Introduction

Phthalate esters (PAEs) are restricted to the ortho form of benzenedicarboxylic acid prepared by reaction of phthalic acid with a specific alcohol to form the desired ester. Most of the esters are colorless liquids, have low volatility, and are poorly soluble in water but soluble in organic solvents and oils (Autian 1973). PAEs are endocrine disruptors and have been used as plasticizing agents in cellulosics and elastomers (Graham 1973). In addition, PAEs are widely applied in many daily products including floor tiles, various types of furnishings for households and transportation vehicles, food packaging systems, industrial tubing and conduits, medical tubing, catheters, blood containers, certain types of dental materials, drug coatings, and numerous other products. As a result, PAEs are ubiquitous pollutants in the environment and have been widely detected in the air, water, soil, and sediments (Peijnenburg and Struijs 2006; Teil et al. 2006). For example, according to the Toxics Release Inventory (TRI) database, the total release of PAEs in the USA in 2012 was 1, 492, 674 kg, with releases of 1, 354, 968 kg to air, 237 kg to water (direct), 1, 457 kg released on-site to land, 2, 204 kg transferred off-site to land, and 136, 010 kg to off-site disposal or other releases (USEPA 2012).

Due to their possible teratogenic, mutagenic, and carcinogenic effects (Koch and Calafat 2009; Meeker et al. 2009; Oehlmann et al. 2009; Talsness et al. 2009; Shaxson 2009), PAEs are classified as priority pollutants by the United States Environmental Protection Agency (USEPA) and several other countries (CCME 1989; USEPA 1980). Although several regulatory measures have been initiated at local, regional, and global levels to control the production and use of some of these chemicals (Andrady and Neal 2009), PAE use is not regulated in Asia. Furthermore, the domestic demand for PAEs in China has increased at an annual rate of 7.70 % from 2010 to 2015 (Emanuel 2011), and the total demand of PAEs was approximately 1.36 × 106 tons in 2010 (Wang et al. 2008a, b). In addition, China is currently the world’s largest PAEs importer, aggravating PAE contamination.

Effective environmental management practices for pollutants are dependent on the clear understanding of the ecological risk of chemicals (Lei et al. 2008). Therefore, more and more studies have investigated the potential ecological risk from PAEs in the environment (Staples et al. 2000; Xia et al. 2011). Few studies, however, have evaluated the ecological risk of PAEs, even though PAEs have been detected in different types of media (water, soil/sediment, and air) in recent years. The toxicity of PAEs to aquatic organisms generally increases with increasing alkyl chain length up to the point where a critical body burden cannot be attained due to the low aqueous solubility of the ester (Adams et al. 1995). For phthalate esters with alkyl chain lengths of six or more carbons, the results of numerous acute and chronic aquatic toxicity studies using many species indicate no toxicity at the solubility limit (Staples et al. 1997a). Therefore, the aim of this study was to reveal the ecological risk for the most detected PAEs with lower molecular weights (DMP, DEP, DBP, BBP, DHP, and DEHP) in aquatic environments in China.

Materials and methods

Data collection strategies

To reflect the overall status of PAE research in aquatic environments of China, a systematic literature review was performed using an electronic search of Elsevier, Springer, Google Scholar, ISI Web of Knowledge, and PubMed. Literature published in Chinese was retrieved from the China Knowledge Resource Integrated Database and the Wanfang of E-Resources for China studies, with master’s theses and doctoral dissertations included. Given the large number of studies in the literature, our study focused on those that are most relevant to Chinese aquatic environments. Studies that failed to report details of occurrence data and/or geographical information were excluded. Data on the toxicology of PAEs to non-target organisms were retrieved from the USEPA ECOTOXicology database (USEPA 2012) and supplemented by journal articles screened by the criteria of accuracy, relevance, and reliability (Klimisch et al. 1997). The collected toxicity data should be obtained under the guideline of good laboratory practice.

