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. 2025 Aug 25;59(35):18892–18903. doi: 10.1021/acs.est.5c07859

Spatial Distribution, Seasonal Variation, and Ecological Risk Assessment of Benzotriazole UV Stabilizers in Waters from Canadian Tributaries and Coastal Systems

Amina Ben Chaaben †, Mathieu Babin †, Frank Wania ‡, Hayley Hung §, Magali Houde ∥, Liisa Jantunen ⊥, Frank Gobas #, Huixiang Xie †, Zhe Lu †,*
PMCID: PMC12424182  PMID: 40854170

Abstract

Benzotriazole UV stabilizers (BZT-UVs) are industrial additives of emerging environmental concern, with UV-328 recently listed under the Stockholm Convention on Persistent Organic Pollutants and several congeners listed as Substances of Very High Concern in Europe. However, their distribution and fate in coastal environments remain poorly understood. This study investigated the spatial and seasonal variations of dissolved and suspended particulate matter (SPM)-bound BZT-UVs in surface water from the St. Lawrence River, Estuary and Gulf (SLREG) and the coast of Vancouver and Victoria, spanning Canada’s east and west coasts. BZT-UV contamination was higher in the SLREG, with peak UV-328 levels in July, likely due to increased summer use. Most congeners were more abundant in the SPM from July to October in the St. Lawrence Estuary, while elevated UV-329 levels in April suggest a distinct source, possibly related to snowmelt. These seasonal variations may influence the exposure of local species to BZT-UVs. While the concentrations of the dissolved BZT-UVs in most samples are expected to pose minimal ecological risks, the concentrations of some BZT-UVs in a few samples from the upper estuary of the SLREG may pose moderate to high risk in summer, highlighting the need for further assessment.

Keywords: UV absorbents, surface water, suspended particulate matter, partitioning, St. Lawrence River and Estuary, Salish Sea, hazard quotient

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1. Introduction

Benzotriazole UV stabilizers (BZT-UVs) are additives used in various industrial and consumer products such as plastics, paints, textiles, personal care products, cosmetics, sunscreens, rubbers, and lubricants to inhibit the degradation and discoloration of materials when exposed to sunlight. , Recent studies have revealed the widespread occurrence and biomagnification of some BZT-UVs in aquatic environments. − Exposure to BZT-UVs is known to cause various adverse effects, such as hepatotoxicity, immunotoxicity, reproductive and developmental toxicity in zebrafish (Danio rerio), − neurotoxicity in rainbow trout (Oncorhynchus mykiss), as well as endocrine toxicity in humans. In addition, BZT-UVs are known to cause increased reactive oxygen species production and lipid peroxidation in green algae (Chlamydomonas reinhardtii). Due to these properties, BZT-UVs are of emerging environmental concern. The Stockholm Convention on Persistent Organic Pollutants (POPs) has included 2-(2H-benzotriazol-2-yl)-4,6-di-tert-pentylphenol (UV-328) in Annex A for phase-out by 2026 and global prohibition by 2044. Thus, it is essential to investigate the distribution and fate of BZT-UVs in aquatic environments, assess their potential ecological risks, and establish a baseline for evaluating the effectiveness of regulations on these chemicals. However, such knowledge remains limited for coastal systems, which are complex transitional zones between terrestrial and marine environments, including estuaries and semienclosed marine waters influenced by both freshwater inputs and oceanic processes. To our knowledge, while previous studies have investigated BZT-UVs in the water of rivers, estuaries, or coastal systems separately, no study to date has systematically examined these contaminants in both dissolved phase and suspended particulate matter (SPM) in the full freshwater-estuary-marine continuum. This leaves a critical gap in understanding the transport, partitioning, and potential ecological risks of BZT-UVs across connected aquatic compartments, particularly in regions where urban and industrial activities influence multiple locations along this continuum.

The St. Lawrence River, Estuary, and Gulf (SLREG) and the Salish Sea are part of Canada’s east and west coasts, respectively. The Great Lakes-St. Lawrence system represents one of the largest river systems in the world. It is >3000 km long from Lake Superior to Cabot Strait and the drainage basin covers an area of about 1.6 million km2. , Approximately 25% of the world’s freshwater drains through this system, transporting contaminants from the Great Lakes watershed to the Atlantic Ocean. With an average discharge of 12,309 m3/s, the St. Lawrence River contributes about 88% of the freshwater inflow to the St. Lawrence Estuary. The Saguenay River, another important tributary, contributes about 12% with an average discharge of 1750 m3/s. To date, only limited BZT-UVs data have been collected from the SLREG, exclusively from the surface water of the fluvial portion of the SLREG near the outfall of Montreal’s wastewater treatment plant (WWTP), fish (lake sturgeon (Acipenser fulvescens), northern pike (Esox lucius), and deepwater redfish (Sebastes mentella)), and beluga whale (Delphinapterus leucas) tissues. ,,

The Salish Sea is a marginal sea of the Pacific Ocean, situated between British Columbia (BC), Canada, and Washington State. It covers approximately 18,000 km2 along with its surrounding watersheds. Vancouver and Victoria are major cities that discharge treated wastewater to the Salish Sea. In addition, the coastal waters of the Salish Sea are influenced by other anthropogenic activities, including transportation, shipping, and industries. Although BZT-UVs have not been reported in the Salish Sea, one compound in this group, namely 2-(2H-benzotriazol-2-yl)-p-cresol (UV–P), has been detected in air over the Salish Sea and its adjacent areas, suggesting that BZT-UVs may be present within the Salish Sea.

