Review of Environmental Occurrence and Toxicity of Benzotriazole Ultraviolet Stabilizers

Abstract

Benzotriazole ultraviolet stabilizers (BUVs), extensively used synthetic UV absorbers, are established global contaminants exhibiting persistence and bioaccumulation. Ubiquitously detected in diverse environmental matrices (water, wastewater, sediment, air, dust, and soil) and organisms, their distribution shows significant regional and seasonal variations driven by urbanization and wastewater discharge. Notably, UV-328, UV-329, UV-234, and UV-P occur at elevated concentrations, with air and dust being the most contaminated media (even reaching levels ng/g to μg/g dw). Toxicological studies demonstrate that BUVs induce multifaceted biotoxicity, including endocrine toxicity, reproductive and developmental toxicity, immunotoxicity, hepatotoxicity, and neurotoxicity. Endocrine toxicity is the most extensively characterized effect. Moreover, BUVs disrupt critical human physiological processes and impair plant photosynthesis. Biotoxicity is mediated through molecular mechanisms: AHR pathway, ER/ERR activation, oxidative stress, lipid metabolism, and apoptosis. Among the BUVs, toxic mechanisms of UV-P, UV-234, UV-329, and UV-326 have been studied comprehensively. Notably, UV-P shows the highest estrogenic activity. While current environmental risks are assessed as low to moderate, future ecological disturbances are plausible due to intensifying anthropogenic pressures or bioaccumulation. Despite growing evidence, knowledge gaps persist regarding environmental fate and toxicology mechanisms. This review synthesizes BUVs contamination profiles, ecotoxicological impacts, and risk assessments, informing strategies to mitigate BUVs persistence and toxicity.

1. Introduction

Benzotriazole ultraviolet stabilizers (BUVs) are a class of aromatic organic chemicals that can effectively absorb ultraviolet light within the spectral range of 280–400 nm. With the core 2-hydroxyphenyl benzotriazole structure, BUVs exhibit high bioaccumulation potential. With octanol/water partition coefficients (logK _ow) ranging from 3.00 to 12.46 (Table ), BUVs exhibit hydrophobicity, facilitating their partitioning from aqueous phases into solid and organic phases, thereby influencing biological systems. These chemicals are incorporated into personal care products such as sunscreens, lotions, liquid foundations, and lipsticks. , Industrially, BUVs are utilized in the production of plastics, rubber, fuels, lubricants, construction materials, and electronics. , The global UV absorber market size was valued at USD 1.56 billion in 2024, and is projected to grow significantly, to reach an estimated value of USD 2.41 billion by 2032, driven by growing awareness among consumers and manufacturers about the harmful effects of UV radiation.

1. Names, Structures, Formulas, and LogK _ow of the Studied BUVs.

^aLogK _ow values are predicted by the software ECOSAR 2.2.

^bLogK _ow values are derived from experimental measurements.

As a class of recently identified emerging pollutants, BUVs have been officially reported because of their great environmental hazards. In 2023, UV-328 was officially included in Annex A of the Stockholm Convention as persistent organic pollutants (POPs). Furthermore, the European Chemicals Agency (ECHA) classified UV-320, UV-327, and UV-350 as possessing a high potential for bioaccumulation and significant resistance to degradation. In addition, both UV-328 and UV-329 have been designated by the Organization for Economic Cooperation and Development (OECD) as high-volume production chemicals. However, even though BUVs have attracted growing attention, the work of systematically combing through the environmental occurrences and biotoxic effects of multiple BUVs is still scarce.

In natural environments, BUVs demonstrate environmental persistence due to their structural stability, exhibiting resistance to hydrolysis and oxidation. Nevertheless, photodegradation may partially attenuate their persistence under specific conditions, although this process is highly chemical-specific and influenced by environmental factors (e.g., pH, dissolved organic matter (DOM), oxygen, inorganic ions, and suspended particulate matter). − Studies have demonstrated that BUVs photodegradation occurs through two pathways: direct photolysis and indirect photolysis. , Direct photolysis exhibits extremely low efficiency with negligible degradation rates, while indirect photolysis serves as the dominant pathway, relying on DOM to mediate energy transfer for generating reactive intermediates that facilitate degradation. , Despite this potential attenuation, previous studies have reported the widespread detection of BUVs in diverse environmental matrices. In surface water and wastewater, their concentrations are typically at the ng/L level. In sediment, air, dust, and soil, BUVs are detected at the ng/g dry weight (dw) level. Moreover, they are detected in organisms and even in various human samples, including breast milk, adipose tissue, urine, and serum. ,,− These studies confirm that BUVs efficiently transfer from environmental matrices to organisms, leading to significant bioaccumulation.

In addition, BUVs are characterized by high bioconcentration factors (BCF) and long half-lives, which makes them prone to accumulate in organisms and cause toxic effects over an extended period of time. − Recently, numerous previous studies have verified the biotoxicity of BUVs, including neurotoxicity, hepatotoxicity, endocrine toxicity, reproductive and developmental toxicity, and immunotoxicity. − Despite the increasing number of studies on BUVs, several significant knowledge gaps remain. Prior investigations have predominantly focused on BUVs known to be toxic, such as UV-P, UV-320, UV-326, UV-327, UV-328, UV-329, and UV-350. , However, for BUVs whose toxicity is not clearly defined, further research on their environmental fate and biotoxicity is urgently required. Furthermore, current environmental monitoring has limited geographic coverage, and temporal concentration trends are rarely documented. Toxicity studies primarily rely on aquatic models, and several key toxic effects remain inadequately studied. These critical knowledge deficits hinder a comprehensive assessment of the ecological and human health risks posed by BUVs and impede the development of effective regulatory strategies. Therefore, conducting a systematic review of the environmental occurrence and ecotoxicological effects of BUVs, while identifying current research gaps, represents an urgent priority.

This review comprehensively synthesizes recent data on the environmental concentrations and ecotoxicity of BUVs. First, it compiles and compares reported environmental concentrations of BUVs to identify chemicals with heightened environmental prevalence, while evaluating the impacts of temporal variations and regional disparities on concentration patterns. Subsequently, the ecotoxicological effects of BUVs and their underlying molecular mechanisms are systematically discussed. The primary objective of this work is to identify unresolved issues and research gaps by consolidating existing systematic data on environmental concentrations and current ecotoxicity studies of BUVs, thus assessing their environmental risks to guide future toxicological investigations.

2. Methodology

To comprehensively explore the research landscape regarding BUVs, a systematic literature search was conducted in the Web of Science and PubMed databases up to June 2025. The search utilized specific topic terms, including “benzotriazole ultraviolet stabilizers” (or “BUVs/BUVSs”), “benzotriazole”, “UV filters”, “environmental concentration”, “toxicity”, and “risk”. Through a comparison of chemical types, environmental characteristics, and toxic effects, these studies were filtered based on the following criteria: (1) Exclusion of studies not centered on BUVs; (2) Exclusion of studies that did not report the environmental detection concentrations, the biological toxic effects, or the ecological risk of BUVs. Ultimately, 45 articles reporting environmental concentrations of BUVs and 26 studies on their toxic effects were selected to summarize detection status across environmental media, organismal toxicity, and associated mechanisms, with 30 risk assessment studies analyzed for current environmental risks.

