds, while exposure via medications is difficult to determine, due to the varying use of preservatives in formulations, although it is likely to account for less than one-third (Fransway et al., 2019). Only 1% of parabens enter the body via food (Andersen & Larsen, 2013; Fransway et al., 2019). Studies in rabbits have shown that approximately 60% MeP and 40% EtP penetrates the rabbit skin. In general, skin penetration of parabens decreases with increasing chain length (El Hussein et al., 2007; Pedersen et al., 2007). In rats and rabbits, parabens given orally are largely absorbed and then rapidly metabolised (Cowan-Ellsberry & Robison, 2009). Presystemic intestinal metabolism can limit systemic exposure to paraben esters (Lakeram et al., 2007). Parabens are reasonably stable in plasma in vitro, whereby MeP and EtP were shown to still be 95% intact after 24 h, while the concentrations of parabens with longer chain lengths decreased by 50% (Abbas et al., 2010). Under in vivo conditions, esters are hydrolysed by carboxyesterase 1 and carboxyesterase 2. The rate of these reactions decreases with increasing alkyl chain length (Jewell et al., 2007). When absorbed through the skin, the esters are partially hydrolysed to PHBA in keratinocytes (Handa et al., 2006). The second phase of biotransformation occurs in the liver, where parabens are conjugated with glucuronide, sulphate and glycine to facilitate their elimination (Moos et al., 2016). Indeed, more than 90% of parabens are excreted in the urine, mostly in these conjugated forms (Wang et al., 2013).

The evidence for systemic absorption of parabens is provided by their detection in various tissues (as given in Table 2). First, in blood samples, the concentrations of individual parabens range from 0.005 to 59.6 µg/L, with MeP reaching the highest concentrations (Frederiksen et al., 2011; Li et al., 2020b). Similarly, the concentrations of parabens in menstrual blood range from 0.19 to 17.7 µg/L (Jiménez-Díaz et al., 2016); interestingly, the concentrations of MeP in maternal serum reached 420 µg/L (Assens et al., 2019). Secondly, the highest paraben concentrations are found in urine. Particularly high concentrations have been reported in urine samples in Korea, compared to other Asian countries, and to the USA and Greece (mean MeP: 112 µg/L in Korea; 2.92–11.1 µg/L in other countries) (Honda et al., 2018). Also, children and pregnant women appear to be more exposed to parabens than the general population, as concentrations of individual parabens can reach up to 4890 µg/L (MeP) in children and 5380 µg/L (PrP) in pregnant women, while in the general population the sum of the parabens reaches 766 µg/L (Hajizadeh et al., 2020; Murawski et al., 2021; Pycke et al., 2015; Renzy-Martin et al., 2014; Shirai et al., 2013; L. Wang et al., 2015b; Ye et al., 2006). Paraben concentrations in pregnant women are thus an order of magnitude higher than in the general population, while the levels of exposure are also higher in the EU and the USA than in China or India (Wei et al., 2021). As women generally use more cosmetics products than men, the 12-fold higher urinary paraben concentrations determined in women are largely to be expectedly (Calafat et al., 2010). Thirdly, PHBA is the major paraben metabolite in blood and urine (Soni et al., 2005), although it has also been detected in human adipose tissue at up to 17400 ng/g (L. Wang et al., 2015a), while the concentrations of the other individual parabens did not exceed 30.6 ng/g (Artacho-Cordón et al., 2018; L. Wang et al., 2015a). Parabens have also been detected in hair samples (EtP and PrP: 0.40–1.98 µg/g) (Martín et al., 2015) and semen (0.06–0.99 µg/L) (Frederiksen et al., 2011).

Table 2

Occurrence of parabens in human tissue.
Tissue	Paraben	Concentration	Reference
Serum from men	MeP	up to 59.6 µg/L	(Frederiksen et al., 2011)
	EtP	up to 20.8 µg/L
	PrP	up to 5.50 µg/L
	BuP	up to 0.87 µg/L
	BzP	up to 0.29 µg/L
Serum from pregnant women	MeP	up to 14.8 µg/L	(A. Li et al., 2020b)
	EtP	up to 9.23 µg/L
	PrP	up to 0.053 µg/L
	BuP	up to 0.005 µg/L
	BzP	up to 0.063 µg/L
	Sum of parabens	up to 18.6 µg/L
Maternal serum	MeP	up to 420 µg/L	(Assens et al., 2019)
	EtP	up to15.13 µg/L
	PrP	up to 19.51 µg/L
	BuP	up to 7.67 µg/L
Menstrual blood	Parabens	0.19–17.7 µg/L	(Jiménez-Díaz et al., 2016)
Adipose tissue	PHBA	up to 17400 ng/g	(L. Wang et al., 2015a)
	MeP	up to 22.3 ng/g
	EtP	up to 30.6 ng/g
	PrP	up to 18.2 ng/g
	BuP	up to 1.05 ng/g
	MeP	up to 4.45 ng/g	(Artacho-Cordón et al., 2018)
	EtP	up to 2.99 ng/g
	PrP	up to 15.19 ng/g
	BuP	up to 2.74 ng/g
Placenta	MeP, BuP, BzP	up to 11.7 ng/g	(Valle-Sistac et al., 2016)
Breast milk	MeP, EtP, PrP	1.26–2.18 µg/L	(Schlumpf et al., 2010)
Hair	EtP, PrP	0.40–1.98 µg/g	(Martín et al., 2015)
Semen	Parabens	0.06–0.99 µg/L	(Frederiksen et al., 2011)
Urine samples from different countries		Mean concentrations	(Honda et al., 2018)
	MeP	2.92–11.1 µg/L; in Korea: 112 µg/L
	EtP	0.19–2.74 µg/L, in Korea: 32.7 µg/L
	PrP	0.07–1.24 µg/L, in China: 9.91, in Korea: 47.4 µg/L
Urine	MeP, EtP, PrP, BuP	up to 43.9 µg/L	(Ye et al., 2006)
Urine of urban residents	Sum of parabens	2.87–766 µg/L	(L. Wang et al., 2015b)
Urine samples from children	MeP	up to 4890 µg/L	(Murawski et al., 2021)
	EtP	up to 959 µg/L
	PrP	up to 702 µg/L
	BuP	up to 31.6 µg/L
Urine of pregnant women	MeP	297 µg/L	(Pycke et al., 2015)
	PrP	75 µg/L	(Pycke et al., 2015)
	MeP	up to 1238 µg/L	(Shirai et al., 2013)
	EtP	up to 2022 µg/L
	PrP	up to 5380 µg/L
	PHBA	up to 81.5 µg/L
	MeP	up to 2950 µg/L	(De Renzy-Martin et al., 2014)
	EtP	up to 156 µg/L
	iPrP	up to 13.3 µg/L
	nPrP	up to 346 µg/L
	nBuP	up to 62.7 µg/L
	BzP	up to 1.91 µg/L
	MeP	892 µg/L	(Hajizadeh et al., 2020)
	EtP	233 µg/L
	PrP	244 µg/L
	BuP	123 µg/L

Of particular importance, parabens can cross the placenta and can also be absorbed by the developing fetus (Frederiksen et al., 2008). Accordingly, MeP, BuP and BzP have been detected in placental samples, with the highest concentration of 11.7 ng/g fresh weight seen for MeP (Valle-Sistac et al., 2016). Parabens have also been reported for breast milk (1.26–2.18 µg/L) (Schlumpf et al., 2010), which suggested that infants can be exposed to 0.95 to 1.64 µg parabens per day, assuming they drink 750 mL breast milk per day. Therefore, collectively, parabens are present in various tissues, with the highest concentrations in urine, although significant amounts are also found in breast milk, which indicates that breast-fed infants can be exposed to these preservatives.