Understanding the spatial distribution and seasonal dynamics of BZT-UVs in surface waters, in both the dissolved phase and SPM, of the SLREG and the Salish Sea is critical for comprehensive and accurate ecological risk assessments. Identifying hotspots and understanding the sources of these contaminants are essential steps in elucidating their transport mechanisms and fate in these sensitive coastal environments. Given the limited data on BZT-UV levels in the SPM and the potentially significant influence of seasonal variations on their presence and concentrations, detailed spatiotemporal studies are necessary.

To better elucidate the contamination and impacts of BZT-UVs in aquatic environments, the present study aimed to (i) investigate the distribution of dissolved and SPM-bound BZT-UVs in the surface water from the SLREG and the coast of Vancouver and Victoria, as well as the seasonal variation in their concentrations in the St. Lawrence Estuary; (ii) study the partitioning of BZT-UVs between the dissolved phase and the SPM; and (iii) assess ecological risks of dissolved BZT-UVs in the surface water.

2. Materials and Methods

2.1. Sampling

For the spatial distribution study, 54 and 7 surface water samples were collected from the SLREG and the coast of Vancouver and Victoria, respectively (Figure ). Specifically, surface water (<50 cm) samples (n = 15) relatively distant from the shores were collected from 15 sites of the St. Lawrence Estuary and Gulf between September 6 and 10, 2019 aboard the CCGS Amundsen. Between August 15 and 26, 2020, nearshore surface water samples were collected from 39 sites along the St. Lawrence River and Estuary coast by wading operation. The SLREG sampling covered the fluvial section (n = 6), fluvial estuary (n = 12), upper estuary (n = 11), lower estuary (n = 10), gulf (n = 12), and the Saguenay River (n = 3) (Figure ). In the Salish Sea, samples were collected between August 10 and September 1, 2021, from nearshore water close to the cities of Victoria (n = 5) and Vancouver (n = 2) (Figure ) for comparison with the east coast.

1.

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Surface water sampling sites in (A) the St. Lawrence River, Estuary and Gulf (SLREG) (Quebec) and (B) the coast of Victoria and Vancouver (British Columbia). The blue dots represent the sites. The maps were created using QGIS 3.38 and MapChart.

In addition, surface water samples (n = 6 for each month; n = 42 total for 7 months) were collected from 6 sites in the St. Lawrence Estuary near Rimouski, Trois Pistoles, and Rivièredu-Loup between April and October 2023 around the middle of each month aboard the vessel La Macoma (Cadorette Marine Pro 220) to elucidate the seasonal variations of BZT-UVs. The coordinates of 6 sampling sites are shown in Text S1. Samples were not collected in winter due to ice conditions.

For each sample, 2 L of water was collected into a precleaned amber glass bottle. Samples were transported on ice and stored at 4 °C before processing and analysis. Water samples were stored at 4 °C in amber glass bottles to minimize background contamination from plastics and to avoid freezing-induced breakage (e.g., under −20 °C) and potential alteration of phase partitioning between dissolved and particle-bound contaminants. The bottles were filled with no headspace to minimize volatilization and degradation. For the seasonal variation study and the samples received from British Columbia, we processed the samples immediately after collection. Due to limited laboratory access and travel restrictions during the COVID-19 pandemic, samples from the St. Lawrence for spatial analysis were processed up to one year after collection. Previous studies have shown that BZT-UVs are stable for at least 100 days at 20 °C in a water–particle system. The duration of the study, not chemical instability, limited the stability result in that experiment. However, we acknowledge that this extended storage period could have resulted in an underestimation of BZT-UV concentrations in these samples. Since all St. Lawrence samples were stored and analyzed together under the same conditions, any potential losses would be systematic and unlikely to affect the spatial distribution patterns.

2.2. Chemicals and Reagents, Sample Processing, Instrumental Analysis, and QA/QC

The full names, CAS registry numbers, acronyms, and selected physicochemical properties are listed in Table S1. The structures of the target BZT-UVs are shown in Figure S1. Sample processing followed published methods with some modifications. Briefly, water samples were filtered using glass fiber filters (Whatman, Cytiva, Marlborough), and the dissolved phase was extracted using a glass ENVI-18 solid-phase extraction (SPE) cartridge (Sigma-Aldrich, Oakville, Canada). SPM on filters was extracted using an ultrasound bath in a solvent mixture of dichloromethane and n-hexane (v/v, 3/1), followed by cleanup with a glass LC-Si SPE cartridge (Sigma-Aldrich, Oakville, Canada). The final extract was analyzed using gas chromatography-triple quadrupole mass spectrometry in the multiple reaction monitoring mode. Additional details and QA/QC results are provided in Text S1 and Table S2.

2.3. Data Analysis

Data were analyzed using GraphPad Prism 10.4.1 (La Jolla) and R 4.4.1 with RStudio v2024.04.2. Descriptive statistics were calculated for BZT-UVs with detection frequency ≥ 50% based on NDExpo estimation (https://expostats.ca) when sample sizes >5. Half the method detection limit (MDL) was used to replace nondetects when sample sizes ≤5 and detection frequency ≥50%. SPM concentrations are reported based on dry weight (dw). Before the comparisons, the data were logarithmically transformed to approximate normal distributions. Given the unequal variances observed in the dataset, a Welch ANOVA, followed by Dunnett’s T3 multiple comparisons, was employed to assess the levels of contaminants across multiple groups. The differences in contaminant concentrations between the two groups were evaluated using an unpaired t-test with Welch’s correction based on log-transformed data. Correlations between BZT-UV congeners were tested using the clikcorr package (v1.0) in R, which is a method for testing correlations including censored data. The statistically significant level was set as p < 0.05.