3. Occurrence of BUVs in the Environment and Organisms

3.1. Environmental Media

BUVs are detected ubiquitously in various environmental matrices worldwide, such as surface water, wastewater, sediment, air, dust, and soil (Figure ). BUVs have been detected in East Asia, North America, and Europe, but occurrence data from Africa, South America, Oceania, and polar regions remain unreported. Their distribution exhibits regional and seasonal variations, driven by urbanization, wastewater management, hydrological conditions, and treatment inefficiencies. Despite generally low detected concentrations, elevated levels of BUVs are observed in industrialized and urban regions.

1.

Environmental distribution of BUVs on a global scale. The schematic diagram illustrates the global distribution of BUVs in surface water, sediment, air, and dust matrices, combined with comparative concentration profiles across representative nations. The concentration data of BUVs in various countries is the arithmetic mean of the BUVs’ concentrations reported in different literature studies of the country. The bar chart is arranged in descending order according to the detected concentration. ^aKorea presents the median values to represent the concentration of BUVs.

3.1.1. Surface Water

Most of the current studies reporting the concentrations of BUVs in surface water are limited to Asia. The concentration data for surface water (including river water, lake water, and seawater) from various countries around the world, as presented in Table S1, show that the overall concentration of BUVs in surface water is relatively low (at the ng/L level). BUVs’ concentrations varied across studies due to region and measurement timing. Comparative analysis showed lower total levels in seawater than river and lake waters, likely due to seawater’s larger volume and greater water exchange. In Taihu Lake and Qiantang River, China, UV-327 dominated with average concentrations of 16 ng/L and 14 ng/L, respectively, whereas significantly lower levels (UV-327: 0.09 ng/L) were observed in the Yellow Sea and East China Sea. In these marine systems, UV-P reached its highest concentration at 1.42 ng/L, yet remained substantially lower than its detected concentrations in Taihu Lake (15 ng/L), Qiantang River (5.9 ng/L), Yangtze River (2.01 ng/L), and Pearl River (7.65 ng/L) in China. However, an exception was observed for UV-928, which was detected exclusively in southern coastal bays of South Korea at the concentration reaching 1.75 ng/L.

By comparing the data, it can be found that the concentrations of BUVs exhibit significant seasonal characteristics. For instance, a study conducted at the Yangtze River estuary, China, revealed that among the nine BUVs (UV-P, UV-PS, UV-320, UV-326, UV-327, UV-328, UV-329, UV-350, and UV-234), their concentrations were highest in summer (mean: 15.04 ng/L), followed by autumn (mean: 10.16 ng/L), and the lowest concentrations were observed in winter-spring period (mean: 4.40 ng/L). Similarly, the arrival of the monsoon and subsequent dilution during the rainy season led to a significant decrease in the concentrations of BUVs (mean = 7.14 μg/L before the monsoon and mean = 4.50 μg/L after the monsoon) in three Indian rivers.

Urbanization and industrial activities significantly shape BUVs’ concentrations. Discharges from urban Guangzhou, China, significantly influence BUVs’ concentrations in the Pearl River Basin, evidenced by strong positive correlations between water quality parameters (e.g., total phosphorus) and UV-327 and UV-P levels. In highly urbanized cities such as Wuxi and Hangzhou, China, BUVs’ concentrations have notably risen due to surface runoff and wastewater discharge.

3.1.2. Wastewater

Studies indicate generally low BUVs’ concentrations in wastewater, occasionally falling below detection limits. − Table S2 presents a summary of the concentrations of BUVs in the influents and effluents of diverse wastewater treatment plants (WWTPs) in recent years. A comparison of these data reveals that the purification capacity of WWTPs significantly impacts the concentrations of BUVs. At the Lüneburg WWTP in Germany, UV-326 had the highest average concentration at 320 ng/L in the influent, followed by UV-329. In contrast, only UV-328 was detected in the effluent. Similarly, Lu et al. analyzed samples from nine WWTPs in Canada and found significant disparities in BUVs’ levels between influent (UV-328 peaked at 45.1 ng/L) and effluent (UV-328 peaked at 3.61 ng/L). In Norway, the total concentration of BUVs in the effluent of a WWTP ranged from 300 to 8,900 ng/L, while the concentration in the sludge was 2 orders of magnitude higher, reaching 5 to 51 μg/g dw.

The treatment processes employed in WWTPs significantly influence the concentration of BUVs in wastewater. Primary sedimentation contributes notably to BUVs removal; however, its purification efficiency when used alone is low, hence it is generally applied only as the initial treatment stage. The most prevalent core treatment processes currently are the activated sludge process (including conventional activated sludge and anaerobic-oxic configurations) and biofilm-based processes. Studies indicate that activated sludge processes can achieve BUVs removal rates of about 90%. Nevertheless, they face challenges such as reliance on sludge adsorption for pollutant transfer, weak biodegradation, and inhibition of microbial activity by high BUVs’ concentrations. Biofilm processes exhibit good adsorption capacity for hydrophilic pollutants and are utilized as a core removal method in some WWTPs. Furthermore, technologies such as microfiltration/reverse osmosis, advanced oxidation processes, and ultraviolet disinfection are frequently employed during the advanced treatment stage. These methods can further reduce residual BUVs in the effluent. ,,

3.1.3. Sediment

The concentrations of BUVs in sediments vary significantly, ranging from not detected (ND) levels to several ng/g dw, as presented in Table S3. In multiple river systems in China, UV-327 and UV-329 have been found to have the highest environmental concentrations. The concentration of UV-327 in the Bohai Sea, China, was 0.119 ng/g dw, in Qiandao Lake, China, it was 15.4 ng/g dw, and the concentration of UV-329 in the Pearl River basin (North River), China, was 10.8 ng/g dw. While in many rivers in Europe and North America, UV-328 has been proven to have the highest concentration in sediments. ,, Especially in Mimico in Toronto, Canada, the concentration of UV-328 has reached 240 ng/g dw.

Terrestrial runoff and ocean currents substantially influence sediment BUVs’ types, concentrations, and distribution. In the Yangtze River estuary and adjacent areas, China, the average concentration of BUVs in sediments was 1.06 ng/g dw, with UV-PS and UV-327 being the most prevalent chemicals. Remarkably, higher concentrations were observed near the shore, while lower concentrations were found offshore. Similarly, sediments in Bohai Sea, China, showed elevated nearshore levels, primarily due to intensive BUVs use in Bohai Bay and polluted inputs from the Haihe, Xuanhui, and Majia rivers.

Human activities and industrial operations play pivotal roles in shaping the concentrations and distribution of BUVs in sediments. In small rivers near industrial zones in Tianjin, China, UV-326, UV-327, and UV-329 have been detected at a frequency of 100%, except for UV-320 (94.9%), indicating their widespread use. All sediment samples from Qiandao Lake in Hangzhou, China, contained BUVs, with UV-327 dominating and showing significant recent concentration increases, likely driven by rising anthropogenic usage. Moreover, in the Sanitary and Shipping Canal of Chicago, USA, UV-234 (100%), UV-328 (80%), and UV-P (70%) were frequently detected, demonstrating that shipping traffic influences the release of BUVs into the environment.

