Introduction

Phthalates esters (PAEs) or esters of phthalic acid (1,2-benzenedicarboxylic acid) have been commonly used as plasticizers to increase flexibility, transparency, and durability of plastic materials since the 1930s (Farahani et al. 2008; Tsumura et al. 2002). PAEs are produced all over the world around 6.0 million metric tons per year (Jia et al. 2014). Generally, dimethyl phthalate (DMP), diethyl phthalate (DEP), dipropyl phthalate (DPrP), dibutyl phthalate (DBP), and di-2-ethylhexyl phthalate (DEHP) are commonly used plasticizers. DEHP contributed about 54% market share in the year 2010 and was categorized as a high manufacturing chemical (Market Study: Plasticizers 2011; Gallart-Ayala et al. 2013).

PAEs are considered to be potential mutagenic, carcinogenic, and endocrine disrupters, with fetal animals being particularly sensitive (Chopra et al. 2014; Favrat et al. 2014). Although the intake of phthalates may originate from many sources, there is special interest in monitoring the contamination of water, especially for human beings (Schulpen et al. 2014; Trasande et al. 2013; Völkel et al. 2014).

In general, the contents of PAEs in plastics materials, such as polyvinyl chloride (PVC), polyethylene terephthalate (PET), polyvinyl acetates (PVA), and polyethylene (PE), vary from 10 to 60% by weight (IARC 2000). The quantity of plasticizer depends on the required flexibility for plastic usage. The extent of the plasticizer and more prominently its bulk (taking into consideration the existence of branches and side chains) will affect mutually the efficiency and the temperature behavior of the polymer (Amato et al. 2001). Mostly, these plastic materials are used for packaging and processing and for different household wares such as water coolers, water bottles, and even baby feeders. Since PAEs are not chemically bound to the plastic matrix (Matsumoto et al. 2008; Staples et al. 1997), they tend to migrate into drinking water from plastic containers like coolers, bottles mostly of mineral water, and baby feeders and they also become part of air, groundwater, and soil. As a result of such sources of PAE contamination, they move into the food chain if absorbed by plants and animals (Castillo and Barceló 2001). In addition, blood bags and medical devices are also alarming sources of PAE pollution and human health at risk due to the presence of PAEs in life-saving resources (Tickner et al. 2001; Shaz et al. 2011; Simmchen et al. 2012; Bernard et al. 2014).

It is common observation in our country that water coolers are used for drinking water storage purposes and mostly available at home, schools, and hotels. Most probably, water coolers are used in highway side hotels and kept in an open atmosphere on benches exposed to sun radiations. On the other hand, stocks of mineral water bottles are also kept outside shops for several days due to lack of cooling facilities and storage space. Some of the sun-exposed bottles are subsequently placed in fridges for selling on customer demand (Dawn News. 2015). These are alarming sources for leaching of PAEs in drinking water, which have a negative impact on human beings. Sterilization of baby feeders is a normal practice of parents especially mothers to save their children from a more serious bout of food-borne illness (vomiting and diarrhea), and to produce a germ-free environment for their babies. Sterilizing babies’ feeding equipment can help to keep them healthy (Naomi 2014), but it is also a risk factor towards leaching physically bound PAEs into water, milk, etc.

The above discussed facts are potential sources of PAE toxicities in drinking water and tend to be removed, which can pose toxic impacts and hence require elimination from water bodies. PAEs can be removed or degraded by various natural processes such as hydrolysis and photolysis (Surhio et al. 2014), but unluckily these processes have been reported to be slow and insignificant (Staples et al. 1997). However, the metabolic breakdown of PAEs by bacteria plays a major role in the environmental degradation of PAEs (Jianlong et al. 1999). Previously, several authors have reported the leaching of water-soluble PAEs from plastic containers into soft drinks and mineral water (Bošnir et al. 2007; Erythropel et al. 2014), household waste (Bauer and Herrmann 1997), food packaging (Muncke 2009), bottled water (Saeed et al. 2010), and toys (Wilkinson and Lamb 1999). In addition, aqueous leaching of DEHP from PVC blends has also been reported (Kastner et al. 2012).