SPM-water distribution coefficients K d (L/kg) were derived from the equation K d = C SPM/C water, where C SPM and C water are the concentration in paired SPM-bound (ng/kg, dw) and dissolved BZT-UV (ng/L) for each sampling site, respectively. For target contaminants detected in both the dissolved phase and SPM, the TOC-normalized SPM-water distribution coefficient (K OC) for each sampling site was estimated by the equation K OC = K d × 100/f oc, where f oc (%) is the organic carbon content fraction for each sampling site by weight. We note that deriving water-SPM partitioning K OC values may involve greater uncertainty due to the highly variable composition of SPM, which differs from the soil and sediment matrices typically used to develop partitioning models. SPM-bound concentrations were expressed in μg/g dw for consistent comparison across sampling sites with different SPM loads. Expressing SPM-bound concentrations in ng/L would introduce variability driven by fluctuations in SPM concentration across sites and seasons, making it difficult to distinguish whether observed differences are due to changes in contaminant loading or simply changes in particle abundance.

The hazard quotient (HQ) was determined using the equation HQ = MEC/PNEC, where MEC is the measured environmental concentration of dissolved BZT-UV in surface water and PNEC is the predicted no-effect concentration (Table S3). PNEC was calculated using the equation PNEC = ChV/100, where ChV is the geometric mean of the no observed effect concentration (NOEC) and the lowest observed effect concentration (LOEC) for the chronic toxicity of BZT-UV to fish, Daphnia magna, and green algae, and 100 is the assessment factor. The ChV was estimated by the benzotriazole class in the Ecological Structure–Activity Relationships (ECOSAR, v2.2) model.

3. Results and Discussion

3.1. Occurrence and Spatial Distributions

3.1.1. Dissolved BZT-UVs

All target BZT-UVs were detected in the dissolved phase of SLREG surface water, with 2-(2H-benzotriazol-2-yl)-4-(1,1,3,3-tetramethyl butyl)­phenol (UV-329) (detection frequency; mean ± SD; median: 72%; 1.29 ± 0.95 ng/L; 1.05 ng/L), 2-(2H-benzotriazol-2-yl)-4-(tert-butyl)-6-(sec-butyl)­phenol (UV-350) (61%; 0.74 ± 1.30 ng/L; 0.22 ng/L), 2-tert-butyl-6-(5-chloro-2H-benzotriazol-2-yl)-4-methylphenol (UV-326) (57%; 0.54 ± 0.37 ng/L; 0.62 ng/L) and 2,4-di-tert-butyl-6-(5-chloro-2H-benzotriazol-2-yl) phenol (UV-327) (56%; 0.76 ± 0.86 ng/L; 0.60 ng/L) found in more than 50% of the 54 samples collected for spatial distribution study (Tables and S4). The concentrations of the four congeners differed significantly, following the order of UV-329 > UV-327 ≈ UV-326 > UV-350 (Figure S2). A positive correlation was found between UV-326 and UV-329 (r = 0.34, p = 0.02, n = 54), suggesting a potential common source for these two congeners in the dissolved phase of SLREG water. For the coast of Vancouver and Victoria, target BZT-UVs were detected less frequently, with only UV-P (43%; <MDL-5.23 ng/L; median is not available due to detection frequency <50% (NA)), UV-327 (14%; <MDL–0.60 ng/L; NA), and UV-328 (29%; <MDL–0.15 ng/L; NA) being observed (Tables and S4).

1. Summary of Detection Frequency (DF) and Concentrations of Dissolved and SPM-Bound BZT-UVs in the Surface Water from the St. Lawrence River, Estuary and Gulf (Quebec) and the Coast of Vancouver and Victoria (British Columbia) .
St. Lawrence River, Estuary and Gulf (Quebec, n = 54)
  dissolved phase (ng/L) SPM (μg/g,dw)
compound DF mean median max DF mean median max
UV-P 39% NA NA 37.3 2% NA NA 19.0
UV-9 17% NA NA 2.22 30% NA NA 2.72
UV-320 43% NA NA 4.83 4% NA NA 0.74
UV-350 61% 0.74 ± 1.30 0.22 6.35 2% NA NA 2.08
UV-326 57% 0.54 ± 0.37 0.45 1.91 4% NA NA 0.89
UV-327 56% 0.76 ± 0.86 0.60 5.86 11% NA NA 0.67
UV-328 19% NA NA 22.0 28% NA NA 1.91
UV-329 72% 1.29 ± 0.94 1.05 4.19 50% 9.70 ± 27.8 0.29 184
UV-090 43% NA NA 20.5 13% NA NA 5.00
∑BZT-UVs   9.92 ± 10.5 5.74 42.6   10.7 ± 28.1 1.39 186
Coast of Vancouver and Victoria (British Columbia, n = 7)
UV-P 43% NA NA 5.23 43% NA NA 2.86
UV-9 0% NA NA <MDL 43% NA NA 0.41
UV-320 0% NA NA <MDL 43% NA NA 0.14
UV-350 0% NA NA <MDL 43% NA NA 0.56
UV-326 0% NA NA <MDL 14% NA NA 0.11
UV-327 14% NA NA 0.60 29% NA NA 0.13
UV-328 29% NA NA 0.15 14% NA NA 0.18
UV-329 0% NA NA <MDL 29% NA NA 1.43
UV-090 0% NA NA <MDL 43% NA NA 2.99
∑BZT-UVs   3.43 ± 1.19 3.11 5.33   1.88 ± 3.09 0.10 8.40
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a

NA: not available because of detection frequency <50%.