3.1. Regulatory status

In the EU, BPA is classified as a substance that adversely affects reproduction, irritates the respiratory tract, causes serious eye damage, and can trigger skin allergies. Companies supplying BPA to the EU must classify and label it, and mixtures containing it, according to the harmonised classification. Germany has proposed additional harmonised classifications that include the risk to the aquatic environment (acute and chronic). The Committee for Risk Assessment (RAC) that operates within the framework of the ECHA has supported this proposal. Since March 2018, the use of BPA as a stand-alone substance and in mixtures intended for consumer use has been restricted in the EU. The use of BPA in thermal paper was further restricted in 2020. Henceforth, thermal paper containing 0.02% or more BPA is no longer allowed on the EU market. France and Sweden have proposed to restrict over 1000 skin-sensitising chemicals in clothing, footwear and other articles with similar skin contact, including BPA. The RAC has supported this proposal, and a final decision is to be taken by the Commission together with the Member States (Bisfenol A - ECHA). In the EU, BPA can be used in food contact materials, but only 0.05 mg/kg is allowed to leach from the materials into the food. Of note, BPA has been banned in infant bottles throughout the EU since June 2011. Later, in 2018, it was also banned from plastic bottles and packaging containing food for babies and children under 3 years of age. Moreover, France has banned BPA in all food packaging, containers and utensils. The EFSA has also limited the amount of BPA that can leach from toys for children up to 3 years of age, and from toys intended to be placed in a child's mouth, to 0.04 mg/L (Scientific Opinion on the Risks to Public Health Related to the Presence of Bisphenol A (BPA) in Foodstuffs, 2015). In its draft re-evaluation in December 2021, the EFSA proposed lowering the TDI from 4 µg/kg bw/day to 0.04 ng/kg bw/day (Bisphenol A, EFSA 2021). The most recent ECHA market survey on the use of BPA and other thermal paper developers confirmed that paper manufacturers continue to substitute BPA with bisphenol S (BPS). Accordingly, in 2019, 187 kilotonnes of BPS-based thermal paper were placed on the EU market, and it is expected that by 2022, 61% (307 kilotonnes) of all of the thermal paper in the EU will be BPS based. The widespread use of BPS in thermal paper also constitutes a cause for concern, as BPS is believed to also affect the human reproductive and endocrine systems. To this end, Belgium has proposed to harmonise the classification and labelling of BPS as toxic for reproduction (Bisphenol S has replaced bisphenol A in thermal paper - All News - ECHA). Much less information on the restrictions of BPA is available for the USA, however. Since 2012, BPA has no longer been allowed in polycarbonate resins for baby bottles and ‘sippy’ cups. The FDA also banned BPA-based epoxy resins as coatings in packaging for infant formula (Bisphenol A (BPA): Use in food contact application | FDA). Moreoever, BPA should not be present in cell phone cases in concentrations over 3 ppm, and thermal paper at over 10 ppm. Other countries have restricted the use of BPA in food contact products (i.e., Argentina, Australia [voluntary phase out], Brazil, Canada, China, Costa Rica, Ecuador, Japan, Malaysia, South Africa, South Korea, Taiwan, Turkey), especially in those for infant use (BPA Bans and restrictions in food contact materials | SGS).

3.2. Major sources

Bisphenols are composed of two phenol moeties, and several prominent examples are shown in Figure 1. They are generally used in the manufacture of polycarbonate plastics and epoxy resins. Synthetic polymers of BPA are characterised by good mechanical properties, low moisture absorption and thermal stability. Therefore, they are used in various products such as water pipes, food containers, bottles, toys, nipples, medical devices, dental products, electronic devices and CD/DVD disks (Michałowicz, 2014). The annual production of BPA in the USA ranged from 940000 tonnes to 1030000 tonnes from 2000 to 2019, but the Asia-Pacific region dominates the bisphenols market (Bisphenol A production USA 2019). As well as polycarbonate plastics and epoxy resins, BPA is used in the paper industry as a colour developer in thermal paper (Geens et al., 2012). In the thermal paper used for receipts, BPA is the most frequent of all the bisphenols, at concentrations from 5.6 to 30.4 mg/g, while its alternative BPS has been reported at lower concentrations, from 3.3 to 13.2 mg/g (Goldinger et al., 2015). However, since January 2020, thermal paper containing more than 0.02% BPA is no longer allowed on the EU market (Bisfenol A - ECHA).

As indicated, BPA is still the most widely used among the bisphenols, but due to increasing restrictions, the use of other bisphenols has increased, and especially of BPS. Since the discovery of BPA-related adverse effects on the kidneys and mammary glands, the EU has temporarily reduced the daily intake of BPA from 50 to 4 mg/kg body weight/day, and many countries have prohibited its use in children's products. As a result, many companies have switched to using other bisphenol analogues, such as BPS and bisphenol F (BPF). Similar to BPA, BPS is used in polycarbonates, epoxies, thermal paper, and certain consumer products that can thus be labelled as ‘BPA-free’. It is also used in personal care products and 13 categories of food (Liao & Kannan, 2013; Zheng et al., 2018). The European consumption of BPS is estimated to be between 1000 and 10000 tonnes per year (Lehmler et al., 2018).

3.3. Level of environmental contamination

The widespread use of bisphenols in everyday life has led to their occurrence in the environment. This has been confirmed for both WWTP influent and effluent water, as well as for surface waters, including rivers, lakes and seas, which has consequently resulted in contamination of the biota (e.g., accumulation in fish) (see Table 3).

Table 3

Occurrence of bisphenols in the environment.

Sample/Location	Bisphenol	Concentration	Reference
Effluent water from paper recycling plant (Japan)	BPA	up to 370 µg/L	(Fukazawa et al., 2001)
Influent water from WWTP (Romania)	BPS	1160-1688 ng/L	(Chiriac et al., 2021)
	BPE	483-744 ng/L
	BPA	6422-9140 ng/L
	BPF, BPB, BPC	n.d.
Effluent water from WWTP (Romania)	BPA	6.32-73.6 ng/L
Effluent water from WWTP (Romania)	BPS, BPE, BPF, BPB, BPC	n.d.
Influent water from WWTP (Slovenia)	Sum of bisphenols	6.66 – 16.4 ng/L	(Česen et al., 2018)
River water near chemical industrial zone (China)	BPA	average: 415 ng/L	(Lan et al., 2019)
River water (Spain)	BPA	up to 817 ng/L	(Bolívar-Subirats et al., 2021)
River water (China)	BPS	up to 18.99 ng/L	(Yang et al., 2014b)
	BPA	up to 74.58 ng/L
	BPAF	up to 245.69 ng/L
River water (Romania)	BPS	6.15-8.23 ng/L	(Chiriac et al., 2021)
	BPA	74.5-135 ng/L
	BPE, BPF, BPB, BPC	n.d.
Influent water (Canada)	BPA	up to 1000 ng/L	(Muir et al., 2017)
Harbour water (Canada)	BPA	up to 10 ng/L	(Muir et al., 2017)
Surface water samples (Japan, China, Korea, India)	BPF	up to 2850 ng/L in Japan	(Yamazaki et al., 2015)
	BPA	up to 1950 ng/L in India
	BPS	up to 7200 ng/L in India
19-23 freshwater sites; 8-14 marine locations (Netherlands)	BPA	up to 21 µg/L	(Belfroid et al., 2002)
Surface sediments (USA, Japan, Korea)	BPA	117 ng/g	(Liao et al., 2012b)
	BPS	12.37 ng/g
	BPF	69.7 ng/g
	Others	2 ng/g or less
Sediment near chemical industrial zone (China)	BPA	average 521 ng/g dw	(Lan et al., 2019)
Sediment (China)	BPF	up to 30.17 ng/g	(Yang et al., 2014b)
	BPA	up to 42.76 ng/g
	BPAF	up to 2009.8 ng/g
Indoor dust (US, China, Japan, Korea)	BPA	1.6 µg/g	(Liao et al., 2012a)
	BPS	0.36 µg/g
	BPF	0.096 µg/g
	Other	less than 1 ng/g
Indoor dust (China, Columbia, Greece, India, Japan, South Korea, Kuwait, Pakistan, Romania, Saudi Arabia, US, Vietnam)	BPA	1000 ng/g	(W. Wang et al., 2015)
	BPF	1000 ng/g
	BPS	220 ng/g
	BPAF	3.1 ng/g
	Others	less than 1 ng/g
Food	Bisphenols	4.38 ng/g	(Liao & Kannan, 2011)
Fish	BPA	liver: 2–75 ng/g	(Belfroid et al., 2002).
		muscle: 1–11 ng/g	(Belfroid et al., 2002).
		muscle: up to 59.3 ng/g	(Ademollo et al., 2018)
		liver: up to 105.3 ng/g	(Ademollo et al., 2018)
		liver: 36.2–51.4 ng/g	(Errico et al., 2017)
		muscle: 18.8–84.8 ng/g	(Errico et al., 2017)
		brain: 31–46 ng/g	(Ros et al., 2016)
		muscle: 20–28 ng/g
		liver: 49–97 ng/g
		muscle: 16.2 ng/g	(Barboza et al., 2020)
	BPB	muscle: 7.3 ng/g	(Barboza et al., 2020)
Canned seafood	BFDGE	up to 4.2 µg/g	(Theobald et al., 2000).