3.1.4. Air and Dust

BUVs undergo gas-particle partitioning, with low-volatility congeners (such as UV-328 and UV-329) predominantly adsorbed onto fine particulate matter, while volatile species (UV-P) exist significantly in the vapor phase. There are significant differences among different studies in terms of the reported environmental concentrations (Table S4). Considering the prevalence of BUVs in industrial and cosmetic products, their concentrations are relatively higher in industrial air and dust. For instance, street dust from industrial areas showed significantly higher total BUVs levels than cultural, residential, and transportation areas, in Hefei, China. Wu et al. found that the sum of median concentrations of BUVs in Canadian e-waste dust and residential dust was 447.84 and 98.2 ng/g dw, respectively, with that in e-waste dust significantly higher than in residential dust. In Tianjin, China, the sum of median concentrations of the four BUVs (UV-320, UV-326, UV-327, and UV-329) in residential areas, new urban areas, and cultural areas was 5.73, 5.6, and 5.58 ng/g dw, respectively. All these values were lower than the 6.16 ng/g dw recorded in industrial areas. Shunthirasingham et al. extensively collected air samples from urban and rural areas in Canada. They found that except for UV-P, which was present in all samples due to its strong volatility, the other BUVs were only detected in urban areas, with concentrations ranging from 10 to 240 pg/m³.

In nonindustrial areas, BUVs’ concentrations show substantial regional variations and are influenced by human activities. In the Malate residential area, Philippines, the total concentration of the seven BUVs (UV-P, UV-9, UV-234, UV-320, UV-327, UV-326, and UV-328) in house dust spanned from ND to 1,020 ng/g dw. In contrast, in the Payatas municipal dumping site, Philippines, the range was from ND to 277 ng/g dw. Notably, UV-326 and UV-327 concentrations were significantly higher in Malate, indicating household products as primary indoor contamination sources. A novel comparison of dust from plastic sports fields and indoor areas in universities in Beijing, China, revealed that the average concentration of ΣBUVs in indoor badminton courts was 11,023 ng/g dw, far exceeding those in basketball courts (4777 ng/g dw), plastic tracks (3779 ng/g dw), synthetic turf (1920 ng/g dw), and tennis courts (689 ng/g dw). Friction from track materials, shoe soles, and personal care products were identified as potential sources.

3.1.5. Soil

Currently, data on the concentrations of BUVs in soil are still extremely limited (Table S5). In the whole city soils of Tianjin, China, the sum of median concentrations of four BUVs (UV-320, UV-326, UV-327, and UV-329) was only 3.45 ng/g dw, which was much lower than that in dust (5.75 ng/g dw). , Yao et al. analyzed nine BUVs (UV-P, UV-PS, UV-320, UV-326, UV-327, UV-328, UV-329, UV-234, and UV-360) in Chinese agricultural soils and found that the average concentrations of ∑9BUVs in nonfilm-mulching soils, biodegradable film-mulching soils, and PE film-mulching soils were 0.48, 68.2, and 30.0 ng/g dw, respectively. This finding was consistent with the previous study reporting that the concentration of light stabilizers in ground-mulched soils was much higher than in agricultural soils.

3.1.6. Summary of Environmental Occurrence

Conclusively, UV-328, UV-329, UV-234, and UV-P have high detected concentrations in the environment (Figure ). Specifically, UV-328 has the highest average concentration in surface water (Figure A); UV-329 has the highest average concentrations in wastewater and sediments (Figure B,C); and UV-234 has the highest average concentrations in air and dust (Figure D). This is consistent with the fact that BUVs such as UV-P, UV-328, UV-329, and UV-234 are predominant in industrial areas and personal care products. ,

2.

Average concentrations of BUVs in four environmental matrices. The data for generating the graphs are, respectively, derived from Tables S1–S4, which are sourced from (A) surface water, (B) wastewater, (C) sediment, and (D) air and dust matrices. The unit most frequently used in the reported literature was selected, and the discrete values in the original data were removed. Each bar represents the arithmetic mean concentration, and the error bars indicate the standard deviation.

Notably, the concentrations of BUVs in air and dust are significantly higher than those in other environmental media, even reach levels on the order of μg/g dw. This phenomenon can be attributed to multiple factors. First, BUVs are directly released into air through industrial activities and personal care applications (e.g., plastic abrasion, cosmetic spraying), with subsequent dust deposition. Second, BUVs undergo preferential adsorption onto airborne particulates due to their high hydrophobicity and low water solubility. Third, BUVs resist hydrolysis, photolysis, and oxidation due to their chemical stability, particularly in microbially deficient environments like air and dust. Moreover, airborne particles facilitate long-distance BUVs transport via adsorption, enhancing their persistence in air and dust environments.

Existing research indicates clear seasonal variations in environmental BUVs’ concentrations. In addition to the surface water and sediments discussed previously, ,, airborne UV-P peaks in summer and minimizes in winter, whereas UV-328 exhibits an opposite seasonal pattern. These divergent trends mechanistically reflect chemical-specific volatility and particle adsorption coefficients. Moreover, BUVs’ distribution correlates inversely with distance from anthropogenic sources, with consistently elevated concentrations near industrial and residential zones across all media. ,, Additionally, urban wastewater discharge constitutes a key pollution source. Enhancing wastewater purification capacity and efficiency can mitigate environmental BUVs’ release, reducing their environmental concentrations. ,

However, comparative studies reveal a predominant research focus on sediment and atmospheric environments for BUVs, with significant gaps in surface water, wastewater, and soil. Geographical distribution of studies also exhibits marked imbalances, particularly in aquatic environments where concentration data remain largely confined to Asia. This limited spatial coverage impedes systematic understanding of BUVs’ environmental behavior. Furthermore, existing environmental concentration data suffer from both spatial blind spots and discontinuous temporal measurements, hindering cross-regional comparisons and long-term accumulation or temporal trend analyses.

3.2. Organisms

Influenced by diverse exposure pathways and metabolic capacities, the accumulation of BUVs varies significantly among different species (Table S6). In North American ecosystems, UV-328 manifested the highest concentration in the blood plasma of common carp (Cyprinus carpio) (median: 776 pg/g wet weight (ww)), followed by gizzard shad (Dorosoma cepedianum) (median: 762 pg/g ww), and then brown bullhead (Ameiurus nebulosus) (median: 411 pg/g ww). Granados Galvan et al. detected multiple BUVs in the livers of at least two Arctic wildlife species, with seabirds exhibiting notably higher exposure and accumulation susceptibility than other taxa.

BUVs exhibit significant variation in bioaccumulation, which can be partially reflected by BCFs (Table S7). Studies demonstrate high BCFs and bioaccumulation potential for UV-327, UV-328, and UV-326 during chronic exposure. ,, However, BCFs exhibit species-specific differences: Leubner et al. reported consistent steady-state and kinetic BCFs indicating moderate bioaccumulation of UV-P in rainbow trout (Oncorhynchus mykiss), whereas Wu et al. observed low bioaccumulation in aquatic organisms in Miyun Reservoirs, Beijing. These discrepancies likely stem from species-specific feeding habits. BUVs primarily accumulate in fish livers and kidneys, likely due to high protein content facilitating binding. , Current research on BUVs metabolites in organisms remains limited, primarily assessing metabolic persistence through half-life estimations, with hepatic and intestinal metabolism predominating.