The dissolved ∑9BZT-UVs in SLREG surface water (9.92 ± 10.5 ng/L; median 5.74 ng/L) were significantly (p = 0.02) higher than those from the coast of Vancouver and Victoria (3.43 ± 1.19 ng/L; median 3.11 ng/L) (Figure ). Excluding the freshwater samples from SLREG and considering only those from the St. Lawrence Estuary and Gulf (9.81 ± 11.4 ng/L; median 4.78 ng/L), ∑9BZT-UVs concentrations did not differ significantly (p = 0.13) between the east and west coasts. The average composition of BZT-UVs in SLREG samples showed greater diversity, with UV–P, UV-329, UV-350, 2-[3-(2H-benzotriazol-2-yl)-4-hydroxyphenyl]­ethyl methacrylate (UV-090), and UV-327 contributing more than other congeners (Figure ). In contrast, samples from the coast of Vancouver and Victoria were predominantly composed of UV–P (Figure ). However, the limited sample sizes from the west coast introduce uncertainties in these comparisons.

2.

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Dissolved (A) and SPM-bound (B)∑9 BZT-UV concentration and average composition in surface water from the St. Lawrence River, Estuary and Gulf (Quebec (QC)) and the coast of Vancouver and Victoria (British Columbia (BC)).

As a recently listed POP by the Stockholm Convention, dissolved UV-328 was less frequently detected in the dissolved phase of the surface water from both SLREG and the coast of Victoria and Vancouver (Table ). Because UV-326 and 2-(2H-benzotriazol-2-yl)-4-methyl-6-(2-propenyl)­phenol (UV-9) are not on the Canadian Domestic Substances List (DSL), their detection in SLREG surface water suggests that they may be coming from imported materials or products containing these additives. In addition to UV-328, the European Chemical Agency (ECHA) has also classified 2-benzotriazole-2-yl-4,6-di-tert-butylphenol (UV-320), UV-326, UV-327, UV-329, and UV-350 to the Candidate List of Substances of Very High Concern (SVHC) for Authorisation, and UV-320 has been banned in Japan due to its potential persistence, bioaccumulation, and toxicity in the environment. Thus, it is worthwhile to further evaluate the impacts of these BZT-UVs on the coastal environment of eastern Canada. Furthermore, to our knowledge, UV-9 and UV-090 have rarely been reported in surface waters and have not previously been included in targeted analyses of coastal environments. Their detection in this study highlights the need to expand future monitoring and risk assessment efforts to cover a wider range of BZT-UVs in aquatic environments.

The spatial distribution of BZT-UVs in the SLREG is shown in Figure S3. Some sites in the St. Lawrence River, on the south coast of the St. Lawrence Estuary, and in waters of the Gulf of St. Lawrence showed elevated concentrations of ∑9BZT-UVs (Figure S3). However, this pattern was primarily influenced by the high UV-P levels at some sites (Figure S3). As samples were collected from two separate sampling campaigns, this introduced uncertainties that complicate spatial comparisons. To reduce such uncertainties, we further analyzed the spatial pattern for each compound based on the six geographic zones of the SLREG: the fluvial section (St. Lawrence River), fluvial estuary, upper estuary, lower estuary, Saguenay River, and St. Lawrence Gulf. Four general spatial patterns were observed in SLREG. UV-328 was more frequently detected in the surface water of the St. Lawrence River (50%; 0.84  ±  1.64 ng/L; 0.18 ng/L) than in other zones, suggesting that the river serves as a major transport pathway for UV-328 to downstream SLREG (Figure A). For UV-P and UV-090, both the St. Lawrence River and the Saguenay River exhibited higher levels than other zones, indicating that these rivers are key transport pathways for delivering these contaminants throughout the SLREG (Figure B). Additionally, higher levels of UV-P in samples from Quebec City (28.1 ng/L), Rivièredu-Loup (37.3 ng/L), and Rimouski (25.6 ng/L) compared to other coastal sites (<7.73 ng/L) suggest that these three cities are hotspots for the release of UV-P in the St. Lawrence Estuary (Figure S3). Two sites (28.5 ng/L and 31.5 ng/L) distant from the shores in the St. Lawrence Gulf also had higher concentrations of UV-P than others, which may be due to shipping traffic in these areas (Figure S3). For example, one site (S46) was near the ferry route between the north and south coasts of the St. Lawrence Estuary (Figure S3). Ships may release UV-P or other BZT-UVs from coatings and building materials, as well as with treated wastewater potentially containing residues of sunscreens, cosmetics, or other personal care products that include UV-P or other BZT-UVs. In addition to UV-P, UV-329 also showed elevated levels at the two offshore sites compared to other locations (Figure S3). However, the correlation between the two compounds was not statistically significant (r = 0.28, p = 0.08, n = 54), suggesting that they may not share the same dominant sources in these areas and/or may undergo different environmental partitioning or transformation processes that influence their distribution patterns in water. UV-320 and UV-350, which are isomers, showed similar spatial distribution patterns across the SLREG (Figure C). The Saguenay River may be an important transport pathway for both congeners to downstream zones (Figure C). In addition, activities in Port-Cartier and Sept-Îles may contribute to UV-320 and UV-350 contamination in the St. Lawrence Gulf, as elevated levels were found in samples collected from the coasts of these two towns compared to other areas in the Gulf (Figures C and S3). UV-326 and UV-327 had relatively higher concentrations downstream of Montreal, but their levels remained relatively consistent across other sections of the SLREG (Figure S3). This pattern likely reflects substantial inputs of these two congeners from the Montreal area, followed by dilution and/or stabilization of concentrations through a balance between inputs and removal processes downstream. Although elevated concentrations were found for a few hotspots (Figure S3), UV-329 showed no apparent spatial variation among the six zones, suggesting that there are various sources around the system and/or its input and removal may have reached a steady state in the dissolved phase of the St. Lawrence surface water (Figure D). For the Vancouver and Victoria coast samples, downtown Victoria had the highest levels of BZT-UVs compared to the other sites (Figure S4). Although Vancouver is a larger city than Victoria, the higher BZT-UV levels at the downtown Victoria site are expected due to its closer proximity to urbanized areas than the two Vancouver sampling sites. In addition, more shipping activity and the McLoughlin Point WWTP, approximately 1.8 km west of downtown Victoria, may also contribute to the higher concentrations observed.