Bacteriological breakdown is the major route of removal of PAE isomer and their esters from aquatic and terrestrial systems such as sediments, soils, sewage, and surface waters (Staples et al. 1997). The presence of microbes significantly advances the biodegradation of the PAEs, which prevails as the leading alteration route within the upper landfill layers (Ejlertsson et al. 1996; Staples et al. 1997; Peterson and Staples 2003; Di Gennaro et al. 2005). The PAE biodegradation has been broadly studied and well recognized (Elder and Kelly 1994; Kleerebezem et al. 1999; Jianlong et al. 2000; Juneson et al. 2001; Cousins et al. 2003; Liu and Chi 2003; Chang et al. 2004). Mostly, the PAE degradation pathway involves primary biodegradation from dialkyl phthalate to monoalkyl phthalate and then to phthalic acid and eventual biodegradation of phthalic acid to CO2 and/or CH4 (Jianlong et al. 2000). This degradation procedure happens under both aerobic and anaerobic conditions (Cousins et al. 2003).

Therefore, the purpose of this study was to describe the leaching concentration of PAEs in drinking stuffs, i.e., mineral water bottles, water coolers, and baby feeders. Sunlight-exposed samples of water coolers and mineral water bottles were addressed, as in our region the temperature fall is up to 46–48 °C in peak summer timing (Sharif et al. 2015). Furthermore, to evaluate the potential leaching of plasticizers from baby feeders, different house-based procedures were followed such as boiling, electric steam sterilizing, and microwaving. Moreover, for the first time, this study evaluated the degradation of leached PAEs from water with the biochemical cooperation of novel Bacillus thuringiensis (B. thuringiensis) strain. Detection and quantification of leached PAE mixture were carried out by GC-FID and LC-MS/MS techniques. The present study provides sufficient knowledge of leached PAE exposure from water coolers, mineral water bottles, and baby feeders under common conditions of use and potential application of bioremediation for these leached PAEs from drinking water.

Materials and methods

Reagents and samples

DMP, bis(2-methoxyethyl) phthalate (DMEP), bis(2-ethoxyethyl) phthalate (DEEP), DEP, DBP, diallyl phthalate (DAP), benzyl butyl phthalate (BBP), bis(2-ethylhexyl)adipate (DEeP), and DEHP, all with 99% analytical standards, were obtained from Scharlau (Barcelona, Spain). DPrP of 98% purity was purchased from Sigma-Aldrich (Steinheim, Germany). Luria-Bertani (LB) medium was procured from Caisson Laboratories, Inc. (Utah, USA). n-Hexane was obtained from Sigma-Aldrich (Steinheim, Germany). All other chemical reagents used were of analytical grade.

Water cooler samples, different brands of mineral water bottles, and baby feeders were purchased from local stores of Hyderabad, Sindh, Pakistan, in the peak summer season (temperature ranges 38–48 °C, sunlight 10–12 h day−1). Special care was taken to avoid contact of solvents and reagents with plastic materials. To minimize the risk of secondary contamination, glass materials were used in place of plastic materials. All glassware was consecutively rinsed with ethyl acetate, iso-octane, and purified water twice each before use according to the recommendations specified in US EPA Method 506. All solvents were checked for the presence of phthalates before use. Water samples serving as blanks were taken directly from the well of bottling companies for each brand into 1.5-L glass bottles, and PAE levels were measured and used as the baseline values for blank controls prior to the leaching experiment.

Leaching study

Leaching study of phthalate esters from different samples was divided into two parts.

Mineral water bottles and water coolers

Water coolers and mineral water samples were kept in an open environment treated with exposure to sunlight (May 2016) for a period of 7 days (10 h day−1; 8:00 a.m. to 6:00 p.m.). After completion of the leaching period, water was drawn from bottles for analyzing PAEs leached and their subsequent biodegradation.

Baby feeders

Baby feeders underwent three types of potential leaching practices like boiling, autoclaving (Philips Avent 3-in-1 electric steam sterilizer SCF284/01), and microwave oven sterilization (Super Asia Microwave Oven SM-124-D Digital GRTII-AUTO). Boiling of baby feeders was carried out in a 2-L beaker for 15 to 20 min (min). Electric steam sterilization (autoclave) was carried out for 20 min, and baby feeders were oven sterilized in microwave for 1.5 min.

Preparation of culture medium for degradation of leached phthalate esters

The mineral salt medium (MSM) for bacterial culture growth consisted of (NH4)2SO4 1 g L−1, KH2PO4 0.8 g L−1, K2HPO4 0.2 g L−1, MgSO4·7H2O 0.5 g L−1, FeSO4 0.01 g L−1, CaCl2 0.05 g L−1, NiSO4 0.032 g L−1, Na2B4O7·10H2O 0.021 g L−1, (NH4)6Mo7O24·H2O 0.0144 g L−1, ZnCl2 0.023 g L−1, CoCl2·H2O 0.021 g L−1, CuCl2·2H2O 0.01 g L−1, and MnCl2·4H2O 0.03 g L−1. The pH of MSM was adjusted to 7.0 with 0.1 M HCl or NaOH.

