Chapter 1 : Concentrations of organophosphate esters in drinking water from the United Kingdom: Implications for human exposure

 Concentrations of organophosphate esters in drinking water from the United Kingdom: Implications for human exposure

Abstract

Data on the presence of organophosphate esters (OPEs) in drinking water and its significance as a pathway of exposure are limited. In this study, we measure for the first time, concentrations of eight OPEs in 50 UK drinking water samples. Arithmetic mean concentrations of ∑8OPEs were: 6.4 and 11 ng/L in bottled (n = 25) and tap water samples (n = 25), respectively. Concentrations of ∑8OPEs in tap water (mean: 11 ng/L) exceed significantly those in bottled water (mean: 6.4 ng/L) (p ˂ 0.01). Moreover, UK tap water is more contaminated with chlorinated, aryl-, and alkyl-OPEs than bottled water. The predominant OPEs detected were: tris (butoxyethyl) phosphate (TBOEP), tris (2-chloroethyl) phosphate (TCEP), and tris(2-chloroisopropyl) phosphate (TCIPP) with arithmetic mean concentrations in the two water sample types ranging between (3.5–3.8 ng/L), (0.60–3.0 ng/L), and (1.02–2.9 ng/L), respectively. Estimated daily intakes (EDIs) (mean and high-end exposure) via drinking water for different sectors of the UK population were: infants (0.93 and 6.4 ng/kg bw/day) ˃ toddlers (0.46 and 3.1 ng/kg bw/day) ˃ children (0.35 and 2.3 ng/kg bw/day) ˃ adults (0.28 and 2.1 ng/kg bw/day). Based on these data, exposure to Σ8OPEs via drinking water is much lower than via: food, indoor dust ingestion, inhalation, and dermal uptake for adults and toddlers. Reassuringly, our EDIs were lower than relevant reference dose (RfD) values. However, combining our drinking water ingestion data with exposure via other pathways revealed overall exposure to 2-ethylhexyl diphenyl phosphate (EHDPP) and TCIPP to approach health-based limit values for UK toddlers under a high-end exposure scenario.

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Keywords:

Daily intakes;Tap water; Reference dose; and high-end exposure scenario.

1. Introduction

Organophosphate esters (OPEs) have been widely used as a flame retardant (FR) in various industrial and consumer products [1]. Other uses include as stabilisers, plasticisers, antifoaming, and wetting agents, as well as additives in lubricants and hydraulic fluids [2,3]. OPEs are used as additive FRs, meaning that they are not chemically bound to consumer products [4], thus these compounds can be released during production, use, and end of life product management by leaching, volatilisation, and abrasion [5]. Global usage of OPEs as a replacement for restricted BFRs has increased rapidly in recent years from 680,000 t in 2015 to 2,800,000 t in 2018 [1,6]. In western Europe, OPEs have been estimated to account for 20% of total FR consumption [3]. Such widespread and substantial use has led to the detection of OPEs in various environmental matrices including: food, water, air, dust, and sediment, as well as in human tissues such as: breast milk, placenta, blood serum, urine, hair, and nails [[7], [8], [9], [10], [11]].

Such evidence of human exposure is of potential concern given reports that OPEs can cause adverse effects. For instance, tri(2-chloroethyl) phosphate (TCEP) has been reported to decrease red blood cell cholinesterase activity, disrupt the thyroid endocrine system, and elicit neurotoxicity; while triphenyl phosphate (TPhP) has been linked to decreased red blood cell cholinesterase activity, along with neurotoxicity, contact allergenic effects, and impaired fertility. Moreover, tributoxyethyl phosphate (TBOEP) was associated with decreased red blood cell cholinesterase activity, and triethyl phosphate (TEP) linked with disruption of the thyroid endocrine system [2,3]. Also of concern, both TCEP and tris (1-chloro-2-propyl) phosphate (TCIPP) are potentially carcinogenic [3,12]. The presence of OPEs in human tissues demonstrates that human exposure occurs, but while research to date has highlighted that this arises via: dermal contact, dust ingestion, inhalation, and dietary intake [7]; relatively few studies have examined drinking water as a pathway of exposure to OPEs. Such studies are, hitherto, limited to: China [[13], [14], [15], [16], [17]], the United States [18,19], Spain [20], and South Korea [21,22]. Against this background, the current study reports concentrations of eight OPEs in 52 samples of drinking water collected in the UK, with the primary objective to assess the significance of drinking water as an exposure pathway of the UK population to OPEs. To the best of our knowledge, this is the first study of the occurrence of OPEs in UK drinking water, and as such will form a valuable baseline against which efforts to minimise this occurrence may be evaluated.

2. Materials and methods

2.1. Chemicals and materials

All chemicals, reagents and materials used in this study are provided as supplementary information (SI) (section 1).

