Organophosphate Esters in Children and Adolescents in Liuzhou City, China: Concentrations, Exposure Assessment and Predictors

Meng Yu Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Xiang Li Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Bingqing Liu Women's Hospital School of Medicine Zhejiang University Yaping Li Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Ling Liu Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Limei Wang Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Lulu Song Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Youjie Wang Tongji Medical College of Huazhong University of Science and Technology: Huazhong University of Science and Technology Tongji Medical College Liqin Hu Wuhan Childrens Hospital: Wuhan Women and Children Medical Care Center Surong Mei (  surongmei@hust.edu.cn ) Huazhong University of Science and Technology Tongji Medical College


Introduction
Organophosphate esters (OPEs) are increasingly used as ame retardants and plasticizers in consumer and industrial products due to the phase-out of some polybrominated diphenyl ethers (PBDEs). It was reported that phosphorous and brominated ame retardants accounted for 11% and 23% of the total global consumption of ame retardants in 2008, respectively. However, the percentage contribution of phosphorous ame retardants increased to 30%, and brominated ame retardants attenuated to 20% in 2017 (Wang et al., 2020). Because OPEs are frequently physically mixed with polymers, they can easily released from treated products into the external environment, leading to a frequent detection of OPEs in air , sediments and waters , foodstuffs (Li et al., 2019a), and biota (Pantelaki and Voutsa, 2020). Concerns regarding these replacement OPEs have mounted, because toxicological studies suggested that some OPEs are considered endocrine disruptors (Liu et al., 2012), and carcinogenic- (Böckers et al., 2020), neurodevelopment- (Rock et al., 2020), and reproductive toxicants (Xu et al., 2017b).
Humans can be exposed to OPEs through different pathways, including dust ingestion and inhalation, dermal absorption, and dietary intake (Cequier et al., 2015;Liu et al., 2017). After being absorbed, OPEs could be rapidly metabolized (several hours to days) and excreted into the urine (Hou et al., 2016). Urinary OPE metabolite concentrations re ect integrated exposure from different pathways and have been extensively recognized as biomarkers of exposure to OPEs. A growing body of studies have estimated the total daily intake of OPEs through the monitoring of urinary OPE metabolite concentrations in the general population (Zhang et  nondietary exposure, which was considered one of the most important exposure pathways. To identify individual's exposure from indoor microenvironments, hand wipes could be an effective metric because it re ects the exposure from dermal contact with dust and other media (e.g., indoor consumer products) (Larsson et al., 2018). Similar dermal exposure estimates were obtained for several halogenated ame retardants (HFRs) when the exposure assessment was based on direct measurements from hand wipes or the indirect measurement from indoor dust (Tay et al., 2018). Human dermal absorption of OPEs could be estimated using the mass of OPEs in hand wipes, which has been reported in biomonitoring studies Xu et al., 2016).
Although several studies have previously suggested that dermal contact with dust and inhalation are two more important pathways (He et al., 2018c;Wei et al., 2015;Xu et al., 2016), there is increasing concern regarding the occurrence of OPEs in foods, and the importance of diet as an exposure source of OPEs has been studied (Li et al., 2019a). Zhang et al. (2016) monitored the residue levels of OPEs in 75 commonly consumed foods in China and detected OPEs in 100% of the food samples. He and Wang (2018) collected 87 food samples from Southeast Queensland, Australia, and found that OPEs and their metabolites were both frequently detected in various foodstuffs. However, the associations between dietary factors and the body burden of OPEs in humans are unclear. In a Norwegian adult cohort, Xu et al. (2017a) detected OPEs in duplicate diet samples (n=61) and found that meat intake was the primary dietary exposure source of OPEs. A study conducted in eastern China showed that postpartum women (n=50) who consumed more vegetables before pregnancy had a higher tris (2-chloroisopropyl) phosphate (TCIPP) level in placentas than those who consumed few vegetables (Ding et al., 2016). A study of toddlers aged 15-18 months in the USA (n=41), using dietary survey data, reported that meat and sh consumption may be linked with increased bis(1,3-dichloro-2-propyl) phosphate (BDCIPP) and diphenyl phosphate (DPHP) concentrations in urine (Thomas et al., 2017).
Nevertheless, research using a larger sample size is warranted to verify the previous results.
Children and adolescents have more rapid nutrient absorption, metabolic and ventilation rates, and larger ratio of surface area to body weight as compared to adults (Van den Eede et al., 2015), thus, the body burden of OPEs in children and adolescents may be higher. The monitoring surveys in China suggested that children and adolescents are ubiquitously exposed to OPEs, especially chlorinated alkyl phosphates Ding et al., 2019). Childhood and adolescence are in the critical periods of development, and developing organ systems are more susceptible to chemical exposure, including OPEs (Scheuplein et al., 2002). Considering increasing exposure to OPEs and their potential health risks, determining the factors that may impact OPE exposure in children and adolescents has been of growing concern. However, the related studies are very limited, especially in