n.d. not detected

Bisphenol A was detected in wastewater from a paper recycling plant in Japan at concentrations of up to 370 µg/L (Fukazawa et al., 2001). In contrast, a more recent study reported lower concentrations of bisphenols in influent water from a WWTP in Romania, as BPS (1160-1688 ng/L), bisphenol E (BPE; 483-744 ng/L) and BPA (6422-9140 ng/L). As expected, the concentrations were two to three orders of magnitude lower in the effluent water, where BPA ranged from 6.32 to 73.6 ng/L, while BPS was below the limit of quantification, and BPE was not detected (Chiriac et al., 2021). Similar concentrations of bisphenols were detected in influent wastewater in Slovenia (6.66-16.4 ng/L) (Česen et al., 2018). River samples generally contain lower concentrations of bisphenols than influent wastewater, with the exception of a river in China near a chemical industrial zone (mean BPA, 415 ng/L) (Lan et al., 2019) and in a river in Spain (bisphenols up to 817 ng/L) (Bolívar-Subirats et al., 2021). Samples from different rivers have been reported to contain individual bisphenols at maximum concentrations from 6.15 to 245.69 ng/L (Chiriac et al., 2021; Lan et al., 2019). In Canada, BPA has been reported for surface waters near WWTPs at up to 1000 ng/L, while in harbour waters, BPA levels were two orders of magnitude lower (up to 10 ng/L) (Muir et al., 2017). In Japan, surface waters have shown BPF at 2850 ng/L, and in India, BPA and BPS in surface waters were 1950 ng/L and 7200 ng/L, respectively (Yamazaki et al., 2015). In addition, in samples of surface water collected during the summer of 2002 in The Netherlands, BPA was reported at up to 21 µg/L, although at that time the regulation of bisphenols was not as stringent (Belfroid et al., 2002).

In sediments from industrial areas in the USA and Japan, BPA was reported as a few ng/g, while for those in Korea, the mean BPA was 567 ng/g (maximum, 13,370 ng/g). The mean BPF and BPS in Korea were 69.7 and 12.37 ng/g, respectively. Other bisphenols were detected at concentrations of 2 ng/g or less (Liao et al., 2012b). Similarly, Lan et al. (2019) detected BPA in sediment samples in China at 521 ng/g. However, again in China, bisphenol AF (BPAF) was reported to be predominant in sediment samples with a maximum concentration of 2009.8 ng/g, which can be attributed to its use in the production of phenolic resins and fluoroelastomers (Yang et al., 2014b). BPA and other bisphenols have also been reported to contaminate indoor dust. The most commonly detected bisphenols were BPA, BPS and BPF, at mean concentrations from 0.096 to 1.6 µg/g. However, in certain cases, BPA reached 32 µg/g and BPF reached 110 µg/g, as it is an important substitute for BPA (Liao et al., 2012a; W. Wang et al., 2015). As thermal paper is recycled, these substances can find their way back into everyday life, as has been shown for BPA. Therefore, they can occur in other paper products, such as magazines, leaflets, newspapers, paper towels and toilet paper, with a concentration of 14.4 µg/g reported (Liao & Kannan, 2011).

Bisphenol A and its analogues were also found in 75% of food samples in one study, with a mean concentration of 4.38 ng/g fresh weight (Liao & Kannan, 2011). BPA was found predominantly in liver and muscle tissue of fish, where it ranged from 1 to 105 ng/g, generally as higher in liver than muscle tissue. Overall, the highest BPA concentrations have been reported for Greenland shark (105 ng/g), which is at the top of the food chain (Ademollo et al., 2018; Barboza et al., 2020; Belfroid et al., 2002; Errico et al., 2017; Ros et al., 2016). Furthermore, bisphenol-F-diglycidyl ether has been reported for canned sardines (0.1-3.8 µg/g), tuna (0.1-4.2 µg/g), anchovies (up to 4.1 mg/kg), mackerel (up to 1.6 µg/g), mussels (up to 0.5 µg/g) and oysters (up to 0.2 µg/g) (Theobald et al., 2000). To provide another perspective on these concentrations, if a person eats half of a can of sardines (i.e., 100 g), they will be exposed to up to 380 µg bisphenol-F-diglycidyl ether, or if they consume 150 g fish, they can be exposed to up to 12.7 µg BPA.

Overall, the bisphenols are ubiquitous, with the highest concentrations found in wastewaters, and lower concentrations in rivers and seas, as expected, albeit with some notable exceptions. On the other hand, unexpectedly high concentrations have been reported for indoor dust. The various bisphenols in fish and canned food (as mainly BPA) is another matter for concern, as this represents direct exposure of humans to bisphenols.

3.4. Fate in the environment

Bisphenols have different stabilities under different conditions. The clogP values of BPA, bisphenol B (BPB), BPE, BPF and BPS are 3.06, 3.40 3.02, 2.58 and 1.88, respectively. D. Chen et al. (2016) ranked the degradation of bisphenol analogues under anaerobic conditions in an aquatic environment, where BPF was the most biodegradable, followed by BPS and BPA, then BPE, and the least biodegradable was BPB (i.e., BPF > BPS, BPA > BPE > BPB). In an aquatic environment, BPF degrades faster than BPA, while BPS, BPB and bisphenol P (BPP) tend to be more resistant to biodegradation than BPA. When comparing various conditions, bisphenol analogues are less biodegradable in sediments (half-life [t_½], 135-1621 days) than in soil (t_½, 30-360 days) and water (t_½, 15-180 days) (D. Chen et al., 2016).

3.5. Human exposure levels

The major routes of exposure to BPA are ingestion, inhalation or dermal ingestion. Exposure can occur via food packaging, as BPA can leach from food and beverage containers, dust, dental materials, medical devices, thermal paper and toys and articles for children and infants. The daily intake of BPA is higher than that of the other bisphenols. Based on the urinary excretion data, the estimated daily intakes of BPA, BPF, BPS, BPP, bisphenol AP, BPB, bisphenol Z (BPZ), and BPAF have been reported as 2.53, 0.68, 0.60, 0.41, 0.36, 0.29, 0.24 and 0.06 μg/day, respectively (Wang et al., 2020). Most of the ingested BPA is absorbed orally (up to 77%) (Gayrard et al., 2019), while only 1.7% to 3.6% is absorbed through the skin (Toner et al., 2018). Similarly, BPS is almost 100% absorbed in the gastrointestinal tract. Both BPA and BPS undergo glucuronidation either in the intestine (BPA: 44%) or the liver (BPA: 53%), whereby BPS is glucuronidated (41%) (Gayrard et al., 2019) or sulphated; subsequently they are excreted via the kidneys (Fenichel et al., 2013).