At the ecosystem level, BUVs can migrate and transform within the food chain through feeding. Research synthesis indicates that BUVs trophic magnification occurs more frequently in marine ecosystems, while trophic dilution predominates in freshwater ones. For example, UV-329 demonstrated trophic magnification in apex predators such as beluga whale (Delphinapterus leucas). UV-326 and UV-327 have been reported to exhibit trophic magnification characteristics in marine food chains. In contrast to marine ecosystems, UV-326 and UV-327 showed trophic dilution within the food chain of Chaohu Lake, China. Neither Wu et al. nor Peng et al. observed trophic magnification of BUVs in freshwater ecosystems. ,

Human activities directly influence BUVs’ occurrence in ecosystems. The concentrations of BUVs in coastal organisms were correlated with the intensity of sunscreen use, especially during the peak tourism seasons. The absence of UV-320 in the biota of Chaohu Lake aligned with its limited regional application. Lu et al. discovered that organisms in aquatic environments receiving WWTP effluent had higher detection frequency and concentration of UV-328 in their plasma than those from low-impact areas.

BUVs enter the human body via multiple pathways, primarily dietary intake and inhalation. BUVs in food packaging materials may migrate into foodstuffs, thereby entering humans via dietary consumption. Recent studies have detected the highest BUVs concentrations in high-fat foods, particularly in beef. Emissions from plastic track materials and athletic shoe soles substantially elevate airborne BUVs levels in both indoor and outdoor sports facilities, subsequently leading to human inhalation exposure. ,, In addition, research has found that UV-328 exhibited a strong capacity for skin penetration, indicating that skin contact is also an important pathway for BUVs to enter the human body. Eight BUVs (UV-P, UV-PS, UV-320, UV-326, UV-327, UV-328, UV-329, and UV-234) were detected in the urine of Chinese adults. Among them, UV-P was the most prevalent, with a mean value of 1.6 μg/g creatinine and a detection frequency of 75%. It was followed by UV-327 (mean: 0.42 μg/g creatinine, detection frequency: 88%), UV-PS (mean: 0.21 μg/g creatinine, detection frequency: 67%), UV-328 (mean: 0.18 μg/g creatinine, detection frequency: 83%), and UV-326 (mean: 0.15 μg/g creatinine, detection frequency: 60%). In Japan, the concentration of UV-P reached a peak of 21 ng/g lipid in breast milk; in Philippines, it was 71 ng/g lipid. In Philippines, the concentration of UV-326 reached 34 ng/g lipid, and in Vietnam, the concentration of UV-9 was the highest, at 360 ng/g lipid.

The bioaccumulation potential of BUVs in human is increasingly evident, though significant metabolic knowledge gaps remain. Studies show their metabolic efficiency depends strongly on chemical structure; nonhindered BUVs (UV-P) undergo rapid Phase II enzyme-catalyzed clearance, exhibiting low bioaccumulation risk. In contrast, sterically hindered BUVs (UV-327 and UV-328) primarily rely on Phase I enzyme-catalyzed reaction (catalyzed by cytochrome P450), resulting in slow metabolism and high bioaccumulation. − Research has demonstrated that BUVs have the capability to interact with transport proteins and may act as substrates or inhibitors of P450 enzymes, thereby affecting metabolism. Although the liver (metabolizing large lipophilic molecules) and kidneys (processing small hydrophilic chemicals) are primary metabolic organs for BUVs, their contributions are limited, exemplified by UV-328 renal excretion accounting for <1% of administered doses.

Furthermore, BUVs exhibit plant-specific bioaccumulation patterns with implications for food chain transfer. In lettuce (Lactuca sativa), the root concentration factors for UV-320 ranged from 47.9 to 464 mL/g, while in radish (Raphanus sativus), they ranged from 194 to 787 mL/g, demonstrating species-dependent accumulation. Additionally, emerging evidence shows tissue-specific BUVs accumulation in plants. In reed (Phragmites australis), aerial tissues accumulated BUVs preferentially over roots, with UV-329 concentration in leaves exceeding UV-320 and UV-326. Analogous to phthalates and polybrominated diphenyl ethers, BUVs probably enter the root system through its hydrophilicity, translocated via xylem, and distributed to aerial tissues. This uptake mechanism establishes plants as critical vectors for BUVs entry into terrestrial ecosystems, posing ecological risks and necessitating urgent research on their plant distribution and transportation.

4. Toxicity of BUVs in Organisms

BUVs induce toxic effects in a variety of species, encompassing endocrine toxicity, reproductive and developmental toxicity, immunotoxicity, hepatotoxicity, neurotoxicity, and phytotoxicity. These toxic effects are mediated through several mechanisms, such as the aryl hydrocarbon receptor (AHR) pathway, estrogen receptor (ER)/estrogen-related receptor (ERR) activation, oxidative stress, lipid metabolism, and apoptosis. Pathway interplay causes organ damage, endocrine disruption, and developmental abnormalities, reflecting complex BUVs-induced toxicity across biological systems (Figure ). Among studied BUVs, UV-P, UV-234, UV-329, and UV-326 have relatively comprehensive toxicological characterization, whereas others remain understudied (Figure B). Current biotoxicological research predominantly uses freshwater vertebrate models (excluding endocrine toxicity) to extrapolate effects on higher animals. Direct evidence from mammalian systems or human cells remains scarce, demanding immediate expansion of such investigations.

3.

Graphical representation of the main toxic effects of BUVs in relation to molecular mechanisms. (A) Schematic illustration of major toxic effects and molecular mechanisms of BUVs. The five types of biotoxicity (endocrine toxicity, developmental toxicity, immunotoxicity, hepatotoxicity, and neurotoxicity) are mainly caused by the effects of seven toxicology mechanisms (e.g., the AHR pathway, ER/ERR activation, and oxidative stress). (B) Sankey diagram illustrating the association between 11 BUVs and their corresponding toxic effects with molecular mechanisms.

4.1. Endocrine Toxicity

The endocrine toxicity of BUVs has been a focal point of research, particularly their interference with estrogen, androgen, and antiandrogen activities. Already in 2003, Kawamura et al. used yeast two-hybrid assays to demonstrate that food packaging-derived BUVs disrupt endocrine function via ERα interference. In recent years, research has clearly shown that BUVs significantly disrupt the endocrine balance, warranting in-depth study.

Multiple studies indicate that UV-P shows the strongest activation ability for human estrogenic activity among all BUVs. Luciferase reporter gene assays for estrogen-related receptors (ERRγ/ERRα) revealed significant changes in receptor activity upon BUVs exposure. For ERRγ, BUVs (1 μmol/L) significantly enhanced its transcriptional activity, with UV-P showing the strongest induction, followed by UV-329, UV-320, UV-234, UV-328, UV-326, and UV-327. For ERRα, BUVs (1 μmol/L) except for UV-234 and UV-328 effectively activated it, and the agonistic activity was ranked as UV-P > UV-329 > UV-320 > UV-326 > UV-327. Similarly, MVLN assay demonstrated that UV-P, UV-PS, and UV-9 showed dose-related estrogenic activity, with the potency ranking: UV-P > UV-PS > UV-9. In addition, Sakuragi et al. discovered that UV-P, UV-PS, UV-090, and UV-329 (10 and 30 μmol/L) had significant agonistic effects on ERα, with the agonistic potency ranking: UV-P > UV-090 > UV-PS > UV-329. Similarly, UV-P, UV-PS, and UV-090 also showed agonistic activity on ERβ, with the potency ranking: UV-P > UV-090 > UV-PS.