Dietary assessment
Information on dietary factors was obtained using food frequency questionnaires, which is frequently used in large sampling campaigns. The questionnaire was provided to obtain the dietary patterns for the study participants, and they completed the questionnaire with their parents and were guided by trained researchers. In the questionnaire, seven frequently consumed food groups were administrated in the food items, including four foods of animal origin (meat, sh, eggs, and dairy products) and three foods of plant origin (vegetables, fruits, and beans). The consumption rate (per week) of the seven food groups is detailed in Table S1 and was categorized into tertiles for further analysis (Table 1). Additionally, the daily source of drinking water was recorded, and a four-category option was set in this item, including tap water, barrelled water, puri ed water, and others (e.g., well water).

Sample collection
During the school visit, a 5 mL polypropylene container marked with a unique identi cation number was given to each participant for spot urine collection. The collected urine samples were placed in carton and transferred to the laboratory.
The collection of hand wipes was performed as described in our previous study (Tao et al., 2018). Brie y, a sterile gauze pad (6×8 cm) soaked in 5 mL of isopropanol was used to wipe the palm and back of each subject's hands. Subsequently, the collected hand wipe samples were enclosed in 50 mL polystyrene centrifuge tubes. Meanwhile, eld blank samples were collected from all the schools.

Sample analyses
Prior to urine analysis, speci c gravity (SG) was detected using refractometer (Atago PAL-10S, Tokyo, Japan), and urinary OPE metabolites were adjusted for dilution before statistical analyses. Eight urinary OPE metabolites, including tris (

Exposure assessment
Estimated daily intakes (EDIs) were assessed using urinary OPE metabolite concentrations to determine the total daily intake of OPEs in children and adolescents. The calculation for EDIs (ng/kg bw/day) was described as: where C mOPEs (ng/mL) is the SG-corrected concentrations of individual OPE metabolites in case of this study; V urine (mL/day) are the molecular weights of the parent compound and its metabolite respectively. The parameters for estimating EDI were showed in Table S2-S3.
Dermal exposure dose (DED, ng/kg bw/day) of OPEs was estimated based on the amount of OPEs in hand wipes and was calculated as: where C S (ng/m 2 ) is individual OPE concentration on hand surface; SA (m 2 ) is the exposed surface area (USEPA, 2011); AF (unitless) is absorption factor adopted from human ex vivo skin experiments (Abou-Elwafa Abdallah et al., 2016); ED (h day −1 ) is the exposure time, which was assumed as 24 h; and bw (kg) is the body weight. The parameters of SA among different age group and AF of individual OPEs were showed in Table S4-S5.