Bisphenols have been detected in blood, adipose tissue, placenta, breast milk and urine samples (see Table 4). In blood samples of pregnant women, BPA has been reported to range from 0.4 to 86 µg/L, with the sum of bisphenols at 144 µg/L (Cambien et al., 2020; González et al., 2019; Jin et al., 2018; Li et al., 2020°). In urine samples from children, women and men, bisphenols can range from 0.32 to 1.84 μg/L (Radwan et al., 2018; Snoj Tratnik et al., 2019; Venisse et al., 2014; Yang et al., 2014a). Interestingly, a higher mean bisphenol concentration was reported for children (1.81 μg/L) compared to their mothers (1.02 μg/L) and fathers (0.32 μg/L) (Snoj Tratnik et al., 2019). BPA has been detected in adipose tissue of humans at a mean concentration of 5.83 ng/g (Fernandez et al., 2007). It is worth noting that BPA has also been detected in placental tissue, at up to 22.2 ng/g (BPA) (Jiménez-Díaz et al., 2010). Of particular concern, the colostrum from women in France has also been reported to contain BPA (1.87 μg/L), thus even infants are exposed to bisphenols (Migeot et al., 2013).

Table 4

Occurrence of bisphenols in human tissues.

Tissue	Bisphenol	Concentration	Reference
Blood from workers at hazardous waste incinerator	BPA	0.46 µg/L	(González et al., 2019)
Plasma	BPA	0.4 µg/L	(Jin et al., 2018)
	BPS	0.15 µg/L
	BPAF	0.073 µg/L
Serum of pregnant women	Sum of BPs	up to 144 µg/L	(A. Li et al., 2020a)
Serum of pregnant women	BPS	0.113 µg/L	(A. Li et al., 2020a)
Blood	BPA	0.266 – 86.831 µg/L	(Cambien et al., 2020)
Adipose tissue	BPA	5.83 ng/g	(Fernandez et al., 2007)
Placenta	BPA	5.7-22.2 ng/g	(Jiménez-Díaz et al., 2010)
Colostrum	BPA	1.87 µg/L	(Migeot et al., 2013)
Urine	BPA, BPS, BPF, BPAF	up to 4.38 µg/L	(Yang et al., 2014a)
	BPA	children: up to 1.81 µg/L	(Snoj Tratnik et al., 2019)
		mothers: up to 1.02 µg/L
		fathers: up to 0.32 µg/L
	BPA	up to 1.84 µg/L	(M. Radwan et al., 2018)
	BPA	up to 1.378 µg/L	(Venisse et al., 2014)

n.d. not detected

4.1. Regulatory status

Sunscreens represent a particular challenge in international harmonisation, as they are considered drugs under US law and cosmetics under EU law. However, there is interest in examining the approval procedures for active sunscreen ingredients in different countries, and in sharing information on test procedures for the measurement of Sun Protection Factors (Overview of International Activities for Cosmetics | FDA).

In the EU, the maximum concentration allowed for an individual organic UV filter is 3% to 10%, while that for inorganic ZnO and TiO₂ is 25% (Cosmetic Products Regulation, Annex VI - ECHA). There are currently 28 organic UV filters allowed in the EU, while there are only 22 allowed in the USA. In addition, there are some differences in the concentrations allowed between countries, especially for benzophenones. Oxybenzone (BP-3) is allowed at 6% in the USA and the EU, but at 10% in Australia and 5% in Japan. Dioxybenzophenone (BP-8) is only allowed in the USA and Australia at 3%, while benzophenones-1 (BP-1), -2 (BP-2) and -6 (BP-6) are only allowed in Japan at 10%. European and American regulations also appear to be stricter on other UV filters, such as avobenzone (BMDM), which is allowed in the EU at 5% (w/w) and in the USA at 3% (w/w), while in Japan it is allowed up to 10% (w/w). Furthermore, OMC is allowed in the EU and Australia at 10% and in the USA at 7.5%, but in Japan up to 20% (Sabzevari et al., 2021). By amending Annexes II and VI of Regulation (EC) No. 1223/2009 of the European Parliament and of the Council on Cosmetic Products, the EU banned 3-benzylidene camphor (3-BC) from the market to ensure safety for human health (European Commission, 2014).

4.2. Major sources

UV filters are largely used in personal care products (see Fig. 2 for representative organic UV filters). In the USA, their annual use ranges from 9513 to 32507 tonnes, while in Europe it ranges from 1000 to 10000 tonnes per year (Kwon & Choi, 2021). According to ECHA, the UV filters approved in the EU include both organic and inorganic filters (Cosmetic Products Regulation, Annex VI - ECHA). Organic UV filters absorb UVA and/or UVB radiation, while inorganic UV filters mainly scatter and reflect the radiation. In most cases, organic UV filters are used in combinations, as none of them can ensure a sufficiently high Sun Protection Factor at the concentrations allowed (Serpone et al., 2007). Their uses vary from country to country. For example, Manová et al. (2013) detected 22 of 28 UV filters approved in Switzerland in 116 products, and discovered 17 different UV filters. The most common were butylmethoxydibenzoylmethane, OMC and octocrylene, which were detected in 71%, 51% and 43% of the samples, respectively. The number of UV filters contained in PCPs ranged from 1 to 7 (mean, 3). Sunscreens contained between 2 and 7 UV filters in a single product (mean, 4), while face creams contained between 1 and 5 (mean, 3). Lip-care products, lipsticks, hand creams, aftershaves and liquid make-up foundations also contained UV filters. BMDM was prevalent in sunscreens (mean concentration, 3.6%). Hand and face creams also contained BMDM predominantly. On the other hand, OMC was not as frequent in sunscreens; it was predominantly found in liquid make-up foundations (mean concentration, 4.4%) and in lipsticks (mean concentration, 5.2%), and was frequently detected in lip-care products and face creams (Manová et al., 2013). In addition to their use in PCPs, UV filters are also used to protect materials such as plastics and rubber from sunlight (Ramos et al., 2016). There are also some technologies that use UV filters in nanoparticles with ethyl cellulose, to obtain UV protection in textiles (Vílchez-Maldonado et al., 2014).

4.3. Level of environmental contamination

UV filters can enter the environment not just directly, in particular during the summer season through people swimming in the sea, lakes, rivers and swimming pools, but also indirectly, through sewage treatment plants (see Table 5 for environmental occurrence) (Silvia Díaz-Cruz et al., 2008). As expected, the concentrations of UV filters were higher in sludge from WWTPs than in other samples, as seen for 4-methylbenzylidene camphor (4-MBC; 1780 ng/g), octocrylene (4840 ng/g) and octyltriazone (5510 ng/g) (Plagellat et al., 2006). Benzophenones are also found frequently in soil samples from industrial drainage (up to 0.204 ng/g), and at particularly high concentrations in soil (16.550 ng/g) and sediments (18.380 ng/g) (Jeon et al., 2006). UV filters are also found in lakes, with the highest concentrations during the holiday season, which have been reported as octocrylene (250 ng/L), BP-3 (125 ng/L) and 4-MBC (82 ng/L) (Balmer et al., 2005; Jeon et al., 2006; Poiger et al., 2004; Rodil et al., 2008). In river samples, however, the reported concentrations are lower, such as for BP-1 at 47 ng/L (Jeon et al., 2006; Rodil et al., 2008). In seawater, the concentrations of UV filters can exceed those in rivers and lakes. For example, total concentrations in Japan ranged from 505 to 1770 ng/L, with BP-3 contributing up to 1340 ng/L (Tashiro & Kameda, 2013). In the USA, high concentrations have been seen for octocrylene (up to 3730 ng/L), BMDM (up to 1298 ng/L), octyl methylcinnamate (OMC; up to 341 ng/L), homosalate (HMS; 625.7 ng/L) and BP-8 (up to 2203 ng/L) (Bratkovics et al., 2015; Mitchelmore et al., 2019). Interestingly, similar concentrations of UV filters have been detected in swimming pool water. In Greece, BP-3 in swimming pools ranged from 2.4 to 3.3 µg/L, while in shower effluents it was 2–3-fold higher (Lambropoulou et al., 2002). Similar concentrations have also been reported in water from reservoirs in Australia for octocrylene (1380-4660 ng/L), 4-MBC (436-1240 ng/L) and BMDM (424-1130 ng/L) (O’Malley et al., 2021).