Analytical methods for PAEs in environmental media

This section will briefly introduce several PAE analysis methods. Cao (2010) reviewed the analytical methods that have been reported for the determination of PAEs in food. The main strategies developed included sample preparation, extraction, cleanup, separation, and detection. Extraction and cleanup are the most challenging steps for phthalate analysis in foods and are often the critical steps that determine the detection limits of the overall methods. The extraction method can be subdivided into solvent-based or liquid–liquid extraction (Page and Lacroix 1995) and solid-phase microextraction (SPME) (Arthur and Pawliszyn 1990). For separation and detection, studies have mainly focused on techniques such as liquid chromatography–mass spectrometry (LC–MS) (Lin et al. 2003), gas chromatography–mass spectrometry (GC–MS) (Farahani et al. 2007), and other methods (Kozyrod and Ziaziaris 1989; Cai et al. 2003; Sorensen 2006; Hogberg et al. 2008).

The analysis of PAEs in environmental media is generally similar to the analysis of PAEs in food, with the exception of the sample treatment process. Cai et al. (2003) developed a new analytical method for the analysis of PAEs in surface waters using solid-phase extraction, quantitative desorption with acetonitrile, and determination by high performance liquid chromatography (HPLC). In another study (Cortazar et al. 2005), a method to determine PAEs in sediment using solid phase extraction, desorption with MeOH, and GC–MS was developed. For atmospheric samples, Wang et al. (2008a, b) used liquid–liquid extraction and concentration, desorption with cyclohexane, and determination by GC–MS.

The fate of PAEs in ecosystems

Water solubility

Water solubility is an extremely important property that influences the biodegradation and bioaccumulation potential of a chemical as well as aquatic toxicity. Water solubility is also a determining factor that controls the environmental distribution of chemicals (Staples et al. 1997b). PAEs have water solubilities ranging from approximately 2014–0.00 mg/L. Most of the higher molecular weight phthalate esters (alkyl chain length of C6 or greater) are actually mixtures of closely related isomers. The physical and chemical characteristics of 16 PAEs are shown in Table 1.

Table 1 Physical and chemical characteristics of 16 PAEs
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Soil/sediment sorption

The sorption of phthalate esters to soil, sediment, or suspended solids is partially governed by the relative hydrophobicity of the chemical. The sorption is not always linear with the chemical concentration in the soil, and it may vary considerably with the particular solid used (Carlberg and Martinsen 1982). A number of authors have published soil or sediment and water partition coefficients (Table 1). In addition, several authors have examined the dissolved versus suspended particulate-bound fraction of phthalate esters in surface water samples. For example, Ritsema et al. (1989) used centrifugation to separate SPM from surface water samples collected from Lake Yssel and the Rhine River (Netherlands). The geometric mean of SPM values ranged from 4.00 to 100 mg/L; 98.0 % of the DBP present was dissolved, while only 2.00 % was SPM-bound.

Air–water partitioning

The equilibrium distribution of a chemical between water and air serves as a guide to estimate the tendency of a substance to escape from water into air. The ratio of the vapor pressure to the molar water solubility is an estimate of Henry’s Law constant, which is a measure of the equilibrium distribution coefficient (Thomas 1982).

For lower molecular weight phthalate esters (DMP, DEP, DAP, DPP, DnBP, DiBP, and BBP), H values ranged from 1.2E−7 to 8.8E−7 atm-m3/mole. Compounds with H values in the range of 1.0E−7 atm-m3/mole are generally considered to have negligible volatility (Howard et al. 1985). For all higher molecular weight phthalate esters except BOP, the calculated H values ranged from approximately 1.7E−5 to 5.5E−4 atm-m3/mole. The higher H values are due to a greater decrease in water solubility relative to vapor pressure with increasing alkyl chain length. The H values for some of the higher molecular weight phthalate esters (DiNP, DiDP, and DTDP) are difficult to calculate due to the extremely low vapor pressures and water solubilities that are not accurately known.

Degradation

In both aquatic and terrestrial systems (e.g., sewage, soils, sediments, and surface water), microbial action is thought to be the principal mechanism of PAE degradation (Staples et al. 1997b). Microorganisms that degrade PAEs can be aerobic (Wang et al. 1995), anaerobic (Shelton et al. 1984), or facultative. Precise measurements of biodegradation rates are considered important to accurately forecast the fates of potential pollutants and assess risk.