Statistical analyses
Extracted samples with OPE concentrations below the LOD were substituted with LOD/√2. Compounds with a detection frequency >50% were included in all statistical analyses, whereas we did not analyze TEHP in hand wipes due to the greatly varied concentrations of TEHP in the QC samples. The concentrations of urinary OPE metabolites were natural log-transformed due to the lack of normally distributed. Spearman correlation coe cients (r s ) were used to determine the associations between OPE metabolites in urine and OPEs in hand wipes.
Multivariate linear regression model was applied to investigate the potential predictors of urinary OPE metabolite concentrations.
The following variables were included in the models: subject's sex and age (categorical variable), maternal education, the source of drinking water, the consumption frequency (per week) of seven food groups (meat, sh, eggs, dairy products, fruits, vegetables, and beans), and the concentrations of OPEs in hand wipes. OPE metabolites in urine and their corresponding parent compound in hand wipes were simultaneously entered into multivariable models, as described in Fig S2. Age was categorized as 6-11 years, 12-15 years, or 16-18 years. Maternal education was categorized as less than high school, high school, or college or above. The hand wipe levels were categorized into quartiles, except for TCEP and TBOEP. TCEP and TBOEP were categorized into tertiles and dichotomy, respectively, due to their low detection frequencies. Additionally, we tested multicollinearity using the variance in ation factor (VIF), with all VIFs <5. To interpret the regression models containing ln-transformed outcome variables, the regression coe cient (β) and 95% con dence interval (CI) was converted to percent change (%change) and 95% CI in urinary OPE metabolite concentrations.
All analyses were carried out using STATA 12.0 (StataCorp, College Station, TX, USA). A two-sided P < 0.05 was considered signi cant.

Study population
Of participants, 46.4% of children and adolescents were male and 53.6% were female, with a mean age of 12.5 ± 3.2 years old ( Table 1). A total of 51.9% of mothers attained less than a high school degree, 24.7% attained equal to a high school degree, and 23.4% participated in college. The majority of food groups was generally categorized into tertiles, except for vegetable intake. A total of 49.9% of the participants were divided into the third tertile due to their high proportion of seven meals per week for vegetable consumption. The source of drinking water was mainly tap water (63.1%), followed by barreled water (22.1%), puri ed water (14.0%), and others (0.8%).  Note: SG, speci c gravity; LOD, limit of detection; GM, geometric mean; Max, maximum.

Concentrations of urinary OPE metabolites
Median concentrations of urinary OPE metabolites in several studies were compared in Table S6. BCIPHIPP, BDCIPP, and DPHP are commonly analyzed urinary biomarkers of exposure to OPEs and are frequently detected in urine samples around the world. inhalation of TCEP was considered the greatest contributor of daily intake for TCEP (Xu et al., 2017a). Due to limited data available from multiple environmental compartments, whether a higher body burden of TCEP in children and adolescents of Liuzhou city is attributed to a higher contamination of TCEP in indoor air remains to be con rmed.

Concentrations of OPEs in hand wipes
The concentration distributions of OPEs in hand wipes from children and adolescents were shown in Table 2. Seven OPEs were detected in >50% of the hand wipe samples. EHDPP and TCIPP were detected at the highest median concentrations of 10.9 ng/wipe and 9.49 ng/wipe, respectively.
Concentrations of OPEs in hand wipes from various studies were summarized in Table S7.

Exposure assessment
To investigate the total daily intake of OPEs, EDIs of individual OPEs were calculated and displayed in Table 3  To investigate the contribution of dermal exposure to OPEs, the OPE intake via skin wipes and its proportion to total daily intake were estimated (