Table 5

Occurrence of UV filters in waters, sediments and animals.

Location	UV filter	Concentration	Reference
Sludge (Switzerland)	4-MBC	1780 ng/g	(Plagellat et al., 2006)
	OMC	110 ng/g
	OC	4840 ng/g
	OS	5510 ng/g
Soil and sediment (Korea)	Benzophenones	industrial drainage: up to 0.204 ng/g	(H. K. Jeon et al., 2006)
		ground soil: up to 16.550 ng/g
		sediment: up to 18.380 ng/g
Lake (Switzerland)	4-MBC	28 ng/L	(Balmer et al., 2005)
	BP-3	35 ng/L
	OMC	7 ng/L
	OC	5 ng/L
Lake (Switzerland)	4-MBC	< 2 – 82 ng/L	(Poiger et al., 2004)
	OMC	< 2 – 26 ng/L
	OC	< 2 – 27 ng/L
	BP-3	< 2- 125 ng/L
	BMDM	< 20-24 ng/L
2 lakes 2 rivers Treated wastewater (Spain)	EHS, HMS, 4-MBC, BP-3, OMC, OD -PABA, OC, BMDM	lakes: < LOD – 250 ng/L	(Rodil et al., 2008)
		rivers: < LOD – 42 ng/L
		wastewater: <LOD – 54 ng/L
25 rivers, 6 lakes (Korea)	benzophenones	rivers: up to 47 ng/L	(H. K. Jeon et al., 2006)
25 rivers, 6 lakes (Korea)	benzophenones	lakes: up to 85 ng/L	(H. K. Jeon et al., 2006)
Seawater (Japan)	10 different UV filters	total concentration: 505-1770 ng/L	(Tashiro & Kameda, 2013)
Surface seawater (U.S.)	HMS	53.0 – 625.7 ng/L	Mitchelmore et al., 2019)
	OS	33.1 – 96.0 ng/L
	BP-3	n.d. – 136.2 ng/L
	OC	n.d. – 26.9 ng/L
	OD PABA	n.d. – 11.2 ng/L
	OMC	n.d.
	4-MBC	n.d.
	BMDM	n.d.
Coastal water (U.S.)	OC	up to 3730 ng/L	(Bratkovics et al., 2015)
	BMDM	up to 1298 ng/L
	OMC	up to 341.1 ng/L
	BP-8	up to 2203 ng/L
Swimming pools Seawater Shower wastes (Greece)	BP-3	swimming pool: 2.4 – 3.3 mg/L	(Lambropoulou et al., 2002)
	BP-3	shower waste: 8.2 – 9.9 mg/L
	OD PABA	swimming pool: n.d.–2.1 mg/L
	OD PABA	shower waste: 5.3–6.2 mg/L
Water from reservoir (Australia)	OC	1380–4660 ng/L	(O’Malley et al., 2021)
	4-MBC	436–1240 ng/L
	BMDM	424–1130 ng/L
Mussel (Spain)	4-MBC	up to 49 ng/g dw	(Rodil et al., 2019)
	BP-3	up to 63 ng/g dw
	OMC	up to 94 ng/g dw
	OC	up to 141 ng/g dw
	OD PABA	up to 12 ng/g dw
	BP-4	up to 87 ng/g dw
Mussel (Spain)	4-MBC	0.25–18 ng/g	(Lindo-Atichati et al., 2019)
	OC	0.05–19 ng/g
	BP-4	0.5–11.6 ng/g
Oysters and mussels (U.S.)	BP-3	up to 118 ng/g	(He et al., 2019)
	OMC	up to 192 ng/g
	HMS	up to 158.3 ng/g
	OS	up to 20.4 ng/g
Cyprinid (Greece)	BP-2	11.8–41.9 ng/g	(Díaz-Cruz et al., 2019)
Cyprinid (Greece)	BP-3	n.d–1.8 ng/g	(Díaz-Cruz et al., 2019)
Fish from lake (Switzerland)	4-MBC	166 ng/g	(Balmer et al., 2005)
	BP-3	123 ng/g
	OMC	64 ng/g
	OC	25 ng/g
Coral tissue (U.S.)	OS	210.4–491.0 ng/g dw	(Mitchelmore et al., 2019)
	HMS	188.7–441.1 ng/g dw
	OC	31.3–261.8 ng/g dw
	BP-3	5.8–240.9 ng/g dw
	BMDM	53.5–287.1 ng/g dw
Bogue (Spain)	OMC	1048-3400 ng/g	(Henríquez-Hernández et al., 2017)
Bogue (Spain)	3-BC	20.67-57.98 ng/g	(Henríquez-Hernández et al., 2017)
Cod (liver), Prawn, Crab, Burbot, Whitefish, Perch (Norway)	BP-3	cod liver: up to 1037 ng/g	(Langford et al., 2015)
	BP-3	whitefish: up to 182 ng/g
	OC	cod liver: up to 11875 ng/g
Dolphin (Brazil)	OC	89−782 ng/g lipid weight	(Gago-Ferrero et al., 2013)

n.d. not detected

Most UV filters are lipophilic, and thus they can accumulate in the human body and in aquatic organisms (Fent et al., 2010). Of the UV filters found in biota, BP-3, OMC and 4-MBC occur more frequently and at higher concentrations. Lower concentrations of UV filters (tens of nanograms/gram) have been reported for mussels, oysters and smaller fish, while coral tissue can contain up to 491 ng/g octylsalate and up to 441 ng/g HMS. The highest concentrations were reported for dolphin and cod tissues, where the concentration of octocrylene reached 11.875 µg/g (Balmer et al., 2005; Díaz-Cruz et al., 2019; Gago-Ferrero et al., 2013; He et al., 2019; Henríquez-Hernández et al., 2017; Langford et al., 2015; Lindo-Atichati et al., 2019; Mitchelmore et al., 2019; Rodil et al., 2019).

4.4. Fate in the environment

Due to their lipophilicity (clogP values: BP-3, 2.75; OMC, 4.41; octyl dimethyl para-aminobenzoic acid [OD-PABA], 4.12; octocrylene, 5.45; BMDM, 4.07; 4-MBC, 4.22; octyl salate, 3.87), UV filters are expected to persist in sediments, albeit with varying half-lives (t_½). For example, BP-3 has a low clogP (2.75) and is biodegraded more quickly (t_½: anaerobic: 1.5 days; aerobic: 5.3 days) than octocrylene (clogP: 5.45; t_½ aerobic: 10.3 days; anaerobic: 22 days) and 4-MBC (clogP: 4.22; t_½ aerobic: 33 days, anaerobic: 75 days) (Liu et al., 2013). However, their removal from wastewaters is relatively efficient, with a study in Switzerland reporting high elimination rates for 4-MBC (18%-82%), BP-3 (68%-96%), octocrylene (88%-99%) and OMC (97%-99%) (Balmer et al., 2005).

4.5. Human exposure levels

Due to their frequent use, it is not surprising that UV filters have been detected in many tissues (see Table 6). The exposure to UV filters primarily occurs via dermal absorption, and to a small extent via ingestion (e.g., accidental swallowing of water) or inhalation. The estimated daily intake of organic UV filters for infants and adults ranges from 0.85 to 6.18 ng/kg body weight per day and 0.07 to 0.49 ng/kg body weight per day, respectively (Ao et al., 2017). It appears that BP-3 is the main contaminant among the UV filters.