It is generally accepted that only the truly dissolved phase of a non-polar organic chemical is bioavailable (Steen et al. 1980). Therefore, the partitioning of phthalate esters into colloidal and particulate organic carbon should be considered in the analysis of field samples. Most historical measurements, however, are based on total concentrations, which fail to differentiate free and complex forms. Consequently, the available field data may significantly overestimate bioavailability, especially for the more hydrophobic phthalates that are expected to exist principally in the environment as complex forms (Staples et al. 1997a).

Ecological risk assessment

Ecological risk assessment of PAEs was conducted according to the European Commission’s Technical Guidance Document (EC 2003) and previous studies (Staples et al. 2000). The risk quotient (RQ) approach based on the measured contaminant concentrations in surface waters was used to assess the potential ecological risk. In this study, the RQ was assessed on non-target organisms, as described in previous studies (Staples et al. 1997b, 2000; USEPA 1995). The RQs were calculated as the quotient of the measured environmental concentration (MEC) and the predicted no effect concentration (PNEC). PNEC was estimated as the quotient of toxicologically relevant concentration and a security factor (f). For this purpose, the LC50 or EC50 values for fish, Daphnia, and algae associated with DMP, DEP, DnBP, and BBP were used for the RQ calculations. The RQs of PAEs were calculated as follows:

$$ RQ \, = \,\frac{MEC}{PNEC} \, = \,\frac{MEC}{{\frac{{L(E)C_{50} }}{f}}} $$
(1)

For data interpretation, the maximum probable risk for ecological effects from contaminated water was followed as recommended by Wentsel et al. (1996):

RQ < 1.00 (i.e., the exposure point concentration is less than the risk screening benchmark) indicates no significant risk;

1.00 ≤ RQ < 10.0 (i.e., the exposure point concentration is between one and ten times the risk screening benchmark) indicates a small potential for adverse effects;

10.0 ≤ RQ < 100 (i.e., the exposure point concentration is between ten to one hundred times the risk screening benchmark) indicates a significant potential for adverse effects;

RQ ≥ 100 (i.e., the exposure point concentration is equal to or greater than one hundred times the risk screening benchmark) indicates that potential adverse effects should be expected.

Results and discussions

Production and consumption of PAEs in China

Table 2 lists the total supply and demand volumes for PAEs in mainland China during 2000–2010; the variation trend for apparent consumption volume significantly increased during 2000–2010. Table S1 lists several major manufacturers of PAEs in China; the major manufacturers were located in Shangdong, Guangdong, Zhejiang, and Jiangshu province.

Table 2 The supply–demand situation of PAEs in the mainland China (104 tons)
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PAEs in aquatic environments

Table 3 lists the levels of PAEs detected in aquatic environments in different parts of China compared with other parts of the world. Most studies conducted in China reported PAE concentrations in river and lake waters to be higher than 8.00 μg/L, with the exception of three areas: Guangzhou (Urban Lakes), Beijing (Urban Lakes), and Yangtze River (Jiangsu section). In China, the Environmental Quality Standards for Surface Water (PRC-NS 2002) and Standards for Drinking Water Quality (PRCNS 2006) regulate the concentrations of two (DEHP, 8.00 μg/L and DBP, 3.00 μg/L) and three (DEHP, 8.00 μg/L, DBP, 3.00 μg/L and DEP, 300 μg/L) PAEs, respectively. Therefore, the potential ecological risk of PAEs cannot be ignored (Mo et al. 2001; Lan et al. 2012). Compared with other countries, the PAE concentrations in the waters of China are higher than global PAE levels (higher than 8.00 μg/L).

Table 3 Concentrations of PAEs in aquatic environment in China, compared with other countries
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River water from Ogun River located in Southwestern Nigeria contained a rather high level of 395–4, 775 μg/L PAE, which was 597 times higher than the guideline value for PAEs in environmental waters. With the increasing consumption of PAEs in metropolitan areas, the concentrations of PAEs detected in urban water bodies were obviously higher than those in other areas of China. The data also showed that urban river waters may receive PAE discharges from industrial effluents and landfill leakages without effective treatment; for example, extremely high PAE concentrations were detected in the effluents (up to 182 μg/L) from wastewater treatment plants in Berlin, Germany. Furthermore, PAE pollution of water bodies was found not only in surface waters but also in underground waters; for instance, PAE concentrations in the range of 0.00–6.70 μg/L were detected in underground waters in Dongguan, Guangdong Province, China. However, the PAE concentrations (except DBP) in surface water (rivers, lakes, and reservoirs) were generally higher than those in groundwater (Liu et al. 2014).

