One of the likely consequences of using chemicals in large quantities is their ultimate entry into environmental compartments. BUVs enter the environment through a variety of pathways. Some of them include: i) through wastewater treatment plant (WWTP) effluents; ii) plastic debris and or via weathering of outdoor plastics and products coated with BUVs; iii) personal care products (PCPs) containing BZTs as an additive (Lai et al., 2014a; Braush and Rand, 2011). The first occurrence of phenolic BZTs compounds was reported in the 1980s by Jungclaus et al., (1978) and Avila and Hites, (1980). The key findings of the articles highlighted the release of BZTs and other chemicals from a chemical unit in Cranston, USA that was in operation until 1985. The chemicals produced at this plant were discharged into the Pawtuxet River, eventually reaching Narragansett Bay in the United States. During the study, heavy deposition and accumulation of BZTs were observed in both water bodies. The research findings gave a clear insight of long-term persistence and environmental transport of BUVs after their release from the chemical plant. Figure 3 represents a complete picture of environmental fate of BUVs upon their release in the environment, while Supplementary Table (S3) summarizes overall occurrence of BUVs in different environmental matrices, including biological samples.
3.1 Occurrence in Wastewater treatment plants (WWTPs)
Wastewater treatment plants (WWTPs) are considered to be the main source of BUVs contamination in the environment (Cantwell et al. 2010). BUVs can enter the water treatment units easily via daily “wash-off” from products containing BUVs (Liu et al., 2012). Majority of data on occurrence of BUVs in WWTPs are from Western world, while there exists several findings highlighting its occurrence in Asian countries (Supplementary Table 3).
BUVs usually have high detection rates in both influent and effluent of WWTPs with concentrations ranging from hundreds of nanogram to thousands nanogram per liter (Molins-delgado et al., 2015; voutsa et al., 2006). Because of low hydrophobicity of BUVs, sorption does not play key role for elimination which is in accordance with poor removal rates (<80%) of BUVs during treatment in most of WWTPs (Heeb et al.,2012, Herrero et al., 2014a, 2014b; Karthikraj and Kannan,, 2017). Removal of BUVs in WWTPs is mainly governed by various factors like wastewater loading time, hydraulic retention time, and treatment factors (with or without biological units). Nakata and Shinohara (2010) reported individual concentrations of three BUVs (UV-326, UV-327, and UV-328) from 5.6 to78 ng/L in WWTP of Japan. Study from Portugal and Spain, reported concentration of five BUVs in the range of 24 to 85 ng/L with individual concentration of UV-328 varying from ND to 65 ng/L (Carpinteiro et al. 2012). Liu et al. (2012) reported the presence of UV-326 and UV-329 with mean concentration of 25 ng/L and 310 ng/L respectively in WWTP of Australia.
The estimated levels of BUVs in WWTPs are generally low, but uniform and similar to other persistent contaminants such as Polybrominated Diphenyl Ethers (Anderson and MacRae 2006). UV-328 levels in effluent and sludge samples from WWTPs in Sweden were measured in tens of ug/kg range. UV-328 had a maximum concentration of 3.1µg/kg dry weight in effluent suspended particles, while effluent samples from WWTP from Norway reported concentration in the range of 7 to 57 ng/L (Schlabech et al.2018). Studies have suggested open air system of lagoons and long HRT (<27 days) which may create favorable conditions like photodegradation, transformation and bioaccumulation for BZT degradation in WWTPs (Liu et al., 2012; Herrero et al., 2013). Negative removal rates of BUVs during wastewater treatment have been reported, which could be attributed with the fact that these compounds are predominant in nature and persist for longer duration in the aquatic environment. Further, the effluent from WWTPs can play a key role in polluting the receiving aquatic ecosystem (Zhao et al. 2017).
3.2 Occurrence in Receiving environment
Incomplete and inefficient removal of BUVs from WWTPs leads to its presence in effluent discharged from WWTPs to the receiving water bodies. Water samples from Narragansett Bay were found to be contaminated with UV-328 at concentrations of 7-85 µg/L respectively during 1978 to 1980 (Jungclas et al.1978), whereas water samples from San Francisco Bay were reported to contain UV-328 at a concentration of 17 ng/L (Silva and Muir 2015). UV-328 with other sun-blocking agents was detected in freshwater and seawater samples in Okinawa Islands, Japan. UV-328 was predominantly detected in seawater samples (2.8-287 ng/L) (Tashiro et al.2013). Water samples from heavily polluted river streams in Japan (Saitama prefecture) reported about 4.8×103 ng/L of UV-328 (Ciba 1988), whereas in seawater, UV-328 was reported up to the concentration of 287 ng/L (Tashiro et al.2013). Huge amount of plastic debris was also reported from these areas confirming, plastics as a carrier of these chemicals. Water samples from the Kaveri River in South India were contaminated with several classes of BUVs with UV-329 (31.3 ng/L) being the most prevalent in Indian subcontinent (Vimalkumar et al.2018). Surface water samples from Oslo and the Alna River, Norway were contaminated with UV-328 at concentrations of 0.8-17 ng/L and 1.0-1.9 ng/L after being characterized for BUVs (Schlabech et al.2019). Recently, the St. Lawrence River in Canada was reported to be contaminated by UV-328 at concentrations ranging from 1.2 to 3.0 ng/L (Giraudo et al.2020). Water samples from Philip Bay, Australia were predominantly contaminated with UV-328 with a concentration of 216 ng/L (Allinson et al.2018). To date only Denmark and Australia are the two countries having explicit drinking water quality standards for BZTs (<20 ng/L for CBT and BTri in Denmark and <2400 ng/L for 5-TTri in Australia) (Dummer, 2014).
