Occurrence of TNP in the water source in east Taihu Lake
As shown in Fig. 2, the Ti concentration in WS substantially shifted during the sampling period, achieving an average Ti concentration of 224 ± 59 μg/L during the sampling period. The Ti concentration in sediments was relatively stable (0.70 ± 0.06 mg-Ti/g-dry). Similarly, a few studies have reported the Ti concentration in natural water and sediment. Tong, T. Z. et al. reported the Ti concentration in the sediment of a North Shore Channel in Skokie (USA) as 1.98 ± 0.04 mg-Ti/g-dry (Tong et al. 2015b), and from < 5 to 15 μg/L Ti in the effluent of a wastewater reclamation facility(Kiser et al. 2009). Thus, the Ti concentration in the water of Taihu Lake is significantly higher than the previously reported values, whereas the Ti concentration in sediment is similar. Principle component analysis (PAC) revealed the close correlation of Ti and turbidity (Fig. S2), indicating Ti in water may present in the state of colloid or suspended solids. Accordingly, the high contents of Ti in Taihu Lake is much likely due to the hydraulic disturbance of the sediment as the average depth is merely 1.9 m (Yang et al. 2008).
Taihu Lake from December 10, 2014 to January 11, 2015.
To estimate the presence of TiO2 in the samples, the acid-dissolved fraction was determined, according to Fig. S3. It was found that the major fraction of Ti-containing matter in both WS and WSS could not be dissolved in HCl, implying that the main component of the Ti-containing matter was present as acid insoluble matter, such as TiO2, FeTiO3, Ti(OH)3 and Ti2O3. Interestingly, HNO3 could not decrease the insoluble Ti percentage as much as HCl could in the samples, implying the presence of the Ti-containing matter, which could be oxidized to a species that is insoluble in acid, such as TiO. After being decomposed by an ultrasonic treatment, the size fractionation of Ti-containing matter in water and sediment of the water source was performed (Fig. 3). The main Ti-containing matter in WS and WSS was present in the > 5 μm fraction (approximately 77.3% for WS and 86.2% for WSS). The Ti concentrations between 0–0.1 μm in size in WS and WSS were 0.49 ± 0.62 μg/L and 0.57 ± 0.17 μg/g-dry, respectively. This result confirmed the presence of nano-scaled Ti-containing matter in WS and WSS. SEM analysis confirmed the presence of TiOX and FeTiOX in WS. Most of the Ti-containing particles are present in the μm size, whereas some Ti-containing particles are aggregates of nano-scaled particles as shown in (Fig. S4).
As shown in Fig. 4, the spectra of WS and WSS show the three characteristic peaks in the pre-edge (labeled by the imaginary lines) and the characteristic shapes after edge (Fig. 4a). The three peaks are due to the Ti 4p-Ti 3d hybridization and Ti 4p-O 2p hybridization in TiO2, suggesting the presence of TiO2 in the samples (Lee et al. 2003). Therefore, XANES further confirmed the presence of TNP in the water and sediment of east Taihu Lake.
According to the general estimation of the Ti-containing matter of east Taihu Lake in the above experiments, as well as the previous fitting results of Ti-containing matter in the environment (Tong et al. 2015b), we chose Ilmenite (FeTiO3), Anatase, Rutile, Brookite, and Amorphous TiO2 as the standards to represent the components of Ti-containing matter in the samples. The XANES spectra of the Ti reference compounds, on which the linear combination fitting (LCF) was based, are shown in the supporting information (Fig. S5). LCF was performed from −30 to 80 eV relative to the absorption edge (4966.4 eV) to quantify Ti speciation, and the data and results of the curve fitting are presented in Fig. 4b and Table S1. The best LCF yielded 0.00054, 0.0007, 0.00047, 0.00056 and 0.0036 for the R-factor for WSS, WS and the samples collected after coagulation/sediment, sand filtration and the effluent of WDP, respectively. The goodness of fit reported by the R-factor further provides substantial evidence for the presence of TNP in the samples.
Combined with the Ti concentration on the nano-scale, the TNP concentration in the WS and WSS was calculated using Eq. 1. The TNP concentration in WS was 0.86 μg/L, with a crystal composition of 0.44 ± 0.1 μg/L amorphous, 0.14 ± 0.03 μg/L anatase and 0.28 ± 0.06 μg/L rutile. The TNP concentration in WSS was 0.93 μg/g-dry, with a crystal composition of 0.31 ± 0.09 μg/g-dry amorphous, 0.22 ± 0.07 μg/g-dry anatase and 0.39 ± 0.12 μg/g-dry rutile.
It was found that no brookite was detected in the water samples and sediment. The natural scarcity of Brookite and its poorer production due to the lack of proper usage compared with the other phases may be the primary reasons for this result (Allen et al. 2008). It can also be seen that the amorphous component accounted for a large fraction of the detected TNP in the water samples and sediment. Generally, amorphous TiO2 has not been widely used in product applications due to its poor photocatalytic activity. Thus, amorphous TiO2 may be a natural mineral. Similarly, approximately 24% of amorphous TiO2 was present in the sediments of a river (Tong et al. 2015b). In addition, rutile and anatase were the main components in TiO2. Rutile is widely present in nature as a TiO2-containing mineral (Yu et al. 2013). Moreover, rutile has been massively produced for several decades. Similarly, anatase has a high photocatalytic activity and is used widely in products. Therefore, it is not surprising that rutile and anatase are detected in the samples. Previous studies reported similar environmentally relevant concentrations of TNP based on the exposure modeling. (Boxall et al. 2007; Gottschalk et al. 2009; Mueller and Nowack 2008).