2.2. Sample collection and preparation

Fifty drinking water samples comprising: tap water (n = 25), bottled water (n = 25) were collected from three major cities within the West Midlands conurbation of the UK, between July and August 2021. One litre tap water samples were collected from the kitchen of 25 different homes located within different conurbations (Birmingham, Walsall, and Coventry) in the UK West Midlands; while 25 bottled water samples consisted of single 500–1,500 mL polyethylene terephthalate bottles of popular brands purchased from major grocery stores within Birmingham. A detailed description of how drinking water samples were taken is provided in the supplementary information. Following collection, all samples were stored at 4 °C prior to extraction, which was conducted within 24 h of sample collection.

Approximately 200 mL of drinking water were spiked with 10 ng of isotopically-labelled internal (surrogate) standards (d27-TnBP and d15-TPHP) and extracted using an Oasis® HLB solid phase extraction cartridge (200 mg, 6 cm3; Waters). The cartridges were preconditioned by a sequence of 10 mL each of dichloromethane, methanol, and Milli-Q purified water. After sample loading, cartridges were dried under a gentle nitrogen gas stream, and eluted with 6 mL of dichloromethane. The eluent was concentrated under a gentle stream of nitrogen to incipient dryness. This was reconstituted with 100 μL of iso-octane containing 10 ng of PCB-62 as recovery determination (syringe) standard (RDS). The final sample concentrate was transferred to an amber vial prior to analysis on an Agilent 5975C GC-MS operated in selected ion monitoring electron ionisation mode and fitted with a 30 m DB-5 MS GC column (0.25 mm ID, 0.25 μm film thickness) (Restek, USA). Detailed information on the instrumental analysis conditions is provided as supplementary information (SI, section 2).

2.3. Quality assurance and quality control

All laboratory glassware was washed, rinsed with deionised, distilled Milli-Q water, heated at 460 °C for at least 2 h, then rinsed sequentially with hexane, acetone, and dichloromethane prior to use. For all target analytes, the relative standard deviation of the relative response factors (RRFs) in the five calibration standards (0.05–0.75 ng/μL) were below 8%. Two procedural blanks (n = 2) were included for every batch of five samples with only TCEP detected in blanks at an average concentration of 0.10 ± 0.08 ng/L (Table S3). Concentrations of TCEP in each batch of samples were therefore blank-corrected (by subtracting the average concentrations in the two procedural blanks detected with the samples from each batch). In the absence of an appropriate certified reference material, matrix spiked samples (Milli-Q water) (n = 5) containing 10 ng of each target OPE (equivalent to 50 ng/L in the water sample) were analysed to evaluate method performance. Recoveries of our eight target OPEs in spiked samples ranged from 67 to 123% (Table S3). Moreover, recoveries of the two internal standards were 92 ± 17% and 89 ± 11% for d27TnBP and d15TPHP respectively (Table S3). The limit of detection (LOD) and the limit of quantification (LOQ) were calculated as the amounts of an analyte that yielded signal to noise ratios (S/N) of 3 and 10 respectively based on 11 injections of the lowest concentration calibration standard (0.05 ng/μL, Table S3). For exposure assessment purposes, OPE concentrations below the LOQ were assumed to be present at either zero (lower bound (LB)) or the LOQ (upper bound (UB)).

2.4. Statistical analysis

Descriptive and multivariate statistical analyses were performed using IBM SPSS Statistics 28 (USA) for Windows and Microsoft Excel 365. The data were log10 transformed prior to analysis after a Shapiro-Wilk test showed that the data were not normally distributed. Such data transformation is necessary due to the sensitivity of principal component analysis (PCA) to non-uniformly distributed data [23]. A t-test was used to investigate significant differences in OPE (chlorinated, aryl and alkyl OPEs) concentrations between tap and bottled water. The Spearman rank correlation coefficient (r) is used to reflect the linear correlation between the analytes as well as the correlation between OPE concentrations in bottled and tap water, while PCA was carried out after fulfilling the condition of data to be normally distributed. PCA was performed to investigate the possible factors driving OPE concentrations in the three drinking water sample types. This was carried out based on varimax orthogonal rotation with eigen-value ˃ 1 after determining the Kaiser-Meyer OIkin (KMO) test that measured sampling adequacy and Bartlett's Test of sphericity was adequate and found to be significant (p˂0.01) for the variables [23]. The first four principal components (PCs) with (eigen-values ˃ 1) which explained ˃ 75% of the total variance were retained as the most significant components.

2.5. Exposure and risk estimation

The estimated daily intake (EDI) expressed as nanograms per kilogram of body weight per day of OPEs via ingestion of drinking water was calculated for infants, toddlers, children, and adults using the following equation (1)where water is the concentrations of a given OPE or combination of OPEs in drinking water (ng/L), IR is the average drinking water ingestion rate (L (kg of bw)−1day−1) which is 0.05 L (kg of bw)−1day−1 for infants, 0.026 L (kg of bw)−1day−1 for toddlers, 0.02 L (kg of bw)−1day−1 for children and 0.016 L (kg of bw)−1 day−1 for adults [24]. Both a “normal” exposure scenario (assuming the water consumed was contaminated at the arithmetic mean OPE concentration) and a high-end exposure scenario (assuming the water consumed was contaminated at the 95th percentile OPE concentration) were evaluated for each of the age groups considered.