The agar plates were prepared by adding Luria-Bertani (LB) medium 20 g L−1 and powdered agar 15 g L−1. Minimal salt medium at pH 7.0 was prepared as per our reported method (Surhio et al. 2014). The mineral salt medium (MSM) was used for the degradation tests. A leached PAE sample solution was filter sterilized through a 0.22-mm membrane before use. All the other media and tools were sterilized in an autoclave at 121 °C for 25 min.

Biodegradation experiment of leached PAEs

Three pooled samples from each leached PAE category (water cooler, mineral water bottles, and baby feeders) were subjected to biodegradation. B. thuringiensis strain was grown in the MSM at standard conditions, i.e., 30 °C and 150 rpm. Growing cells were then collected after 48 h of incubation. For carrying out degradation experiment, 80 mL of MSM was prepared in a 250-mL Erlenmeyer flask sterilized and supplemented with 20 mL of leached PAE water samples and then inoculated with newly harvested culture and incubated at 30 °C for 72 h. Following incubation, media containing degraded leached PAEs (20 mL) were extracted with n-hexane (2 × 20 mL) and centrifuged (5000 rpm, 5 min). One milliliter of hexane extract was evaporated and reconstituted for subsequent degraded leached PAE analysis. Leached PAE degradation experiments were conducted in triplicates.

GC-FID conditions for leached PAE sample analysis

Leached PAE concentrations of all water samples were examined by gas chromatography (GC-8700, Perkin Elmer) with a flame ionization detector equipped with a DB-1 (0.25 mm ID, 0.25 μm film), 30-m long non-polar capillary column. Nitrogen was used as a carrier gas, and oven temperature was increased from 130 to 200 °C at a ramp rate of 5 °C min−1 and held for 8 min. Injection volume was 2 μL, and injector and detector temperatures were kept at 240 and 260 °C, respectively.

Valuation of analytical method

Each 10 phthalate esters at different concentrations, i.e., 1.0–25.0 μg L−1, were added into blank control water samples. The sample at each concentration was injected six times repetitively. The data obtained were analyzed statistically. The recovery rates range from 97.0 to 104%, and a good repeatability [calculated as relative standard deviation (RSD) from six repetitive determinations] ranges from 2.16 to 4.54%. Linear range, repeatability, limit of detection (LOD), and limit of quantitation (LOQ) were calculated and are presented in Table 1.

Table 1 Linear range of PAE calibration, repeatability, LOD, and LOQ of standards
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Leached PAE sample analysis by LC-MS/MS

Liquid chromatography was performed on a Merck C18 55 mm × 4.6 mm, 3 μm; column temperature was maintained at 45 °C. Mobile phase composition was methanol and 1% formic acid in water (50:50, v/v). A diode array detector was used at a 254-nm wavelength.

Liquid chromatography mass spectrometry (LC-MS/MS) was performed by using Agilent (Model 6460) triple quadrupole. Mass spectrometry was operated with ESI ionization source, positive mode, mass scanner range of 60–550 amu, fragment at 100 V; drying gas flow was 10 L min−1; nebulizer pressure was 50 psi; drying gas temperature was 350 °C; and capillary voltage was 4000 V. Sheath gas flow and temperature were 8 L min−1 and 250 °C, respectively. The rest of the analytical methods were identical to HPLC. The NIST Mass Spectral Search Program (National Institutes of Standard and Technology, Gaithersburg, MD) was used to identify signals by comparison to the retention time and mass spectra of authentic compounds.

Results and discussion

Identification of leached PAEs by LC-MS/MS

Leached PAEs from water samples were identified by LC-MS/MS technique. A total of 10 PAEs were identified based on their molecular ion peak and mass fragments. DMP was one of the most abundant and primary phthalates showing a molecular ion peak at 195 m/z and mass fragments of 195, 163, and 133 m/z.

Mass fragment 149 m/z is the fragment of protonated phthalic anhydride and is found in many PAE compounds. The phthalate esters with a base peak of m/z 149 included DEP, DPrP, diisobutyl phthalate (DiBP), DBP, bis-(4-methyl-2-pentyl) phthalate (DMPP), dipentyl phthalate (DPP), BBP, dicyclohexyl phthalate (DCHP), DEHP, and di-n-octyl phthalate (DnOP) (Yin et al. 2014). DEHP was also detected (molecular ion peak at 391 m/z) in one of the mineral water samples in the present study at low concentrations. DEHP is characterized as a developmental and reproductive toxicant by several studies (Ye et al. 2014).