Table 6

Occurrence of UV filters in human tissues.

Tissue	UV filter	Concentration	Reference
Serum	BP-1	up to 0.86 μg/L	(Frederiksen et al., 2021)
	BP-3	up to 5.18 μg/L
	4-MBC	up to 10.7 μg/L
	3-BC	up to 1.61 μg/L
Blood	BP-3	up to 3.38 μg/L	(Zhang et al., 2013)
Serum after topical application of sunscreen	BP-3	238 μg/L	(Janjua et al., 2008)
	4-MBC	18 μg/L
	OMC	16 μg/L
	BP-3	300 μg/L	(Janjua et al., 2004)
	4-MBC	20 μg/L
	OMC	20 μg/L
Adipose tissue	BP-3	0.6 ng/g	(Artacho-Cordón et al., 2018)
Human breast tissue	BP-3	up to 26 μg/L	(Barr et al., 2018)
	OMC	up to 58.7 μg/L
	4-MBC	up to 25.6 μg/L
	HS	n.d.
Breast milk	BP-3	up to 799.9 ng/g	(Molins-Delgado et al., 2018)
Breast milk	4-MBC	up to 10.8 ng/g	(Molins-Delgado et al., 2018)
Seminal fluid	BP-1	up to 10.9 μg/L	(Frederiksen et al., 2021)
	BP-3	up to 5.94 μg/L
	4-MBC	up to 16.03 μg/L
	3-BC	up to 2.01 μg/L
Urine of children	Benzophenone type	Up to to 57.1 μg/L	(Lu et al., 2018)
Urine	BP-1	up to 121 μg/L	(Frederiksen et al., 2021)
	BP-3	up to 603 μg/L
	4-MBC	up to 2.58 μg/L
	3-BC	n.d.
	BP-3	up to 46.1 μg/L	(Zhang et al., 2013)
	BP-3	81 μg/L	(Janjua et al., 2008)
	4-MBC	4 μg/L
	OMC	4 μg/L
	BP-3	140 μg/L	(Janjua et al., 2004)
	4-MBC	7 μg/L
	OMC	8 μg/L

n.d. not detected

UV filters have been detected in blood serum samples from volunteers after topical whole body application (2 mg/cm²) of sunscreens containing 10% (v/v) each of BP-3, 4-MBC and OMC. The highest serum concentration was for BP-3 (238 μg/L), with 4-MBC (18 μg/L) and OMC (16 μg/L) more than one order of magnitude lower (Janjua et al., 2008; Janjua et al., 2004). For comparison, the concentrations of the UV filters in control human blood samples (i.e., none applied to the skin) were two orders of magnitude lower again. In line with previous observations, the highest concentration was also obtained for BP-3 (5.18 μg/L) (Frederiksen et al., 2021; Zhang et al., 2013).

Many studies have detected various UV filters in urine samples. Similar to blood samples, BP-3 (up to 603 μg/L) also reached the highest concentrations in urine, followed by lower concentrations for OMC (8 μg/L) and 4-MBC (7 μg/L) (Frederiksen et al., 2021; Janjua et al., 2008; Janjua et al., 2004; S. Lu et al., 2018; Zhang et al., 2013). Moreover, BP-3 was also found in adipose tissue from hospitalised individuals in Spain at up to 40.05 ng/g (Artacho-Cordón et al., 2018). One or more UV filters have also been found in breast tissue samples, mainly as OMC (up to 58.7 ng/g), BP-3 (up to 26 ng/g) and 4-MBC (up to 25.6 ng/g) (Barr et al., 2018). Importantly, BP-3 was also found in high concentrations in breast milk (799.9 ng/g) (Molins-Delgado et al., 2018). In addition, seminal fluid has shown BP-1 (up to 10.9 μg/L), BP-3 (up to 5.94 μg/L), 4-MBP (up to 16.03 μg/L) and 3-BC (2.01 μg/L) (Frederiksen et al., 2021).

5.1. Regulatory status

Nonylphenol ethoxylates are mainly used as detergents or emulsifiers in the manufacture of textiles. Since NPEOs are non-ionic surfactants, they are easily dissolved in water and can be washed out after repeated washing, regardless of the type of textiles. According to EU statistics on textile imports, the annual release of nonylphenol is estimated to be about 21–234 tonnes from textile washing, corresponding to 53–585 tonnes of NPEOs, as an average of 320 tonnes of NPEOs. Although nonylphenol enters the aquatic environment through the release of nonylphenol and NPEOs in wastewater from various sources, textile washing constitutes the main source and contributes to approximately 30%. The use of NPEOs alone or in combination has been restricted to a maximum concentration of 0.1% in the textile sector in the EU since 2005. However, most of the textiles used in the EU are imported from countries outside the EU. The EU risk assessment report states that in the worst-case scenario, 2.5% of the NPEOs discharged in WWTPs are then released into the environment in the form of nonylphenol. The restrictions proposed by the ECHA cover all types of textile materials, but only apply to textile articles that can be washed with water under normal conditions of use. A transition period of 5 years has been proposed to allow the market to adapt to the new conditions. The restriction is expected to reduce the average concentration of NPEO in textiles by about 69%, compared to the estimated 53 mg/kg textile in the baseline scenario. Other important sources that can contribute to the release of nonylphenol and NPEOs include detergents, plastic products, paints and adhesives. According to data from the Swedish Product Register, their annual contribution is estimated at 249 tonnes at the EU level. Many uses of nonylphenol and NPEOs are already restricted, most to < 0.1% by weight. These include detergents, manufacture of textiles and leather, teat rubbers (as emulsifiers), metal processing, pulp and paper manufacture, manufacture of cosmetics and other PCPs, and compounds in pesticides and biocides (4-Nonylphenol, Branched and Linear and 4-Nonylphenol, Branched and Linear, Ethoxylated).

5.2. Major sources

The most common industrial use of alkylphenols is in the manufacture of nonionic surfactants. Several alkylphenols are important in the production of stabilisers that allow coating and durability of plastic articles even at low concentrations. Alkylphenol derivatives can make resins more water resistant and more soluble in oil, and improve electrical properties (CEPAD). APEOs are prepared by reacting alkylphenols with ethylene oxide. The products are not pure, as a mixture of different APEOs is obtained, with varying numbers of ethyoxylate groups. Such ethoxylated derivatives are frequently used as nonionic surfactants in cleaning formulations in industry. APEOs have unique qualities, including good chemical stability and wetting, and emulsifying and detergent properties (CEPAD). Moreoever, they are used as dispersants in paper and pulp production, emulsifiers in latex paints and pesticide formulations, flotation agents, industrial cleaners (metal surfaces, textile processing, food industry), cold cleaners for cars and household cleaners (Acir & Guenther, 2018). In particular, most NPEOs have low surface tension, which makes them excellent wetting agents for rinsing, cleaning and dispersing purposes (Rijkswaterstaat, 2001). Accordingly, the annual demand for NPEOs was estimated to be about 170000 metric tonnes in the USA (Nonylphenol (NP) and Nonylphenol Ethoxylates (NPEs)), while in western Europe, it was estimated to be between 65000 and 80000 metric tonnes per year, although it must be taken into consideration that these data were published more than 20 years ago (Rijkswaterstaat, 2001).