Simultaneously, studies have demonstrated that UV-P (1 × 10^–4 μmol/L) showed remarkable antiandrogenic activity and can effectively inhibit androgen receptor (AR) activity. Zhuang et al. evaluated the effects of seven BUVs (UV-P, UV-234, UV-326, UV-327, UV-328, UV-329, and UV-350) on AR after metabolism and found that UV-P and UV-328 had noticeable toxic effects on organism metabolism at 0.25 μmol/L. Moreover, the recombinant yeast assay’s results showed that UV-P exhibited significant antiandrogenic effects in zebrafish (Danio rerio). Collectively, these findings demonstrate that among all BUVs, UV-P has remarkable capabilities to activate ER and inhibit AR.

What’s worth discussing is whether it is ERR or ER that is the key biomolecule mediating the estrogen-disrupting effects of BUVs. UV-326, UV-327, UV-328, UV-329, and UV-320 have been found to have significant agonistic effects on ERRα/ERRγ, while showing weak or no binding effects on ERα/ERβ. ,, The specific mechanisms still need to be further investigated and elucidated.

4.2. Reproductive and Developmental Toxicity

Abundant evidence demonstrates that BUVs induce developmental toxicity, including embryotoxicity, cardiac dysfunction, and teratogenicity. , The effects of BUVs on embryonic mortality vary among different chemicals and exposure concentrations. Johnson et al. reported a dose-dependent increase in zebrafish embryo mortality following microinjection of UV-P, UV-9, and UV-090, with potency ranking UV-P > UV-9 > UV-090. The 50% lethality (LD₅₀) values were 4,772, 11, 608, and 56, 292 ng/g-egg for UV-P, UV-9, and UV-090, respectively. In contrast, Liang et al. observed no significant mortality or hatching rate alterations in zebrafish embryos exposed to 1, 10, and 100 μg/L UV-234, UV-326, UV-329, and UV-P. ,

BUVs perturb embryonic cardiac function with species- and structure-dependent outcomes. Liang et al. found that four BUVs (UV-234, UV-326, UV-329, and UV-P) significantly decreased the heart rate of zebrafish embryos. Specifically, at concentrations of 1, 10, and 100 μg/L, UV-326 and UV-329 reduced the heart rate by 13.6–21.4%; UV-234 suppressed the heart rate by 13.1% and 16.3% at 10 and 100 μg/L, respectively; and UV-P significantly decreased the heart rate by 11% at 100 μg/L. UV-328 led to a concentration-dependent decrease in the heart rate of zebrafish embryos and larvae. Conversely, Eriksson et al. observed elevated heart rates in UV-327-exposed rainbow trout juveniles compared to controls. These findings necessitate multichemical cross-species comparisons for BUVs cardiotoxicity assessment.

BUVs induce diverse developmental abnormalities across aquatic organisms. Zebrafish embryos exposed to UV-P, UV-9, and UV-090 exhibited yolk sac edema, spinal curvature, and dose-independent deformity rates. Prolonged exposure to BUVs caused axial edema, otolith defects, and pericardial edema in zebrafish. UV-327 reduced the area of the yolk sac utilized for development and the growth length of rainbow trout during development. In Asian clam (Corbicula fluminea), UV-234, UV-326, UV-329, and UV-P provoked digestive gland lesions (including irregular digestive cavities, luminal vacuolation, and disrupted basal membranes) and gill pathology. , Multispecies evidence suggests BUVs teratogenicity may elicit tissue-specific pathological manifestations.

Furthermore, BUVs interfere with thyroid hormone pathways through receptor-level modulation. Zebrafish embryos exposed to UV-234, UV-326, UV-329, and UV-P showed altered thyroid receptor (thraa/thrb) mRNA expression. UV-326 (1 and 10 μg/L) significantly upregulated thraa/thrb. UV-234 (10 and 100 μg/L), and UV-329 (100 μg/L) upregulated thrb, while UV-P (10 μg/L) downregulated thrb. Fent et al. confirmed UV-P-induced thra suppression at 690 μg/L. Transcriptional evidence indicates receptor-mediated thyroid disruption by BUVs may contribute to developmental delays.

Limited evidence suggests paternal UV-P exposure impairs fish reproductive success via spermatogenesis disruption. Microinjection studies demonstrated reduced fertilization in male zebrafish due to inhibited steroid synthesis and spermatogonial meiosis, with no effect on females. This sex-specific toxicity emphasizes the need to evaluate paternal contribution in chemical risk assessments.

4.3. Immunotoxicity

The immune system plays a crucial role in organisms. BUVs exhibit immunomodulatory effects via immune cell regulation. Kubota et al. demonstrated that UV-P and UV-PS dose-dependently stimulate regulatory T cell (Treg) differentiation in mice through both in vitro and in vivo experiments, with UV-PS showing stronger in vivo Treg induction in splenic tissues than UV-P.

BUVs differentially regulate the expression of immune-related genes based on corresponding targets. Under UV-328 exposure, the expression levels of il12 and tnfα in zebrafish larvae increased in a concentration-dependent manner. Li et al. discovered that exposure to 100 μg/L BUVs (UV-234, UV-326, UV-329, and UV-P) led to the suppression of certain immune genes, such as nos2, stat1, pparδ, il6r, and il7r, while upregulating genes like tlr8, ccl2, il13, il22, and tnfsf14. Subsequent research revealed compound-specific immunotoxic mechanisms: UV-P upregulated ahr2 and cyp1a1 to inhibit Interleukin (IL)-22; UV-234 targeted hepatic il17a and il22 causing pathology; UV-329 dose-dependently activated ahr2 signaling, altering il6 and il17a. This suppression or activation pattern suggests BUVs induce immune imbalance through targeted gene interference.

Transcriptomics reveal dose-dependent disruption of immune gene networks by BUVs. Li et al. found that four BUVs (UV-234, UV-326, UV-329, and UV-P) collectively regulated immune-related pathways, encompassing B-cell activation, leukocyte differentiation, and complement activation. UV-326 specifically suppressed brain IL-1β-high-mobility group box 1 (HMGB1) signaling in zebrafish, impairing innate immunity. Under four BUVs (UV-P, UV-326, UV-329, and UV-234) exposure, six conserved pathways associated with inflammatory response, immune system activation, cell proliferation, and myogenesis were activated. The activation of the Retinoic acid-inducible gene (RIG)-I-like receptor signaling pathway by UV-328 led to an enhancement of the actions of downstream inflammatory cytokines. Additionally, Liang et al. observed transcriptional alterations in Toll-like receptor (TLR)-associated genes (e.g., tlr5a and tlr5b) upon exposure to UV-234 and UV-320, except for the suppression of il8 and cxcl-c1c induced by 0.1 μmol/L UV-320.