3.

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Spatial distribution (mean ± standard error) of dissolved BZT-UVs in the surface water from the six geographic zones of the St. Lawrence River, Estuary and Gulf. Compounds are grouped by similarity in their spatial distribution: (A) UV-328, (B) UV-P and UV-090, (C) UV-320 and UV-350, (D) UV-327, UV-326, and UV-329. NA: not available because of detection frequency <50%; ND: not detected.

UV-328 was previously detected in Hamilton Harbor, which is located in Lake Ontario, one of the five North American Great Lakes and upstream of the St. Lawrence River. The median concentration was 0.9 ng/L, higher than the 0.2 ng/L found in the St. Lawrence River in the present study, indicating a decreasing trend of UV-328 from Lake Ontario into the SLREG system. Data on BZT-UVs in large rivers, estuarine, and marine waters are limited and summarized in Table S5. The concentrations of targeted BZT-UVs in the SLREG were comparable to those reported in the Yellow Sea and the East China Sea, but about 1–2 orders of magnitude lower than those reported for the rivers in central India. , In addition, similar to the pattern observed in the present study, UV-329 was also the predominant BZT-UV in the North and West Rivers of the Pearl River Basin in China, which drain into the South China Sea through the Pearl River Estuary. This pattern was also consistent with the water collected during the premonsoon period in the rivers of central India. Furthermore, UV-329 has been identified as the most frequently detected BZT-UV in the marine environment near the island of Gran Canaria, Spain, where nearby WWTPs affected the levels and compositions of BZT-UVs in the receiving water.

3.1.2. SPM-Bound BZT-UVs

Similar to the dissolved phase in water, all target BZT-UVs were detected in the SPM from SLREG, with UV-329 (50%, 9.70 ± 27.8, 0.29 μg/g dw) being the most frequently found congener. UV-328 (28%, <MDL-1.91 μg/g dw, NA) and UV-9 (30%, <MDL-2.72 μg/g dw, NA) were detected more often in the SPM from SLREG than other congeners, contrasting with their occurrence in the dissolved phase (Tables and S6). In the Vancouver and Victoria coast water, the relatively high detection rate of UV-P (43%, <MDL-2.86 μg/g dw, NA) aligned with findings from the dissolved phase (Tables and S6). Conversely, UV-9, UV-320, UV-350, and UV-090 were more frequently detected in the SPM than in the dissolved phase (43% in SPM vs. 0% in the dissolved phase), indicating possible particulate sources of these congeners and/or a strong sorption of these congeners to SPM. Although not statistically significant, the SPM-bound ∑9BZT-UV mean concentrations in SLREG were approximately 5 times higher than those in the coast of Vancouver and Victoria (Table and Figure ). However, the composition of BZT-UVs in the SPM differed between the two regions. UV-329 was the predominant congener in the SLREG, while UV-P and UV-090 were the dominant congeners in the SPM from the coast of Vancouver and Victoria (Figure ).

Figure S5 shows the spatial distribution of SPM-bound ∑9BZT-UV from each site in SLREG. This pattern was primarily driven by the high levels of SPM-bound UV-329 (Figure S5). To better understand the spatial distribution of other congeners detected at relatively high frequencies (i.e., UV-9 and UV-328), we performed a detailed analysis of individual congeners (Figure S5). Additionally, the sites were grouped into six geographic zones to facilitate a more structured and comparative assessment. In the SLREG, UV-9 and UV-328 displayed similar spatial patterns, with elevated concentrations observed in samples from the St. Lawrence River (Figure S6). This suggests that the SPM of this river serves as a vector for transporting these two BZT-UVs downstream, consistent with the spatial distribution of dissolved UV-328. For UV-329, distinct hotspots emerged across various geographic zones. Notable hotspots included Champlain (S8) (186 μg/g dw) in the fluvial estuary, Saint-Jean-Port-Joli (S18) (55.0 μg/g dw) in the upper estuary, and Forestville (S32) (33.2 μg/g dw), Port Cartier (S35) (61.5 μg/g dw), and Sept-Îles (S36) (43.2 μg/g dw) in the lower estuary (Figure S5 and Table S6). Among these, the highest levels of SPM-bound UV-329 were detected in Champlain (Figure S5 and Table S6). This site was located approximately 4 km downstream of an industrial zone and near the Gentilly-2 Nuclear Generating Station, where previous studies have documented extremely high levels of microplastics in sediment, and the water of this site is known to receive municipal or industrial effluents. This site also showed high levels of SPM-bound UV-328 (Figure S5). Given that UV-328 and UV-329 are commonly detected in plastics, their elevated levels in SPM at this location may be linked to significant microplastic contamination. Further spatial analysis was conducted using median values to mitigate potential biases introduced by these highly contaminated sites. The results showed that the Saguenay River was also crucial in transporting UV-329 to the SLREG, in addition to the contributions from the hotspots mentioned above (Figure S6).

In the coastal water of Vancouver and Victoria, a notable hotspot (site SS7) of contamination was identified near Victoria Harbour (Figure S7), approximately 900 m east of the McLoughlin Point WWTP. This finding suggests that the WWTP, along with shipping traffic activities at the port, could be potential sources of SPM-bound BZT-UVs at this site.