5.3. Level of environmental contamination

Due to the intensive use of APEOs, they are frequently detected in effluents from WWTPs, and in surface waters, sediments and biota (see Table 7). The highest concentrations of NPEO (1850 µg/L) and OPEO (78 µg/L) were detected in influent water samples from WWTPs and in sludge samples (135, 12 µg/g, respectively) (Petrovic et al., 2001). In addition, sewage sludge in the USA contained 4-nonylphenol (534 µg/g), NPEO (62.1 µg/g ) and 4-NPE₂O (59.5 µg/g), while octylphenol was below the detection limit due to its less frequent use (Venkatesan & Halden, 2013). Surface waters contain lower concentrations than wastewaters. For example, seawater samples from Europe and the USA contained nonylphenol at up to 211 ng/L, NPEO up to 320 ng/L and OPEO up to 26 ng/L (Bester et al., 2001; Dupuis et al., 2012; Lee Ferguson et al., 2000; Li et al., 2008; Pojana et al., 2007). Unusually high concentrations have been reported for rivers in Germany, in terms of nonylphenol (up to 934 ng/L) and NPEOs (up to 1274 ng/L) (Jonkers et al., 2003), with alarmingly high concentrations in Korea of all nonylphenolic compounds in seawater (up to 36400 ng/L) and in stream waters (up to 3628 ng/L) (Li et al., 2008). Despite heavily polluted surface waters, drinking water contains much lower concentrations of alkylphenols and APEOs, with an average nonylphenol concentration of 27 ng/L (Fan et al., 2013). However, an older study detected 80–430 ng/L nonylphenol in treated water samples from drinking water treatment plants (Petrovic et al., 2003). In addition, sediments contained up to 192 ng/g nonylphenol (Bester et al., 2001; Pojana et al., 2007), while higher concentrations of NPEOs were detected in samples in Jamaica (up to 2120 ng/g) (Lee Ferguson et al., 2000) and Korea (up to 2811 ng/g) (Li et al., 2008). Of note, particularly high concentrations of nonylphenol have been detected in Tokyo Bay (up to 20.7 µg/g; Japan) (Kurihara et al., 2007).

Table 7

Occurrence of alkylphenol (ethoxylates) in waters, sediments and animals.
Sample and location	Alkylphenol	Concentration	Reference
Raw water samples from WWTP	NPEO	1850 µg/L	(Petrovic et al., 2001)
Raw water samples from WWTP	OPEO	78 µg/L
Sludge samples from WWTP	NPEO	135 µg/L
Sludge samples from WWTP	OPEO	12 µg/L
Influent wastewater (Italy)	NP	4.1 µg/L	(Bertanza et al., 2010)
	NP₁EO	3.9 µg/L
	NP₂EO	2.2 µg/L
Sewage sludge	4-NP	534 µg/g	(Venkatesan & Halden, 2013)
	NPEO	62.1 µg/g
	4-NPE₂O	59.5 µg/g
	OP	n.d.
Water samples (German Bight of the North Sea)	NP	0.7–4.4 ng/L	(Bester et al., 2001)
Water samples (German Bight of the North Sea)	NPEO	n.d.	(Bester et al., 2001)
Water (Venice lagoon)	NP	n.d.-211 ng/L	(Pojana et al., 2007)
Waters from the surface (Jamaica Bay (NY))	NPEO	157–320 ng/L	(Lee Ferguson et al., 2000)
Waters from the surface (Jamaica Bay (NY))	OPEO	1.04–26.0 ng/L	(Lee Ferguson et al., 2000)
Water samples (Rhine, Scheldt rivers)	NPE_1 − 2O	4.6–1029 ng/L	(Jonkers et al., 2003)
	NP_3 − 15O	37-1274 ng/L
	NP	31–934 ng/L
	NPEC	95-10879 ng/L
Creek water, seasurface water (Korea)	All nonylphenolic compounds	creek water: up to 3628 ng/L	(Li et al., 2008)
Creek water, seasurface water (Korea)	All nonylphenolic compounds	sea surface water: up to 36400 ng/L	(Li et al., 2008)
Surface and treated water (France)	NP	surface water: 13.5 ng/L − 124.9 ng/L,	(Dupuis et al., 2012)
Surface and treated water (France)	NP	treated water: LOQ − 59.4 ng/L	(Dupuis et al., 2012)
Drinking water (China)	NP	27 ng/L	(Fan et al., 2013)
Raw water samples from DWTP	NPEC	2.0-4.1µg/L	(Petrovic et al., 2003)
	NPE₂C	2.8–15µg/L
	NP	1.1–2.2 µg/L
	NPEO	0.7–1.8 µg/L
	NPE₂O	1.2–2.4 µg/L
Treated water samples from DWTP	NP	80–430 ng/L
Sediment samples (German Bight of the North Sea)	NP	10–153 ng/g	(Bester et al., 2001)
	NPE₁O	< 10 ng/g
	NPEO	< 10–39 ng/g
Sediments (Venice lagoon)	NP	47–192 ng/g	(Pojana et al., 2007)
Sediment (Korea)	All nonylphenolic compounds	up to 2811 ng/g	(Li et al., 2008)
Sediments (Tokyo bay)	NP	142-20700 ng/g	(Kurihara et al., 2007)
Sediment from the surface (Jamaica Bay (NY)	NPEO	110–2120 ng/g	(Lee Ferguson et al., 2000)
Sediment from the surface (Jamaica Bay (NY)	OPEO	5.96–60.7 ng/g	(Lee Ferguson et al., 2000)
Zooplankton and mussels (Gulf of Gdansk)	OP	zooplankton: 27.1 ng/g	(Staniszewska et al., 2014)
	OP	mussels: 6.4 ng/g
	NP	zooplankton: 263 ng/g
	NP	mussels 75 ng/g
Mussels (Venice lagoon)	NP	115–240 ng/g dry weight	(Pojana et al., 2007)
Mussels (Korea)	NP	50.5–289 ng/g dry weight	(Li et al., 2008)
Anchovy mucle (Adriatic sea, Italy)	NP	2.7–1285 ng/g	(Ferrara et al., 2005)
Tuna (Thyrrenian sea, Italy)	NP	695 ng/g	(Ferrara et al., 2008)
Sanddab Pacific ocean (California coast)	NP	2920 ng/g	(Diehl et al., 2012)
Fish (North Pacific Gyre and Callifornia Current)	OP	North Pacific Gyre: 0.66 ng/g wet weight	(Gassel & Rochman, 2019)
	OP	California Current 0.33 ng/g wet weight
	NP	North Pacific Gyre: 9.98 ng/g wet weight
	NP	California Current: 11.3 ng/g wet weight
n.d. not detected

As a direct result of water contamination, nonylphenol and octylphenol can accumulate in fish and other aquatic organisms. Nonylphenol has thus been found in mussels (50–289 ng/g) (Li et al., 2008; Pojana et al., 2007; Staniszewska et al., 2014), zooplankton (263 ng/g) (Staniszewska et al., 2014), tuna (up to 695 ng/g) (Ferrara et al., 2008) and anchovies (up to 1285 ng/g) (Ferrara et al., 2005), with the highest concentrations reported for sanddab (up to 2920 ng/g) (Diehl et al., 2012).

5.4. Fate in the environment

Nonylphenol has lipophilic properties (clogP, 4.28–4.51) and therefore tends to accumulate in sediments. Its t_½ varies significantly under aerobic conditions, with estimates from 1.1 to 99 days, due to many environmental factors; its t_½ under anaerobic conditions is from 23.9 to 69.3 days (Mao et al., 2012). Reports indicate that chlorination removes 60–74% of nonylphenol. This efficiency is practically independent of the reaction time, as 5 min appears to be sufficient (Noutsopoulos et al., 2013). Moreoever, the removal of nonylphenol in WWTPs in Serbia was shown to be from 75.8–99.9% (Bogunović et al., 2021). During wastewater treatment, APEOs are converted into various derivatives, usually as short-chain APEOs, such as mono- and diethoxylates (APE₁O, APE₂O), alkylphenols and alkylphenoxyacetic acids, including carboxyalkylphenol polyethoxycarboxylates and alkylphenol ethoxycarboxylates (Chiu et al., 2010). Biological-based treatment provided 85.5% of the mass loss, while ozonation provided only 41.3%, whereby the whole system removed 96.7–98.4% of all of the target compounds (Berardi et al., 2019). However, NPE₁O and NPE₂O were more resistant, because the removal efficiencies were low even at high chlorine concentrations: 14–37% for NPE₁O, and 22–52% for NPE₂O (Noutsopoulos et al., 2013). In addition to sewage treatment plants, constructed wetlands can remove 52.1–62.8% of such chemicals from water (Toro-Vélez et al., 2016).