Correlation between urine and hand wipes
We found that the TPHP and EHDPP levels in hand wipes were signi cantly correlated with their metabolite DPHP in urine (r s = 0.068-0.142, P < 0.05), whereas no signi cant correlations were observed between other OPEs in hand wipes and the corresponding urinary metabolites (Table 4)  In the present study, OPE intake via dermal absorption could partially explain the body burden of OPEs. To determine the potential predictors for urinary OPE metabolite concentrations, factors (sociodemographic variables, dietary patterns, and OPE levels in hand wipes) that may impact OPE exposure were all included in multivariable linear regression models (Table 5). After adjusting for sex, age, maternal education, and dietary patterns, a strong correlation continued to be found for EHDPP in the highest quartile in hand wipes and urinary DPHP (P < 0.001). Although the association between TPHP and DPHP was no longer signi cant, there was a suggestive dose response relationship between TPHP and DPHP (Fig. 1). Similarly, the levels of other OPEs in hand wipes were not related to their corresponding urinary metabolites in the multivariable models. These results implied that hand wipes could be an effective exposure metric for TPHP and EHDPP, whereas the exposure of other OPEs may be linked to additional predictors. Furthermore, we discussed the associations of sociodemographic and dietary variables with urinary OPE levels after adjusted for other variables in Table 5.  ., 2019). The concentration of urinary TCEP were high in males, while urinary BBOEP was high in females, indicating differences in metabolism and exposure patterns between genders (Hoffman et al., 2015). Maternal education level was a proxy for socioeconomic status that was positively associated with the urinary BBOEP in our participants.
A potential explanation of this result is that newer furniture and electronics entered into these highly educated mothers' households, which led to the transition of indoor exposure patterns from PBDEs to OPEs (Percy et al., 2020  Tap water is the most common and cheapest drinking water in China and contains abundant OPEs, particularly halogeno-OPEs, compared to other water sources (e.g., barreled water, bottled water, puri ed water, and well water) (Ding et al., 2015). A possible hypothesis is that puri ed water, barreled water, and other drinking water may be associated with lower concentrations of urinary OPE metabolites. Due to the limited number of participants in the other drinking water group, the result that other drinking water was associated with a lower TCEP concentration in urine should be interpreted with caution. However, it is unexpected that barreled water drinking was associated with an increase in urinary BCIPHIPP. Li et al. (2019c) analyzed the concentrations of OPEs in tap water samples from 79 cities in China, in which the contamination of OPEs in tap water from Liuzhou (the sample site of our study) was not severe. In addition, most barreled water-producing manufacturers are regional, and different manufacturing techniques are used according to the enterprise scale and cost. In the sample site of our study, regional barreled water could be a potential source of exposure to OPEs for residents. Considering the limited data on drinking water around the world, more research on different types of drinking water is needed.

Limitation and strength
The main strengths of our study include the large sample size for an exposure study of OPEs and the focus on the vulnerable population of children and adolescents. This is a relatively comprehensive study that examines the relationship between urinary OPE metabolites and dietary factors. The dietary factors include the consumption rate of seven common food groups and the sources of daily drinking water. Additionally, the OPE levels in hand wipes were included in the multivariable linear regression models given that dermal contact with dust is commonly considered an important exposure pathway for humans.  ). Therefore, the suitability of usage of organophosphate diesters as metabolites of OPEs are needed further assessment. Second, given that other newly discovered OPEs exhibit great detection frequencies and are abundant in multiple environmental matrices (e.g., bis-(2-ethylhexyl)phenyl phosphate (BEHPP), tris(2,4-di-tertbutylphenyl) phosphate (TDTBPP)), these novel OPEs should be monitored in the forthcoming studies. Third, we could not examine long-term exposure to OPEs using a cross-sectional study, and many of OPEs are rapidly metabolized in the human body; thus, a spot urine sample could not be representative of the average body burden of our participants. Fourth, single hand wipe sample could be a potential weakness for OPE measurements, as it is unknown how OPE levels vary in a single person over time. Herein, we measured OPEs in urine and hand wipe samples, whereas the inclusion of other exposure metrics (e.g., passive air samples and indoor dust) may provide additional information on exposure to OPEs. Last, in the present study, food items combined with food frequency questionnaires were used to estimate human dietary intake

Conclusion
Our study indicated the widespread exposure to OPEs in children and adolescents. OPE intake via skin wipes could be an important contributor to the total daily of OPEs. Additionally, dietary intake is also a signi cant exposure source of certain OPEs, particularly vegetable consumption and barreled water drinking. More research is warranted to verify our preliminary ndings and evaluate the potential health risks from chronic dietary exposure to OPEs.

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