Reddy et al.(2000) reported a very high concentration of BUVs in Pawtuxet River sediments, with concentrations of 4300 mg/kg for UV-P and 5200 mg/kg for UV-327. In comparison, the concentration in Narragansett Bay sediments ranged from 10 to 25 mg/kg, which is lower than that of Pawtuxet River sediments (Avila and Hites 1980). Pruell and Quinn (1985) reported similar findings in sediment cores throughout Narragansett Bay depicting the spatial and temporal distribution of the two BUVs. Phenolic compounds generally occur in two forms, the free form and bound form, the latter is also known as bound phenolics. Reddy et al. (2000) reported the occurrence of both these forms, indicating the sediment binding potential of BUVs in the bay.
Furthermore, Hartman et al.(2005) found two BUVs, UV-P and UV-328, in sediment samples from Narragansett Bay, indicating their extensive dispersion. Conventional WWTPs were ineffective in the removal of BUVs, as shown by the occurrence in mg/L levels of UV-327 and UV-328 in the receiving water. Finally, Cantwell et al. (2010) found that all other BUV compounds in Narragansett Bay were discharged from the same chemical manufacturing facility. UV-328 was also found in every stratum of sediment core samples, taken from Lake Ontario between 1975 and 2013. These data indicate of the persistency of UV-BZTs in sediments, and freshwater bodies. Urban streams in Canada reported UV-328 with concentrations up to 240 ng/g of sediment which was ten times higher than the concentration in rural areas of the state (Parajulee et al.2018).
Four BUVs were found in sediment samples from the Ariake Sea in Japan, in range of 1 to 16 µg/kg dry weight (Nakata et al.2009). In silt sample taken from the Omuta River in an industrialized area of Japan, the individual concentration of UV-328 was 320µg/kg. Individual concentration of three BUVs (UV-326, UV-327, and UV-328) in Songhua River, China ranged between 0.31 and 7.12 µg/kg (Zhang et al. 2011), while it ranged from 0.22 to 22µg/kg in two rivers (Detroit and Saginaw) in US (Zhang et al. 2011).
Pruell et al. (1984) were the first to report the bioavailability of two BUVs (UV-327 and UV-328) in bivalve mollusks (Mercinaria mercinaria) from Narragansett Bay in the range of 1.0 to 8.5 ug/kg and 7.0 to 61.0 µg/kg wet weight, respectively. Tissue burden of BUVs has been accounted for in numerous marine species in a study done by Nakata et al. (2009). UV-327 and UV-328 concentrations in the porpoise blubber varied from 4.5 to 31µg/kg and 11 to 64µg/kg, respectively. Furthermore, Kim et al. (2011) calculated the concentrations of eight BUVs in marine species from Manila Bay, Philippines ranging from not detected (ND) to 211 µg/kg wet weight. Four BUVs were found in blue green mollusks of eleven Asia-Pacific countries, with individual concentrations ranging from ND to 1500 µg/kg wet weight. UV-326 had the highest concentration (ND-1500 µg/kg wet weight), while UV-328 was predominantly found in the concentration range of ND-830 µg/kg wet weight in most of the sites. UV-320 was only accounted in mollusks samples collected from Japan in the range of 39 to 86 µg/kg wet weight (Nakata et al.2012).
3.2.4 Air and Household Dust
BUVs have been found as particulate and vapor phases of urban and indoor air in several regions (Xue et al.2016; Maceira et al. 2020). Six different BUVs were screened in air samples in Sweden in a study by Brorström-Lundén et al.2012, of which four BUVs (UV-320, UV-327, UV-329, and UV-360) were determined in the background and urban air samples. The highest concentration of phenolic BZTs was found in air samples collected in Stockholm, Sweden. Xue et al.2016 reported an average concentration of BUVs ranging from ND- 9.31ng/m3 in the vapor phase and 2.35-395 µg/gm in the particulate phase in New York, USA. A study by Maceira et al. 2018 reported the average concentration of BUVs in urban air at two different locations in Spain: Constanti (2.8ng/m3) and Tarragona harbor (2.2ng/m3). Constanti harbor had an average range of 0.93-5.5 ng/m3, while Tarragona harbor had a range of 0.69-4.1 ng/m3. Another study (Maceira et al. 2020) found that the average maximum concentration of different BUVs compounds was 962pg/m3 and 2248pg/m3. Inhalation of BUVs in ambient and indoor air can be the primary route of direct exposure to human populations.
Household dust samples and vehicles cabin dust were reported to contain BUVs (UV-326, UV-327, and UV-328) in the range of 22-4880 ng/g. The study also provided evidence of the exposure of BUVs to humans via consumer products (Carpinteiro et al., 2010) similar to PBDEs in household dust generated from consumer products. Significantly high quantities of PBDEs were found in WWTPs and household items, indicating their widespread use (Otazo et al.2005). Due to dust inhalation and exposure levels in humans, risk assessment can also be studied in professional drivers, spending eight to twelve hours of duration inside a car.
It is quite certain that BUVs are omnipresent in nature and have been detected in all environmental matrices, as evidenced by the literature. While researchers have documented the release of BUVs, it is clear that BUVs are released post-production, primarily during and after the use of products doped with certain UV stabilizers. What remains uncertain is their transport mechanism in environment through different carriers. The presence of BUVs in various environmental contexts raises legitimate concerns about their potential exposure to the human and animal populations