5.5. Human exposure levels

The major routes of exposure to nonylphenol (ethoxylates) and other alkylphenols are inhalation of polluted air, ingestion of food and indoor dust, and drinking contaminated water. Cleaning agents present another possible route of exposure (Chokwe et al., 2017). Nonylphenol is rapidly absorbed (in rats, peak plasma concentrations are reached in 0.5-1.0 h), and are later metabolised mainly to a nonylphenol-glucuronide, and excreted predominantly in faeces (in rats, after about 48 h) (Green et al., 2003).

Octylphenol and nonylphenol have been detected in various human tissues and body fluids (see Table 8). As expected, APEOs were not found in these samples. Blood levels of nonylphenol (up to 9.38–15.85 µg/L) were generally higher than those of octylphenol (up to 0.74–9.1 µg/L), while data from a study in Korea revealed notable concentrations of nonylphenol and octylphenol, as up to 89.9 µg/L and 62.2 µg/L, respectively (Azzouz et al., 2016; Huang et al., 2014; Jung et al., 2013; Shen et al., 2016). Nonylphenol and octylphenol were also detected in urine samples, at up to 3.7 µg/L and 2.17 µg/L, respectively (Calafat et al., 2005; Park & Kim, 2017; Peng et al., 2016; Shen et al., 2016). Conversely, Pirard et al. (2012) did not detect nonylphenol in any urine samples, but as mentioned earlier, nonylphenol is predominantly excreted through faeces rather than urine. Moreoever, both nonylphenol and octylphenol were found in maternal blood (9.38, 5.46 µg/L, respectively) and amniotic fluid (8.44 5.72 µg/L, respectively) (Shekhar et al., 2017). Similarly, Huang et al. (2014) detected nonylphenol in maternal blood (14.6 µg/L), fetal cord blood (18.8 µg/L), placenta (19.8 ng/g), first month breast milk (23.5 µg/L) and third month breast milk (57.3 µg/L). Lower concentrations of nonylphenol were detected in the breast milk of mothers in Turkey (10.1 µg/L) (Sise & Uguz, 2017). The presence of alkylphenols in breast milk and cord blood indicates a source of exposure in infants. Indeed, both nonylphenol and octylphenol have been found in adipose tissue of children (38.3, 6.3 ng/g, respectively) (Olea et al., 2008). Nonylphenol (12.27 ng/g) was found in decidua, while in the chorionic villi, these concentrations were an order of magnitude higher (128.84 ng/g, respectively) (M. Chen et al., 2016). The most frequently found alkylphenol in human tissues was nonylphenol, as it is more commonly used than octylphenol. The highest concentrations of nonylphenol were found in blood, breast milk and chorionic villi.

Table 8

Occurrence of alkylphenol (ethoxylates) in human tissues.
Tissue	Alkylphenol	Concentration	Reference
Blood	OP	0.74 µg/L	(Shen et al., 2016)
	NP	15.85 µg/L	(Shen et al., 2016)
	OP	up to 9.1 µg/L	(Azzouz et al., 2016)
	NP	up to 12.0 µg/L	(Azzouz et al., 2016)
Serum	OP	up to 62.2 µg/L	(Jung et al., 2013)
Serum	NP	up to 89.9 µg/L	(Jung et al., 2013)
Maternal blood plasma	OP	5.46 µg/L	(Shekhar et al., 2017)
Maternal blood plasma	NP	9.38 µg/L
Amniotic fluid	OP	5.72 µg/L
Amniotic fluid	NP	8.44 µg/L
Maternal blood	NP	14.6 µg/L	(Y. F. Huang et al., 2014)
Fetal cord blood		18.8 µg/L
Placenta		19.8 ng/g
Breast milk in 1st month		23.5 µg/L
Breast milk in 3rd month		57.3 µg/L
Breast milk	NP	10.1 µg/L	(Sise & Uguz, 2017)
Adipose tissue of children	NP	38.3 ng/g	(Olea et al., 2008)
Adipose tissue of children	OP	6.3 ng/g	(Olea et al., 2008)
Decidua	NP	12.27 ng/g	(M. Chen et al., 2016)
Chorionic villi	NP	128.84 ng/g	(M. Chen et al., 2016)
Urine	NP	< 0.1 µg/L	(Calafat et al., 2005)
	NP	0.48 µg/L	(Peng et al., 2016)
	OP	2.17 µg/L	(Shen et al., 2016)
	NP	2.04 µg/L	(Shen et al., 2016)
	4-NP	3.7 µg/L	(Park & Kim, 2017)
	4-t-OP	0.6 µg/L	(Park & Kim, 2017)
	4-NP	n.d.	(Pirard et al., 2012)
n.d. not detected

The parabens, bisphenols, UV filters and APEOs are now ubiquitous, although their use in food, plastics and personal care products has become increasingly limited (e.g. lower permitted levels) due to the more stringent requirements of the competent regulatory agencies. The parabens MeP and PrP, the bisphenols BPA and BPS, the UV filters BP-3, OMC and 4-MBC, and the alkylphenol nonylphenol appear to be the most notable contaminants, as they have been found in the environment and in human tissues at significant concentrations due to primary and/or secondary exposure. As far as their fate in the environment is concerned, in general the short-chain parabens are relatively easily biodegradable, with half-lives of 10–35 h, nonylphenol is moderately persistent while bisphenols, in particular BPS (of the most frequently occurring representatives), tend to be more resistant to biodegradation, with half-lives ranging up to a year. The UV filters are structurally diverse hence their half-lives vary significantly from a few days up to a few months. Moreover, among the most notable pollutants, the UV filters BP-3, OMC, 4-MBC, and nonylphenol show the highest potential to bioaccumulate. Strikingly, other unapproved parabens (e.g., iPrP, iBuP, BzP) and UV filters (e.g., BP-2) have also been detected in surface waters and fish, and even in the urine of pregnant women, which should be of concern as their use has been banned on account of their endocrine disrupting properties. For these reasons, further comprehensive monitoring of environmental pollution and human exposure to ingredients of household and personal care products is of utmost importance in order to improve the risk assessment of HPCPs.

Funding

This work was supported by the Slovenian Research Agency (Grants P1-0208 and P1-0420). The authors thank Dr. Chris Berrie for proofreading the manuscript.

Competing interest

The authors declare that they have no competing financial interests.

Availability of data and materials

Not applicable.

Author Contributions

Conceptualization, Ž.J.; methodology, V.K. and Ž.J.; writing—original draft preparation, V.K. and Ž.J.; writing—review and editing, V.K., M.G. and Ž.J.; visualization, V.K., M.G. and Ž.J.; supervision, M.G. and Ž.J. All authors read and approved the final manuscript.

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Environmental contamination status with common ingredients of household and personal care products

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Abstract

Figures

1. Introduction

2. Parabens

2.1. Regulatory status

2.2. Major sources

2.3. Level of environmental contamination

2.4. Fate in the environment

2.5. Human exposure levels

3. Bisphenols

3.1. Regulatory status

3.2. Major sources

3.3. Level of environmental contamination

3.4. Fate in the environment

3.5. Human exposure levels

4. Uv Filters

4.1. Regulatory status

4.2. Major sources

4.3. Level of environmental contamination

4.4. Fate in the environment

4.5. Human exposure levels

5. Alkylphenol (Ethoxylates)

5.1. Regulatory status

5.2. Major sources

5.3. Level of environmental contamination

5.4. Fate in the environment

5.5. Human exposure levels

6. Conclusions

Declarations

References

Supplementary Files

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