3. Results and discussion

3.1. Concentration of OPEs in drinking water

A statistical summary of concentrations of our target OPEs are shown as Table 1. Similar to previous studies [14,[17], [18], [19], [20], [21], [22],[25], [26], [27], [28], [29]], TCEP was detected in all the studied drinking water samples (detection frequency (DF) = 100%). In contrast, TCIPP, TDCIPP, TPHP and EHDPP displayed varying DFs ranging from 44 to 96% in tap water and 16–92% in bottled water (Table 1). TnBP displayed the lowest DF, ranging from 0% in bottled water to 16% in tap water. With respect to TMTP and TBOEP, DFs were between 28 - 72% and 20–72%, in tap water and bottled water, respectively (Table 1). The higher DFs observed for TCEP, TCIPP, TPHP, and TBOEP may be associated with the wide past and current application of these OPEs in the UK [21,26], as well as their physicochemical properties. Both TCEP and TCIPP have been used widely in: polyurethane carpet backing, furniture foam, and polystyrene building insulation [3]. Meanwhile, TPHP is used as an additive in food packaging materials [30], while TBOEP is widely used in floor polish/wax and as a plasticiser in vinyl plastics and rubber stoppers [3,31]. Moreover, TCEP, TCIPP, and TBOEP have relatively higher water solubilities (exceeding 1000 mg/L at 25 °C) in comparison to other OPEs (Table S1). Average concentrations of target individual OPEs across the two drinking water categories examined (bottled and tap water) were: TBOEP (7.3 ng/L) ˃ TCIPP (3.9 ng/L) ˃ TCEP (3.6 ng/L) ˃ TPHP (1.2 ng/L) ˃ TnBP (0.66 ng/L) ˃ EHDPP (0.39 ng/L) ˃ TMTP (0.25 ng/L) ˃ TDCIPP (0.11 ng/L) accounting for 42%, 23%, 21%, 7%, 4%, 2%, 1%, and 0.7% of ∑8OPEs, respectively. Thus, TCEP and TCIPP were the predominant chlorinated OPEs (Cl-OPEs), TPHP and EHDPP the predominant aryl-OPEs, while TBOEP was the dominant alkyl-OPE (Fig. 1).

Table 1. Statistical summary of OPE concentrations (ng/L) and detection frequency (DF, %) in the three drinking water categories examined in our study.

Several studies have reported higher concentrations of TCEP, TCIPP, and TBOEP in drinking water samples than in our study; specifically in China [[13], [14], [15], [16], [17],25,28,29], in Korea [21,22], and the USA [19]. The mean (95th percentile) concentrations of the chlorinated OPEs, TCEP and TCIPP in UK tap water (TCEP: 3.0 (7.3 ng/L), TCIPP: 2.9 (960 ng/L)) and bottled water (TCEP: 0.60 (1.6 ng/L), TCIPP: 1.0 (1.8 ng/L)) in our study; were lower than the average concentrations reported for drinking water from China (TCEP: 38.8 ng/L; TCIPP: 67.0 ng/L) [21], (TCEP: 9.1 ng/L; TCIPP: 6.7 ng/L) (Ding et al., 2015), (TCEP: 18.7 ng/L; TCIPP: 20.0 ng/L) [28], (TCEP: 27.8 ng/L; TCIPP: 218 ng/L) (Huang et al., 2022), in Korea (TCEP: 39.5; TCIPP; 49.4 ng/L) [22], in US (TCEP: 150 ng/L) [18] and TCIPP (40 ng/L) in drinking water from Spain (Rodil et al., 2012) (Table 2). However, our mean concentrations for the aryl OPEs: TPHP (0.64 and 0.59, ng/L) and EHDPP (0.23 and 0.16) in tap and bottled water, respectively (Table 1) were comparable to those reported in China for TPHP (0.14 and 0.28 ng/L) for ambient temperature water and hot water, respectively [17] but lower than the values reported in other studies from China (TPHP: 21.3 ng/L) [28], (TPHP: 1.11 ng/L) (Huang et al., 2022). Meanwhile, our arithmetic mean concentrations of TBOEP (3.5 and 3.8 ng/L) in tap and bottled water were comparable to those reported in Eastern China [14] and below that reported in eight major metropolitan cities in China (26.1 ng/L) [21]. Among our target OPEs, TBOEP was present at the highest average concentration in: bottled water (3.8 ng/L) and tap water (3.5 ng/L) followed by TCEP with a mean concentration of 3.0 ng/L in tap water and 0.60 ng/L in bottled water (Table 1).

Table 2. Comparison of arithmetic mean (unless indicated otherwise) concentrations (ng/L) of OPEs in drinking water measured in this study with mean concentrations reported from other countries.