4.4. Hepatotoxicity

As the central metabolic hub, the liver readily accumulates BUVs, causing potential damage evidenced by histological alterations. Normal hepatocytes show a polygonal shape with central nuclei and intact architecture. In contrast, UV-234 (100 μg/L) induced nuclear pyknosis and vacuolization in zebrafish liver, while UV-329 (100 μg/L) caused hepatic disorganization and nuclear condensation. Similarly, ruptured hepatocytes, hydropic degeneration, vacuolar degeneration, and leucocyte infiltration were observed in yellow catfish (Pelteobagrus fulvidraco) under UV-234 exposure. Furthermore, Hemalatha et al. observed dose/time-dependent hepatic damage in UV-328-exposed zebrafish, with 42-day exposure (0.01–1 mg/L) causing venous hemorrhage, nuclear pyknosis, sinusoidal necrosis, and hepatocyte degeneration. Notably, at 1 mg/L, blood sinusoid dilation and melanomacrophage aggregates indicated advanced injury.

Hepatocyte metabolic abnormalities indicate cellular dysfunction and may contribute to liver diseases. For instance, Li et al. observed that UV-234 exposure in yellow catfish elevated hepatic superoxide, reactive oxygen species (ROS) accumulation, and malondialdehyde (MDA) levels, while depleting antioxidants including total antioxidant capacity (T-AOC), superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GSH-PX) activities, and glutathione (GSH) content. Similarly, UV-328 in zebrafish dose-dependently changed antioxidant enzyme (including SOD, CAT, and GSH-PX) activities and MDA levels in livers. ,

4.5. Neurotoxicity

Previous studies demonstrated that benzotriazole (BT), a precursor of BUVs, altered expression of neural function-related proteins, including creatine kinase, brain b (CKBB) and glutamine synthetase (GS). , This suggests that BUVs may have similar toxic effects on the nervous system.

A previous study has demonstrated that BUVs impact the nervous system, especially in eliciting visual motion responses (VMR) in darkness and affecting early stage swimming behavior. Specifically, zebrafish larvae exposed to UV-234 (1 μmol/L) displayed reduced locomotor activity during the dark phases, while UV-320 (1 μmol/L) increased this activity during the light phases. When exposed to 1 μg/L UV-234, locomotor acceleration of zebrafish during light/dark stimulations was significantly elevated. Meanwhile, both UV-326 and UV-234 (1, 10, 100 μg/L) increased dark-phase swimming distance but inhibited spontaneous tail coiling. Rainbow trout alevins exposed to UV-327 (107 ng/g egg) exhibited hyperactivity and altered photomotor responses posthatching. Neurobehavioral impairments in nematodes exposed to UV-329 (10–100 μg/L) were also manifested as concentration-dependent decreases in head thrash and body bend frequencies, along with reduced motion wavelength and mean amplitude. These dose/structure-dependent behavioral disruptions highlight unresolved mechanistic complexities in BUVs neurotoxicity.

At the molecular level, the neurotoxicity caused by BUVs involves highly complex pathways. Transcriptomic analysis revealed that UV-327 disrupted neurological pathways in rainbow trout, including circadian rhythms, ion homeostasis, acetylcholine signaling, and energy metabolism. Concurrently, UV-234 and UV-326 differentially modulated neuroinflammatory markers in zebrafish: UV-326 at a concentration of 100 μg/L upregulated il1β and il6, while UV-234 suppressed il1β at the same concentration. Additionally, both UV-234 and UV-326 significantly decreased the expression of pro-inflammatory cytokine tnfα. These molecular divergences underscore BUVs-specific modulation of inflammatory and metabolic networks in neural tissues. Yang et al. demonstrated that UV-328 exposure downregulated chrna7 and ache expression while upregulating gap43 in zebrafish, damaging neurons and the cholinergic system. Meanwhile, significantly increased expression of gria1a and drd1b confirmed that UV-328 induces neurotoxicity via the neuroactive ligand–receptor interaction pathway. Yu et al. found that UV-329 exposure significantly suppressed the expression of dop-3, mod-1, and unc-25 in Caenorhabditis elegans, reducing dopamine, serotonin, and GABA levels and resulting in neurodegeneration.

Furthermore, there is a strong interconnection between mitochondrial dysfunction and neurotoxicity. UV-234 disrupted mitochondrial gene networks, thereby compromising ATP production, calcium homeostasis, and redox balance. Mitochondrial dysfunction represents a critical convergence point for BUVs-induced neurotoxicity, linking compound-specific mechanisms to systemic neural damage.

4.6. Toxicity of Human Cells

Research on BUVs’ human toxicity remains scarce. However, existing evidence demonstrates their health hazards across multiple pathways, including disruption to cholesterol metabolism, epidermal growth signaling, glucose-lipid homeostasis, and cell cycle regulation. Shen et al. found that UV-234 (0.5–10 μmol/L) dose-dependently promoted macrophage foam cell formation. This occurred by degrading hypoxia-inducible factor via the ubiquitin-proteasome system, suppressing liver X receptor expression and its target genes ABCA1 and ABCG1, thereby blocking cholesterol efflux. Concurrently, UV-234 upregulated CD36 receptor expression, increasing cholesterol influx. This imbalance led to cholesterol accumulation, promoting atherosclerosis and elevating cardiovascular disease risk. UV-326 and UV-327 bound to the extracellular domain of epidermal growth factor receptor (EGFR) via hydrogen bonding and hydrophobic interactions, interfering with ligand-induced signal transduction. UV-327 (0.1–10 μmol/L) dose-dependently inhibited epidermal growth factor (EGF)-induced EGFR internalization and downstream extracellular-regulated kinase 1/2 (ERK1/2) phosphorylation. UV-326 (1–10 μmol/L) significantly suppressed ligand-induced DNA synthesis, impairing keratinocyte function. Liang et al. reported that UV-P and UV-329 significantly enriched pathways related to Type II Diabetes Mellitus, Insulin Resistance, and Cholesterol Homeostasis in human hepatocellular carcinoma cells (HepG2). They downregulated the expression of apolipoproteins genes (APOA1, APOC3), insulin-like growth factor, and albumin, thereby disrupting glucose homeostasis and lipid metabolism, suggesting a potential diabetogenic effect. Furthermore, UV-P disrupted the adipogenic/osteogenic differentiation balance in human mesenchymal stem cells (hMSCs) by inhibiting adipogenic markers (LPL, PPARγ) and upregulating osteogenic markers (RUNX2). This process involved nonclassical PPARγ activation and crosstalk with the AHR, posing threats to metabolic and skeletal health. Xu et al. revealed that all tested BUVs induced conserved toxicity by interfering with cell cycle arrest (prolonged G0/G1 phase) and inhibiting DNA replication. Their potency was positively correlated with chemical structural complexity and lipophilicity. Notably, these effects occurred at environmentally relevant concentrations in the μmol/L range.

4.7. Phytotoxicity

As primary ecosystem producers, plants absorb and accumulate pollutants through roots and leaves, playing a key role in pollutant bioaccumulation. Continuous environmental release of BUVs has led to their accumulation in plastic pollution hotspots. While quantification efforts increase, understanding of BUVs accumulation mechanisms in plants and associated toxicity remains scarce (Figure ).

4.