PAEs in soil and sediment

Table 4 lists some of the reported data on the PAE levels in sediment from different parts of China in comparison to other countries. The concentrations (maximum of 450, 000 ng/g, dry mass) of PAEs in metropolitan sediment/soils were generally higher than those from other areas of China. In addition, PAE concentrations in metropolitan sediment/soils were associated with the rather high levels detected in the water of the same area, confirming the influence of discharge from the local chemical plants associated with the manufacturing and processing of PAEs-based products. Compared with other countries, the PAE concentrations in the sediments/soils of China were higher than the global levels (Table 4). However, sediments from the North Sea in the Netherlands contained the highest level of 92.7–727.5 μg/g PAEs.

Table 4 Concentrations of PAEs in sediment/soil in China, compared with other countries
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PAEs in air

To comprehensively understand the PAE pollution level in China, we analyzed many studies that reported PAE concentrations in air. However, the studies concerning PAE concentrations in air are limited. Studies conducted in China have mainly focused on the ambient air concentrations of PAEs or PAEs in PAE-based factories. Table 5 provides several useful data concerning PAE concentrations in the atmosphere in both China and other countries and regions. These data show that the PAE concentrations (0.00–330,000 ng/m3) ranged over six orders of magnitude throughout the world, with a declining trend from continents to remote sites. In China, the level of PAEs was higher in Zhejiang than in the Tibetan Plateau, which is not an industrial district or a densely populated area. This result is reasonable because Zhejiang province has many PAE manufacturers.

Table 5 Concentrations of PAEs in air in China, compared with other countries
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At present, Rudel and Perovich (2009) reported that indoor air PAE concentrations are higher than outdoor concentrations and that concentrations in urban areas are higher than in rural and remote areas. Therefore, household goods and office furnishings have been demonstrated to be potential emission sources of PAEs (Wormuth et al. 2006). At the same time, outdoor sources of phthalates such as the wearing of tires are known to be secondary to indoor sources (Rakkestad et al. 2007). Kong et al. (2013) noted that the emission from cosmetics and personal care products, plasticizers and sewage and industrial wastewater may be important sources of PAEs in atmospheric particulate matter and that PAEs were preferentially concentrated in finer particles. These conclusions are consistent with the data in Table 4. Although no such data concerning PAEs in indoor or outdoor air and dust is available in China, the potential health hazards associated with the continued rise of the indoor use of PAEs, which may result in higher PAE concentrations in indoor air and dust, require more attention.

The ecological risk of PAEs in Chinese aquatic environments

The risk that PAEs pose to aquatic environments is still unknown. Continuous inputs and intrinsic toxicity are the main parameters that influence their effects on ecosystems. A risk assessment following the recommendation of the Technical Guidance Document on risk assessment (European Commission 2003) has been performed considering the L(E)C50 of fish, Daphnia, and algae; this assessment requires at least three trophic levels from the assessed environment. For those compounds, the toxicity data were selected from the review “Aquatic Toxicity of Eighteen Phthalate Esters” (Staples et al. 1997b). RQ values were calculated using the NOEC data along with the lowest L(E)C50 and a factor of 1000 (European commission 2003). Table 6 presents the toxicologically relevant concentrations (LC50, EC50, and NOEC) used for the RQ calculations. Tables 7, 8 and 9 present the RQ results for each compound and location. Because the locations that presented no significant, low or significant potential for adverse effects, and expected potential adverse effects were different for these three aquatic populations, three risk tables were used to illustrate all of these results. Among the PAEs, DBP, DEHP, and BBP presented the main contribution to the ecological risk values. The RQs for DMP varied from 0.00 to 2.78 for Bluegill Lepomis macrochirus, from 0.00 to 25.1 for Daphnia magna, and from 0.00 to 0.660 for Selenastrum capricornutum. In contrast, expected potential adverse effects (RQ > 100) for DEP, DBP, BBP, and DEHP were observed in some locations including Jiangshu-Yangtze River for Lepomis macrochirus populations, Songhua River-Jilin for fish populations, and Beijing-Chaoyang Park Lake for Selenastrum capricornutum populations. Generally, algae are especially susceptible to PAEs, while the RQs for invertebrates (Daphnia magna) are at least two-fold less. With the exception of DMP, DEP, and DHP, most RQs were in the range of 10.0–100, indicating significant risk related to the current predicted concentrations in aquatic environments. However, DMP, DEP, and DHP were found to pose no or low risk towards fish, invertebrates, and algae by growth inhibition (Staples et al. 1997a, b).