Removal of TNP at the drinking water treatment plant
As shown in Fig. 5, significant fraction of both Ti and nano-scaled Ti was removed via coagulation/sediment (99.1% for Ti and 58.8% for nano-scaled Ti). The coagulation/sediment accounted for 61.5% in the total removed nano-scaled Ti. The removed fraction via sand filtration was 20.9%, whereas the disinfection/clear water reservoir removed 17.5% of nano-scaled Ti-containing matter. Additionally, the removal performance deteriorated as the size of the Ti-containing matter decreased. Similarly, Ti larger than 0.7 μm was well removed by a wastewater treatment plant in the USA, while the < 0.7 μm fraction was poorly removed and was present in effluents in concentrations ranging from < 5 to 15 μg/L (Kiser et al. 2009). The effective removal of TNP during coagulation/sediment was also found in the jar test (Abbott Chalew et al. 2013), and the electrostatic interaction was considered to be the main removal mechanism.
As shown in Fig. 4, all the spectra of the effluent at various stages at the drinking water plant contained the three characteristic TiO2 peaks in the pre-edge of TiO2 (Fig. 4a), which indicated the presence of TNP at the various stages at the drinking water plant. For the calculation, the influent Ti species was obtained using the data of the water source. Fig. 6 shows that the TNP concentration in the influent was 0.8 μg/L, with a crystal composition of 0.41 ± 0.05 μg/L amorphous, 0.13 ± 0.02 μg/L anatase and 0.26 ± 0.03 μg/L rutile, whereas the TNP concentration in the effluent was 0.33 μg/L, with a crystal composition of 0.24 ± 0.13 μg/L anatase and 0.09 ± 0.05 μg/L rutile. Similarly, Niall O’Brien et al. predicted that the mean TNP concentrations for the local drinking water treatment schemes ranged from 44.1 to 1450 ng/L (O'Brien and Cummins 2010).
In addition, different removal performances for the various TiO2 crystals were observed during the drinking water treatments. During coagulation/sediment, the removal performance decreased in the order of amorphous, rutile and anatase. Previous studies have reported the different characteristics of some types of TiO2 crystals in the aqueous environment. For example, Xuyang Liu et al. showed that crystallinity and morphology are not influential factors in determining the stability of TNP suspensions; however, the differences in their chemical compositions, notably, the varying concentrations of impurities (i.e., silicon and phosphorus) in the pristine materials, determined the surface charge, therefore determining the sedimentation and aggregation of TNP in the aqueous phase (Liu et al. 2011). Thus, the different removal performances during coagulation may be caused by the varying surface charge of the TiO2 crystals.
Sand filtration did not provide efficient TNP removal. Similarly, the batch isotherms revealed poor adsorption between quartz sand and TNP, and the quartz sand provided nearly zero retention of a 50 mg TiO2 per one liter stream in a column experiment (Rottman et al. 2013).
In terms of exposure pathways to humans, TNP in drinking water is most likely to gain exposure via the gastrointestinal and skin routes (Hagens et al. 2007; Lomer et al. 2002). Previous studies producing lethal and sub-lethal effects in animal toxicity tests revealed the toxicity thresholds of 5 g/kg for rat (Wang et al. 2007), 10 g/m3 for Daphnia magna (Lovern and Klaper 2006), 1 g/m3 for Rainbow trout (Federici et al. 2007) and 40 g/m3 for Algae (Hund-Rinke and Simon 2006).
According to China's water specifications, the drinking water consumption amounts per person per day is 75-220 L (Ministry of Construction of the People's Republc of China 2002). The adsorption and accumulation rate of TNP in humans was assumed as 10%, according to a previous study (O'Brien and Cummins 2010). Therefore, the annual ingestion of TNP through drinking water per person is 26.5 mg when using the highest water consumption amount, which is several orders lower than the above-mentioned toxicity thresholds. Moreover, this intake concentration is 100 times smaller than the concentration to which humans are exposed orally through daily food (0.2–2 mg/kg body weight per day of nano-TiO2/E171 in U.S.) (Weir et al. 2012). Thus, exposure to drinking water is extremely unlikely to result in a nano-specific toxicological response (O'Brien and Cummins 2010).
However, the presence of TNP in the drinking water system may result in secondary pollutants. For example, when TNP is present in the biological treatment, e.g., biological activated carbon filter, it may affect the pollutant removal performance because of its eco-toxicity (Li et al. 2014). Additionally, TNP may exert an influence on the disinfection effectiveness due to its affinity for bacteria (Wei 2011). Furthermore, TNP could alter the potential biological uptake of heavy metal ions, such as arsenic and lead (Miao et al. 2015; Sun et al. 2009). The coexistence of different types of nanoparticles was also reported to alter the original toxicity of individual nanoparticles (Tong et al. 2015a). Thus, comprehensive assessment of potential TNP toxicity in the drinking water system still requires careful integration of complex physicochemical interactions between TNP and other components in water.