Another most abundant PAE, DBP was detected with mass fragments 279, 149, and 205 m/z. Similar mass fragmentation patterns are reported by others (Amiridou and Voutsa 2011; Xu et al. 2005) as identified in the present study. DBP was listed as a priority pollutant by the US Environmental Protection Agency (EPA). The leached PAEs found in the present comprehensive study are summarized in Table 2.

Table 2 Mass fragments of leached PAEs obtained by LC-MS/MS analysis of water stuffs
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Leaching of phthalate esters from water cooler and mineral water bottle samples

Results of PAE leaching from water coolers and mineral water bottles exposed to sunlight are depicted in Table 3. DMP is one of the most abundant leached PAEs ranging from 67 to 112 μg L−1 in water cooler samples, while DEHP was detected in one local mineral water bottle (W7) sample.

Table 3 PAEs leached (μg L−1) from water cooler and mineral water bottles exposed to sunlight
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Variable results for PAE leaching have been reported for different matrices as depicted in Table 4. Bošnir et al. (2007) have reported similar results for DMP residues in soft drinks preserved with different additives as our present findings. In addition, a higher concentration of DMP leaching was observed by Bauer and Herrmann (1997); authors have discussed that a number of PVC products have been used in construction industries. This type of waste material disposed of in landfills possesses a threat to the aquatic environment.

Table 4 Concentrations of PAEs leached from various matrices reported in literature
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Lower levels of DMP were described by Montuori et al. (2008) for water bottles in PET; however, authors concluded that these DMP levels are significantly higher (nearly 20 times) compared to glass bottles.

In the present investigation, DMEP was the second most leached PAE ranging from 2 to 40 μg L−1 with the highest concentration in water cooler samples. Bao et al. (2015) have reported higher DMEP levels in baby care products (baby shampoo); in detail, authors have shown detections in less than 5% of products of some o-phthalates in which one is DMEP (CPSC 2011). DMEP is one of the colorless and oily liquid PAEs, which is water soluble up to 8500 mg L−1. DMEP is predicted to absorb very slowly into human skin, with a steady-state absorption rate of 8 μg cm−2 h−1 (Eastman Kodak 1991). Moreover, DMEP is responsible for acute and subacute toxicity, reproductive and developmental toxicity, genotoxicity, irritation, and sensitization (CPSC 2011).

DEP was detected in all leached water samples with relatively higher concentration in water cooler samples as compared to mineral water samples (average value 26.95 vs. 22.44 μg L−1). DEEP, DAP, DPrP, and DEeP were also detected in some of the leached water samples. On average, leaching of phthalate was higher in water cooler samples, i.e., 186.81 μg L−1 as compared to mineral water bottle samples, i.e., 140.54 μg L−1. The possible reason for leaching is usage of low-quality plastic in our region and storage of these bottles at a higher-temperature environment. In addition, continuous aqueous solubility of low molecular weight phthalate causes higher leaching of these as compared to higher phthalates. Reports of Wagner and Oehlmann (2009) and Saeed et al. (2010) support the present finding and discuss that such PAE leaching is due to low quality of plastic stuff (PET bottles) as well as solubility of these dangerous PAEs in water (total 479 μg L−1).

Though some researchers considered it unnecessary to conduct leaching tests from PET bottles into water at increased temperature (Sánchez-Martínez et al. 2013), others observed that water bottled in PET can be exposed to extreme heat exceeding even at 65 °C in cars parked in sunlight, garages, and other places not equipped with air-conditioning systems (Greifenstein et al. 2013; Westerhoff et al. 2008). Mihucz and Záray (2016) have investigated the short-term effect of water bottled in PET exposed to heat. For that purpose, they selected 0.5-L PET bottles that are ideal not only because this packaging volume is the most popular among consumers but also because several parallel samples can be thermostated or exposed to light simultaneously under controlled laboratory conditions (Keresztes et al. 2013). Casajuana and Lacorte (2004) determined EDCs (bisphenol A (BPA), 4-nonylphenol (NP), DEHP, DEP, DBP, DMP, and BBP) in water samples from PET and PE bottles after exposure outdoors for 10 weeks at temperatures up to 30 °C.

Leaching of phthalate esters via heating treatment of baby feeders

The impact of heating/sterilizing treatment on baby feeders for PAE leaching is presented in Table 5. As evident from results, boiling feeders was a better treatment as compared to autoclave and oven treatment. DEP was the major PAE ranging from 50 to 110 μg L−1 followed by DEEP (19–27 μg L−1) and DMP (15–20 μg L−1). Moreover, DEeP was detected in low concentrations (0.69–6.06 μg L−1), while DMEP was detected during autoclave treatment of feeder bottles. DAP was not detected in oven treatment-leached water samples.