Data on SPM-bound BZT-UVs from comparable aquatic environments, such as large rivers, estuaries, and marine waters, remain scarce. In the present study, sites with relatively higher contamination of SPM-bound BZT-UVs from the two Canadian coastal environments generally exhibited levels comparable to, or in some cases exceeding, those reported for the Yellow Sea and East China Sea, except for UV-329 in the SLREG (Tables S5 and S6). Conversely, sites with lower contamination levels were comparable to those observed in the Yangtze River estuary and its adjacent area in China and five rivers in Germany (Tables S5 and S6). , Mean concentrations of SPM-bound UV-328 in the St. Lawrence River (0.84 μg/g dw) exceeded those reported for the Yellow Sea and the East China Sea (0.29 μg/g dw), Yangtze River estuary and its adjacent area in China (0.01–0.13 μg/g dw), and five German rivers (0.01 μg/g dw) (Table S5). ,, For UV-329, mean values in the SLREG SPM were approximately 1 order of magnitude higher than those in the Yellow Sea and East China Sea, and 2 orders of magnitude higher than those in the German rivers (Tables and S5). However, the elevated concentrations at the hotspot sites discussed above significantly influenced these mean values. When the median was used instead of the mean, the SPM-bound UV-329 concentrations (0.29 μg/g dw) in the SLREG were more consistent with the mean values reported for the Yellow Sea and East China Sea (0.29 μg/g dw), while remaining about an order of magnitude higher than those in the Yangtze River estuary and the German rivers (Table S5). , These comparisons show that the concentrations of SPM-bound BZT-UVs in Canadian tributaries and coastal systems are within or above the range reported for other major riverine and coastal environments in Europe and East Asia. The elevated levels of UV-328 and UV-329 in the SPM from SLREG waters at some hotspot sites further suggest localized sources. UV-329 was the dominant BZT-UV in the SPM of the SLREG water, which is consistent with the prediction that UV-329 would become the dominant BZT-UV in SPM by the mid-2020s in German rivers. These similar results suggest that the dominance of UV-329 in SPM in surface water may not be limited to German rivers but may reflect a broader trend in BZT-UV usage and environmental occurrence, highlighting the importance of continued monitoring and regulatory attention for this compound.

3.2. Seasonality

July showed elevated levels of dissolved ∑9BZT-UVs, largely due to significantly higher concentrations of UV-328 (Table S7 and Figure S8). In addition, July samples showed increased detection of UV-090, UV-320, UV-327, and UV-329 compared to most other months (Table S7 and Figure S8). In contrast, UV-9, UV-350, and UV-326 did not show clear temporal patterns (Table S7 and Figure S8). With respect to SPM, ∑9BZT-UV concentrations were highest in April, primarily due to concentrations of UV-329 (Table S8 and Figure S9). Most of the other congeners, including UV-350, UV-326, UV-327, UV-328, and UV-P, had higher levels between July and October than between April and June (Table S8 and Figure S9). The SPM-bound UV-9 level remained constant throughout the months, except in May and June, when it was undetectable (Table S8 and Figure S9). SPM-bound UV-320 was present in samples from July and August but was not detected in other months (Table S8 and Figure S9).

These results indicate compound-specific seasonal variations, which are likely influenced by fluctuations in BZT-UV input and environmental conditions. A previous study demonstrated that Canadian WWTPs typically have higher mass fluxes of UV-328 in summer influent than winter influent, suggesting increased use and release of this compound during warmer months. This observation is consistent with the presence of BZT-UVs, including UV-328, in sunscreens and personal care products, which are more commonly used in the summer. While direct seasonal data on Canadian WWTP effluent are unavailable, a study in China indicated higher levels of BZT-UV in WWTP effluent in the summer compared to the spring and fall, supporting the hypothesis of seasonal release patterns. In addition, microplastics, another potential source of BZT-UVs, are released in greater quantities from WWTPs during the summer and have been reported at higher abundances on the west coast of Hong Kong, China, during this season. , However, comparable seasonal data are not yet available for Canadian coastal waters. Moreover, BZT-UVs are commonly added to textiles to prevent UV-induced yellowing and degradation. UV-P was detected in cloth made by polyester and cotton, with concentrations up to 11.5 ng/g, and UV-328 was detected in blue cloth labeled as 100% cotton (up to 106 ng/g). The presence of BZT-UVs in clothing textiles may reflect either intentional use of these additives or unintentional contamination from other treated materials used for the cloth production. While there is no direct evidence of seasonal variations in their release from cloth textiles, it is reasonable to hypothesize that warmer months may affect their leaching and release due to the increased clothing use and washing frequency in the summer, greater UV exposure, and the prevalence of lighter and UV-protective clothing, which are often made from polyester. These conditions could potentially increase the release of BZT-UVs from fabrics into the environment through laundry wastewater. However, more research is needed to explore this seasonal pattern.