Toxicity of BUVs in plants. Schematic of BUVs’ key phytotoxic effects: reduced photosynthesis, inhibited stomatal closure, and enhanced metabolism. Soil-absorbed BUVs enter the roots, inducing toxicity. Photosynthesis decline primarily results from chloroplast swelling and thylakoid membrane disorganization.

BUVs disrupt photosynthetic machinery while accelerating carbon metabolic flux. In rice (Oryza sativa), UV-328 downregulated over 50% of chlorophyll-biosynthesis genes and upregulated degradation genes. Concurrently, BUVs enhanced glycolysis and TCA cycle activity, depleting small-molecular carbohydrates and accelerating carbohydrate and amino-acid metabolic flux while compromising antioxidant capacity. This metabolic imbalance between suppressive carbon fixation and hyperactive catabolism critically impairs plant growth.

BUVs exhibit biphasic effects on plant growth dependent on exposure concentration. Low UV-328 doses (50 μmol/L) stimulated thale cress (Arabidopsis thaliana) growth and photosynthesis, whereas high doses (150 μmol/L) caused chloroplast swelling, disrupted thylakoid ultrastructure, and inhibited light-harvesting gene expression (e.g., LHCA1, LHCA4, and LHCB3). Such hormetic responses underscore the ecological risks of concentration-dependent phytotoxicity.

Beyond direct photosynthetic effects, BUVs exposure damages stomata, impairing plant transpiration. BUVs disrupt stomatal regulation by competitively inhibiting Ca²⁺-dependent protein kinases (CDPKs), binding their active sites more strongly than ATP. This blocks ion channel protein phosphorylation, disrupting abscisic acid (ABA)-mediated stomatal closure. This results in uncontrolled transpiration, water loss, and growth retardation. The CDPK-ATP antagonism mechanism expands known pathways of synthetic chemical interference with plant stress adaptation.

5. Toxicology Mechanisms

The toxicology mechanisms of BUVs are attributed to their diverse chemical structures and can be classified into five primary pathways: the AHR pathway, ER/ERR activation, oxidative stress, lipid metabolism, and apoptosis (Figure ).

5.

Five key toxicology mechanisms of BUVs. The schematic diagram illustrates the five key toxicology mechanisms of BUVs, including their roles as AHR agonists, potentiators of endogenous estrogen signaling, disruptors of the antioxidant system, interferers with lipid homeostasis, and modulators of apoptosis-related pathways.

5.1. AHR Pathway

AHR interacts with a diverse array of ligands and signaling proteins, regulating multiple physiological functions, such as neurogenesis, endogenous metabolism, circadian rhythms, and immune modulation. Accumulating evidence indicates that specific BUVs (including UV-P, UV-PS, UV-9, UV-090, UV-326, and UV-329) can serve as AHR agonists, affecting various physiological responses. ,,,

Multiple studies confirm the AHR agonistic activity of BUVs. Research has indicated that UV-P, UV-9, UV-326 and UV-090 exhibit AHR ligand activity. Among them, UV-P, UV-9 and UV-090 are not actively metabolized by CYP1A1, indicating that they will continuously accumulate in the body and pose a persistent threat to the body’s health. Liang et al. demonstrated that UV-P and UV-329 (5, 25, and 50 μmol/L) significantly upregulated AHR target genes (CYP1A1, CYP1A2, UGT1A1, and UGT1A6) in HepG2 cells, with UV-328 elevating UGT1A1 and UGT1A6 only at 50 μmol/L. Luciferase assays verified UV-329 and UV-P as AHR agonists, while UV-328 was not identified as an AHR agonist. Complementing this, Johnson et al. reported that UV-P and UV-9 dose-dependently activated AHR2 and upregulated cyp1a transcripts in zebrafish embryo, whereas UV-090 showed no AHR2 activation. Notably, UV-P exhibited interspecies sensitivity differences that were absent upon exposure to UV-9 and UV-090. Li et al. found that exposure of zebrafish to 10 or 100 μg/L of BUVs can trigger immunotoxic responses via the AHR-IL17/IL22 pathway.

Emerging evidence highlights the need for systematic evaluation of diverse BUVs structures. Fent et al. identified UV-P and UV-326 as AHR activators in zebrafish embryo. Specifically, UV-P dose-dependently induced the expression of cyp1a1, ahr1, and arnt2. In contrast, UV-326 failed to induce ahr1 and arnt2 but significantly upregulated ahr2 and cyp1a1 expression at the concentration of 84 μg/L. Kubota et al. further screened 12 BUVs (UV-P, UV-PS, UV-9, UV-090, UV-234, UV-320, UV-326, UV-327, UV-328, UV-329, UV-350, and UV-360), revealing UV-P, UV-PS, UV-9, and UV-090 as potent AHR agonists. Notably, UV-PS exhibited about 10-fold stronger agonism than other analogs, emphasizing substituent-driven activity variations. AHR activation patterns depend on BUVs structural characteristics. Particularly, UV-PS acts as a high-potency agonist requiring prioritization in research.

5.2. ER/ERR Activation

Existing limited research indicates a close correlation between the endocrine toxicity of BUVs and ER/ERR activation. Feng et al. identified UV-9, UV-P, and UV-PS as weak ERα binders acting as partial agonists. Co-exposure with 1 nmol/L of the endogenous ligand E₂ in MVLN cells revealed dose-dependent synergism, amplifying the estrogen responsive element (ERE)-luciferase activity by 1.6-fold for UV-9, 2.1-fold for UV-PS, and 3.6-fold for UV-P versus E₂ alone. ER antagonist ICI_182,780 abolished this effect, confirming ER-specific mediation. Molecular docking confirmed some BUVs binding to ERα’s ligand pocket. These findings establish BUVs as ER coactivators capable of potentiating endogenous estrogen signaling through allosteric modulation.

Transcriptomic analysis and in vivo evidence revealed broad endocrine disruption of BUVs. Fent et al. demonstrated UV-P’s dual suppression in zebrafish: ectodermal embryo hormone receptor genes (pparα, trα, and hsdb3) were down-regulated, which was correlated with reduced estrogenic and androgenic activities. Such multireceptor interference underscores the necessity for pathway-level assessments in endocrine toxicity evaluation.

However, knowledge gaps persist in BUVs receptor interaction networks. Compared with the multiple experimental evidence of ER activation, , it remains unclear whether BUVs exhibit AR activation. Additionally, the latest research has suggested a close association between BUVs and ERR activation, yet there is a lack of more research reports on this. Therefore, it is urgently necessary to adopt a systems toxicology approach to decipher the complex molecular mechanisms of endocrine disruption by BUVs.