Table 6 Acute toxicity (LC50 or EC50) used for the risk assessment for Fish, Daphnia, and algae
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Table 7 Fish RQ results for DMP, DEP, DBP, BBP, DHP, and DEHP and the Sum of RQs for each location in China (μg/l)
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Table 8 Daphnia magna RQ results for DMP, DEP, DBP, BBP, DHP, and DEHP and the Sum of RQs for each sampled site in River
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Table 9 Selenastrum capricornutum RQ results for DMP, DEP, DBP, BBP, DHP, and DEHP and the sum of RQs for each sampled site in river
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To estimate the joint effects of these PAEs in China, a sum of RQs for each detected compound was calculated for each location. Based on these sums, no significant risk (RQ < 1.00) was observed in Yangtze River Delta-Xuzhou for any of the three populations. The RQ sums varied from 0.160 (Yangtze River Delta-Xuzhou) to 1407 (Jiangshu-Yangtze River) for fish populations, from 0.0400 (Yangtze River Delta-Xuzhou) to 333 (Jiangshu-Yangtze River) for Daphnia magna population, and from 0.310 (Yangtze River Delta-Xuzhou) to 2634 (Jiangshu-Yangtze River) for Selenastrum capricornutum population. The sums indicate that effects (RQ > 100) are expected in Jiangshu-Yangtze River for all of the three populations.

With the exception of Shichahai, the lakes in Summer Places, and Guanting reservoir, which showed no or low significant adverse effects, most of the urban lakes were observed to have expected or significant potential for adverse effects. Most of the rivers presented significant or expected potential for adverse effects, except for Wuhan-Yangtze River High Water Period, which presented no significant risk. For the other aquatic environments, the influents of wastewater treatment plants in Beijing presented expected potential for adverse effects, while most others presented low to significant potential for adverse effects. Thus, the ecological risk of PAEs in Chinese aquatic environments should be considered, and more short-term and long-term toxicological data on the synergistic effects of PAE mixtures in water at relevant urbanization environmental conditions are needed for a more reliable risk assessment.

Some studies have indicated that PAEs can accumulate in biota (Gorsuch et al. 2008; Cheng et al. 2013; Wang and Zhang 2013). Cheng et al. (2013) measured PAEs in 20 fish species collected from Hong Kong market; the ΣPAEs concentration ranged from 1.66 to 3.14 μg/g wet weight (ww) in fresh water fish and from 1.57 to 7.10 μg/g ww in marine fish. DEHP and DBP were the predominant compounds in both freshwater and marine fish. Mo et al. (2009) measured six PAEs in 11 vegetable species from nine farms of the Pearl River Delta; the total concentrations of PAEs ranged from 0.0700 to 11.2 μg/g dw, with a mean value of 3.20 μg/g (dw). The highest levels of PAEs were found in Brassica parachinensis, and the bioconcentration factors of the PAEs ranged from <0.000100 to 0.610. These results indicate that these phthalate esters can accumulate during gastrointestinal digestion in biota.

Conclusions

As one of the world’s most high produced and consumed chemicals, PAEs have become an ecological risk through their widespread and continuous exposure via food and drinking water. In general, the sources of PAEs in China were mainly derived from the manufacturing and processing of PAEs-based materials. Due to the demand for PAEs and PAEs-based materials, PAE pollution is predicted to become more serious in the future. The ecological risk assessment was performed based on measurements of PAE concentrations in aquatic environments and toxicity data. However, long term and bioaccumulation studies for these pollutants in aquatic environments are needed to define the environmental stress produced by the high concentrations of PAEs. There is an urgent need to monitor the sources, fates, toxicity, and ecological risk of PAEs in different environmental media in China, especially in highly urbanized areas or areas with PAE-based industries.