Table 5 PAEs leached (μg L−1) during different treatments of baby feeders
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Davis et al. (2008) have found that heating bottles or pouring hot liquids into bottles, the presence of acidic or basic foods and beverages, and repeated washing have all been shown to increase the rate of BPA leaching from bottles. Moreover, Casajuana and Lacorte (2004) have reported that the most abundant compounds identified in five infant milk samples analyzed, independent of the type of the sterilizing process and packing material, were DEP, DBP, BEHP, and NP with levels from 7.3 to 85.3 μg kg−1, whereas DMP, BBP, and BPA were found at levels ranging from 0.28 to 2.93 μg kg−1. It is likely that the container enhances leaching, depending on when heat is used, while the milk is in a bottle (in-bottle sterilization) or in a larger vat prior to bottling (UHT).

According to Keresztes et al. (2013), phthalate leaching depends strongly on the PET bottle material (virgin vs. polymer containing recycled PET), pH (carbonated vs. non-carbonated samples), packaging volume, and temperature. The phthalate esters, i.e., DiBP, DBP, BBP, and DEHP, in Hungarian mineral water varied as follows: 3.0 ng L−1–0.2 mg L−1, 6.6 ng L−1–0.8 mg L−1, 6.0 ng L−1–0.1 mg L−1, and 16.0 ng L−1–1.7 mg L−1, respectively, for non-carbonated water samples stored exclusively for 90 days at different temperatures (22–60 °C). Phthalate ester concentrations, in virgin PET flake mineral water bottles, were negligible; therefore, the selection of the appropriate material and storage conditions plays a decisive role.

Subsequent investigation of DEHP presence in beverages (as a clouding agent) instead of legal additive has raised attention in 2011. Tolerable daily PAES intake was established as 0.5 mg kg−1 per day for DEPH and 0.01 mg kg−1 per day for DBP. The Taiwan Department of Health has established a standard of 1 mg L−1 for six phthalates (DEHP, DIDP, DINP, BBP, DBP, DNOP), and China established standards of DBP 0.3 mg L−1, DINP 9 mg L−1, and DEHP 1.5 mg L−1. Currently, the International Acceptance Criteria for the daily maximum consumption of phthalates range from 0.6 to 30 mg for a 60-kg adult, depending on the phthalate compound (TFDA 2012; Yang et al. 2013).

Biodegradation of leached PAEs by B. thuringiensis strain

PAE water-leached samples from each category, i.e., water cooler, mineral water, and baby feeder, were subjected to biodegradation by B. thuringiensis strain previously utilized for DMP degradation (Surhio et al. 2014).

Briefly, biodegradation reaction was continued for 72 h at 30 °C. The results are shown in Table 6. B. thuringiensis was able to degrade mixture of leached PAEs in the range of 75–96%, while DMP was degraded mostly >95% and the sequence of degradation was found as DMP > DBP > DMEP > DEP > DEeP > BBP > DAP > DEHP > DEEP.

Table 6 Biodegradation of leached PAEs from water and baby feeder samples by Bacillus thuringiensis
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The present investigation shows that use of B. thuringiensis offers rapid degradation of leached phthalate mixture as compared to abiotic degradation reported by Lertsirisopon et al. (2009). The abiotic degradability of BBP, DBP, and DEHP in the aquatic phase over a wide pH range 5–9 at room temperature was investigated. The experimental findings showed that abiotic degradation of the PAEs occurs mainly by photolysis under acidic and alkaline conditions rather than neutral pH, as a whole trend. However, DEHP did not degrade significantly at any pH. Hence, B. thuringiensis utilized in the present study is capable of degrading PAEs at neutral pH than abiotic strategies.

Conclusion

Leaching study of PAEs from water cooler and mineral water indicates that DMP was the most leached PAE followed by DMEP, DEP, DBP, DAP, and BBP in all samples. Different treatments of baby feeders (boiling, autoclave, and oven) were carried out which suggested that boiling feeders is safer than autoclave and oven treatment with respect to PAE leaching. In addition, degradation of leached PAEs with B. thuringiensis strain showed 75–96% degradation at neutral pH; hence, biodegradation is an effective approach for PAE degradation in water. Detailed information on human exposure to PAEs from PET packaging is required from academia, as well as substantiating or discarding the risk that chronic exposure to PAEs poses for human health.