Hydrological and meteorological factors may also influence the observed seasonal variation. Due to limited detections and missing environmental data, multiple regression analysis was not feasible in the present study. Therefore, we focused on interpreting individual relationships between environmental parameters and BZT-UV concentrations. Negative relationships were found between ln­[UV-329 in SPM] and ln­(temperature) (r = −0.47, p = 0.04, n = 19), and between ln­[UV-327 in SPM] and ln­(pH) (r = −0.47, p = 0.027, n = 22) (Table S9 and Figure S10), suggesting higher levels of these BZT-UVs in SPM at colder temperatures and lower pH conditions. Negative correlations were also found between water salinity and SPM-bound UV-327 (r = −0.75, p = 0.007, n = 11) or UV-9 (r = −0.84, p = 0.002, n = 10) (Figure S10). This trend may be attributed to lower salinity waters, which are typically influenced by greater freshwater inputs and proximity to contaminant sources, resulting in higher SPM-bound BZT-UVs. Similar patterns were reported for UV-P in the eastern shelf seas of China, where terrestrial inputs predominate in low-salinity zones. Rainfall can directly transport BZT-UVs to receiving waters or wash BZT-UVs from urban surfaces (e.g., roadways, plastics, textiles, and industrial coatings) into waterways. , However, depending on the compound and water conditions at different locations, rainfall may have a dilution effect that counteracts the accumulation of contaminants. Snowmelt is likely a significant contributor to the high levels of UV-329 observed in April. , For example, UV-329 was found as a dominant BZT-UV in surface runoff comprising a mixture of rainfall and snowmelt in the Wełna River catchment, Poland, with concentrations ranging from 439  ±  75 to 898  ±  11 ng/L. Although SPM was not measured in that study, such high aqueous concentrations suggest the likely presence of much higher levels in SPM, given the strong sorption affinity of BZT-UVs to particles. However, this is also site-dependent. For example, a separate study conducted in Białystok (Poland), approximately 400 km east of the Wełna River catchment, did not detect UV-329 in either rainfall or snowmelt samples. Another factor to consider is the variability of water volumes within the St. Lawrence Estuary. Freshwater inflow peaks in May and is lowest from July to October. While high freshwater inflows from surrounding rivers can increase the delivery of contaminants to the estuary, they can also dilute their concentrations, which vary by specific compounds and sources. Lower water volumes from July to October may also lead to higher contaminant concentrations. The lack of detectable SPM-bound BZT-UVs in May and June is likely due to the peak water volume and associated dilution effects. The observed differences in seasonal trends between dissolved and SPM-bound BZT-UVs likely reflect a combination of the factors discussed above.

3.3. Water-SPM Partitioning

Very strong sorption of target BZT-UVs to SPM was observed in the present study. The measured log K OC values for BZT-UVs in the SLREG ranged from 6.51 ± 0.75 to 7.57 ± 1.01, exceeding the predictions from the polyparameter linear free energy relationships (PP-LFERs) in the EAS-E Suite (v0.972) (2.47–5.35) and the log K OW-based estimates from the KOCWIN Program (v2.00) in the Estimation Program Interface (EPI; v4.11) (3.10–4.94). The results of PP-LFERs based predictions are likely unreliable for BZT-UVs, as all values fall outside the model’s applicability domain. The field-based log K OC values were more closely matching the Molecular Connectivity Index (MCI) noncorrected estimates (5.65–6.96) in the KOCWIN Program (Figure S11). Our seasonal study observed moderate to strong linear regression results between measured log K OC and predicted log K OC or log K OW for a few sites, although these were not statistically significant. These results indicate that while predictive models can estimate log K OC for organic compound partitioning in controlled conditions, they may not accurately capture the sorption dynamics of BZT-UVs to SPM under field conditions. This is due in part to the complex and variable composition of SPM, which differs significantly from the soil and sediment matrices used to train these models. For example, in the St. Lawrence Estuary, SPM consists of a mixture of terrigenous particulate organic matter (POM) and marine POM, with pronounced seasonal and spatial variability driven by phytoplankton blooms and river inputs. , In addition, the presence of microplastics in St. Lawrence waters may alter SPM composition by acting as carriers of plastic-associated contaminants, potentially increasing field-measured log K OC values beyond model predictions. Environmental factors, such as temperature, salinity, pH, and SPM particle size composition in the SLREG, are likely to affect BZT-UV sorption to SPM but are not fully accounted for in existing models, resulting in an underprediction of BZT-UV log K OC under field conditions. Although the present study found no direct correlation between temperature, salinity, or pH and log K OC, these parameters showed negative correlations with SPM-bound BZT-UV, as discussed above, suggesting their indirect role in sorption dynamics.

The measured log K OC in this study was comparable to previously reported values for the Yangtze River Estuary and its adjacent areas in China (mean ∼5.5 to 6.5, range ∼4 to 9) but higher than those observed in several creeks in Toronto, Canada (4.3–5.8), and the eastern shelf seas of China (mean ∼ 4, range 2.5–5). ,, These variations may be attributed to different environmental conditions. For example, the composition of SPM organic matter may vary significantly across locations, which may influence the sorption of BZT-UVs to SPM. In addition, pH differences may have a modest influence on sorption. The hydroxyl groups of the target BZT-UVs have acid dissociation constants (pK a) ranging from 8.2 to 9.5. The lower pH in the St. Lawrence estuary (pH 7.75 ± 0.12 of samples collected in 2023), compared to the slightly higher pH levels in the Toronto creeks (pH 8.1–8.2), could result in a greater proportion of BZT-UVs existing in their neutral form in the St. Lawrence, thereby enhancing their sorption to SPM. However, the difference in pH is relatively minor compared to the much larger difference in K OC, suggesting that other factors (e.g., SPM composition, particle surface area, or organic matter properties) likely play a more important role in driving the observed differences in sorption behavior.

3.4. Ecological Risk Assessment

The potential ecological risks of BZT-UVs were assessed using HQ values for fish, Daphnia magna, and green algae based on dissolved concentrations in all SLREG water samples (n = 96). SPM-bound BZT-UVs were not included in the risk assessment because predicted ecotoxicity thresholds (e.g., PNECs used in this study) are derived from water-only exposure conditions, and the bioavailability of particle-associated contaminants is lower and more variable, making risk estimation based on the SPM phase uncertain and less comparable.

Most BZT-UVs showed minimal ecological risk. HQ values for UV-P and UV-9 were consistently <0.01, indicating negligible risk to all taxa (Figure S12 and Table S10). Similarly, UV-326, UV-329, and UV-090 posed no risk to green algae (Figure S12 and Table S10). UV-326, UV-329, and UV-090 generally posed no or low ecological risk (HQ < 0.1) to fish and Daphnia magna, although UV-329 showed a potential moderate risk in 3.1% of samples (Figure S12 and Table S10). Similarly, UV-320, UV-350, and UV-327 were associated with no and low risk in most cases, with up to 4.2% of samples reaching moderate risk levels (HQ: 0.1–1) (Figure S12 and Table S10). Notably, UV-328 exhibited moderate risk to fish, Daphnia magna, and green algae in 3.1–8.3% of samples, and high risk (HQ > 1) to fish and Daphnia magna in 3.1–5.2% of samples (Figure S12 and Table S10). High-risk samples associated with UV-328 were found near Quebec City and Rivièredu-Loup in July. We also calculated HQ values for samples from the coast of Vancouver and Victoria, and all HQ values were less than 0.1, indicating no or low risk.

Seasonal analysis revealed consistently no risk or low risk (HQ < 0.1) for most BZT-UVs in the St. Lawrence Estuary, with no significant temporal trends (Figure S12). An exception was UV-328 in July, which showed moderate to high risks due to elevated summer concentrations (Figure S13).

Our findings of minimal ecological risks from BZT-UVs in most surface water samples along the Canadian coasts are consistent with reported assessments for Chaohu Lake and eastern shelf seas in China and rivers in India. ,, A few samples in the SLREG may pose moderate to high ecological risks, consistent with results observed in the Pearl River Basin in China.

Uncertainties arise from the reliance on ECOSAR-derived chronic toxicity data and a fixed assessment factor of 100, which may not fully capture real-world variability. In addition, the co-occurrence of other contaminants could influence toxicity through synergistic or antagonistic effects, whereas this study focused on the estimation of individual BZT-UV risk.

3.5. Implications

The widespread presence of BZT-UVs in both the dissolved and SPM phases in two major Canadian coastal systems raises concerns about their persistence and the extent of their release into the environment. The detection of BZT-UV congeners that are not listed on Canadian DSL (e.g., UV-326 and UV-9) suggests that imported consumer products may be an underrecognized source of these contaminants. This underscores the need to track chemicals used in imported consumer products and highlights a potential regulatory blind spot in the tracking of chemicals present in finished goods. The seasonal variations observed in this study highlight the need for long-term monitoring and improved temporal resolution in sampling campaigns. Short-term or single-season sampling may fail to capture key contamination events, such as spring snowmelt or summer product use peaks. This is particularly important for migratory or seasonal species, such as the St. Lawrence beluga whale, which prefer the upper estuary as a summer habitat. The presence of these species in the upper estuary during the summer may increase their exposure to higher levels of certain BZT-UVs. Additionally, differences between modeled and measured K OC values point to limitations in current sorption prediction tools for SPM-bound BZT-UVs and similar compounds. Further research on SPM sorption mechanisms and specific modeling tools could improve our understanding of the fate and risks of these compounds. Moreover, given the strong sorption of BZT-UVs with SPM, risk assessments that focus only on the dissolved phase may underestimate actual exposure. There is a need to develop or adapt methods that account for particle-bound BZT-UVs to better assess ecological risks.

Supplementary Material

es5c07859_si_001.pdf (2.2MB, pdf)

Acknowledgments

We acknowledge the financial support from Environment and Climate Change Canada Grant and Contribution Agreements (GCXE20S011, GCXE25S047, GCXE20S010, GCXE20S008) under the Whale Initiative 1.0. The GC-MS/MS used in this study was supported by the John R. Evans Leaders Fund of the Canada Foundation for Innovation (#41056). This project was also funded by the Natural Sciences and Engineering Research Council of Canada (NSERC)’s Plastics Science for a Cleaner Future Program (#LLRP-558410-20) and Discovery Grant (#RGPIN-2019-05761; #DGECR-2019-0026). We appreciate the scholarships and travel awards provided by Centre De Recherche En Écotoxicologie Du Québec (EcotoQ), Québec-Océan, UQAR-ISMER and SETAC St. Lawrence Chapter. We are grateful to Dr. Jean-Carlos Montero-Serrano’s team (UQAR-ISMER), Abigaëlle Dalpé Castilloux, Bruno Cayouette, Christian Boutot, Clara Chamot, Elisa Michon, Julie Aquentin (UQAR-ISMER), and Kelsey Lee (Simon Fraser University), for the help with sampling. We also thank Fedia Boussarsar (UQAR), Ingrid-Alejandra Granados-Galvan, Alice Guillot, Clara Chamot (UQAR-ISMER), and Dr. Faqiang Zhan (University of Toronto) for their technical assistance. We thank the crew and fellow scientists of the CCGS Amundsen for the collection of samples in the St. Lawrence River.

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.5c07859.

  • It contains additional information on experimental details, QAQC results, BZT-UV properties and structures, GC-MS/MS analysis parameters, PNEC used for risk assessment, coordinates of all sampling sites, concentrations of BZT-UV in all samples, spatial distribution of BZT-UV in the dissolved phase and SPM, BZT-UV levels reported in the literature, seasonal variation of BZT-UV in the St. Lawrence Estuary, relationships between BZT-UV levels in SPM and environmental parameters, and risk assessment results (PDF)

The authors declare no competing financial interest.

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