5.3. Oxidative Stress

BUVs induction of oxidative stress is well-established in hepatotoxicity, where impaired antioxidant defenses alter SOD, ROS, MDA, etc. Yang et al. discovered that under the influence of UV-328, the expression levels of fgf10a, fgf6a, and mp2k5 in zebrafish were significantly increased. This indicates that UV-328 can promote the production of ROS and induce oxidative stress by regulating the mitogen-activated protein kinase (MAPK) signaling pathway. Zhang et al. found that BUVs inhibited SOD and CAT activity, impaired ROS scavenging, and increased MDA levels. Different BUVs have specific effects on oxidative stress in Asian clam: UV-234 and UV-P (0.1, 1, and 10 μg/L) suppressed SOD activity, whereas low-dose UV-326 (0.1 μg/L) enhanced it, and UV-329 inhibited SOD activity at 1 μg/L but induced it at 10 μg/L. A similar phenomenon was also observed in zebrafish: the MDA content was significantly reduced at 1 μg/L UV-326 and 100 μg/L UV-234, respectively. Li et al. highlighted threshold effects in Asian clam, where UV-329 initially activated but ultimately overwhelmed antioxidant defenses at higher concentrations. Additionally, UV-328 and UV-234 demonstrated chronic toxicity in green algae (Chlamydomonas reinhardtii) and freshwater crustacean (Daphnia magna), significantly elevating ROS and upregulating gpx, sod, cat, and apx expression. These findings establish BUVs as disruptors of redox homeostasis, with effects modulated by chemical structure, concentration, and exposure duration.

Oxidative stress gene networks respond dynamically to BUVs exposure. UV-328 (0.01, 0.1, and 1 mg/L) was found to upregulate the genes associated with oxidative stress in zebrafish, including cyp1a and hsp70. Similarly, under the exposure to UV-328, the expression levels of sod, gst, cat and gpx were up-regulated in a concentration-dependent manner. Hemalatha et al. reported UV-328-induced upregulation of zebrafish hepatic oxidative stress markers (sod, cat, gpx, and gst), aligning with Zhang et al., who observed enhanced gpx-a, gsh-px, and gst activities post-BUVs exposure. , Transcriptomic analyses revealed UV-234, UV-326, UV-329, and UV-P alter expression of genes governing mitochondrial function, oxidative stress, and immunity. Notably, UV-234 uniquely modulated the expression of electron transport chain enzymes, including NADH dehydrogenase, coenzyme Q-cytochrome c reductase, and cytochrome c oxidase. Liang et al. demonstrated that UV-234 and UV-320 can influence mitochondrial bioenergetics and oxidative-stress responses in zebrafish embryo, with sod1 transcription downregulated and sod2 upregulated. These transcriptional signatures serve as early biomarkers for BUVs-induced oxidative stress, revealing both conserved and compound-specific response pathways.

5.4. Lipid Metabolism

At present, the evidence regarding the influence of BUVs on lipid metabolism is rather limited. Transcriptomic studies have revealed that BUVs interfere with lipid homeostasis. Liang et al. identified UV-329 as a potent disruptor in HepG2 cells. After a seven-day exposure at a concentration of 50 μmol/L, GO terms such as cholesterol homeostasis, lipid metabolic process, and liver development were significantly enriched. Complementarily, Li et al. discovered via gene enrichment analysis that UV-234 predominantly impacts biological pathways related to lipid and carbohydrate metabolism, as well as xenobiotic clearance.

Metabolomic profiling uncovers distinct BUVs-driven metabolic rewriting. Zhang et al. exposed zebrafish embryos to UV-234 and UV-326, revealing chemical-specific effects: UV-234 preferentially altered arachidonic acid and tryptophan metabolites, while UV-326 suppressed lipid metabolites but upregulated TCA cycle and oxidative phosphorylation intermediates. This metabolic shift suggested enhanced energy metabolism, potentially exacerbating fatty acid oxidation, thereby increasing lipid supply while decreasing lipid utilization and content. Such metabolomic divergence underscores structure–activity relationships in BUVs-induced lipid metabolic disruption.

5.5. Apoptosis

Apoptosis, a genetically regulated process of cell self-destruction, plays a crucial role in maintaining cellular homeostasis. BUVs can induce apoptosis via multiple mechanisms, including gene expression alteration, intracellular signaling pathway disruption, and mitochondrial damage. ,, As the concentration of UV-328 increased, the expression levels of apoptosis-related genes caspase-3 (0.1, 1, 10, and 100 μg/L), caspase-9 (100 μg/L) and p53 (0.1, 1, 10, and 100 μg/L) in zebrafish larvae gradually increased. However, in Asian clam, UV-329 (10 and 100 μg/L) initially activated Caspase-3/-8/-9 in gills and digestive glands, but suppressed these enzymes in digestive glands at 1000 μg/L, indicating tissue-specific toxicity thresholds. This demonstrates concentration-dependent apoptosis modulation, with high-dose BUVs overriding programmed cell death via cytotoxic overload.

BUVs exhibit a paradoxical capacity to coregulate pro/antiapoptotic signaling networks, resulting in conflicting outcomes in programmed cell death. UV-234 upregulated casp9, while downregulating p53, bax, casp3, and bcl-2, whereas UV-326 at equivalent concentration activated p53 but inhibited bax, casp3, and casp9. Zhang et al. observed concurrent antiapoptotic protein Bcl-2 elevation and pro-apoptotic gene suppression (tnfa, casp3, casp8, casp9, and cyto-c) in BUVs-exposed models. Transcriptomic profiling of UV-329 exposed Asian clams revealed tissue-specific gene activation patterns. In the gills, genes like casp9, dram2, as well as cytoskeletal genes (tuba1a and eif2s1) reached peak expression at 100 μg/L and 1000 μg/L, while digestive glands maximized mec-12 and tuba1a expression at 1000 μg/L. The paradoxical apoptotic signaling regulation may be attributed to negative feedback mechanisms, but the underlying molecular pathways require further investigation.

Multipathway analysis reveals conserved nodal targets. Kyoto Encyclopedia of Genes and Genomes (KEGG) enrichment analysis identified UV-234 as a potent disruptor of the nuclear factor-κB (NF-κB) signaling pathway and apoptosis in zebrafish, while UV-326, UV-329, and UV-P predominantly affected apoptosis and the tumor necrosis factor-α (TNF-α)/NF-κB pathway. Similarly, UV-234 induced an NF-κB-driven inflammatory response with enhanced levels of IL-1β and TNF-α, as well as caspase-mediated apoptosis, which was characterized by enhanced levels of Bax, Caspase-3, CytC, and a reduced level of Bcl-2. UV-328 has been found to induce apoptosis by regulating the p38-MAPK/p53 pathway in zebrafish. UV-329 additionally altered gap junction, autophagy, and necroptosis pathways in Asian clam, with metabolic pathways suppression coinciding with apoptosis activation. Membrane integrity loss and extracellular leakage further implicate microenvironmental dysregulation in BUVs-induced apoptosis.

6. Comprehensive Review of Environmental Risks

6.1. Aquatic Ecological Risks

The ubiquity of BUVs in the environment and growing awareness of their ecotoxicity have raised concerns regarding their ecological risks. Environmental risk assessment integrates the measured environmental concentration (MEC) of BUVs and the predicted no-effect concentration (PNEC) to quantitatively evaluate the threat level of chemical pollutants to aquatic organisms, providing critical data support for hazard risk evaluation. Several studies have adopted the risk quotient (RQ) or hazard quotient (HQ) methodology to assess environmental risks of BUVs, classifying risk levels as follows: RQ or HQ < 0.1 denotes low risk; 0.1 ≤ RQ or HQ < 1 indicates moderate risk; and RQ or HQ ≥ 1 represents high risk (Table S8).

The aquatic ecological risks of BUVs are currently not significant, with low-to-moderate risk areas accounting for 90% of existing studies. − Only isolated regions exhibit higher ecological risks, predominantly originating from sediments or wastewater. In the sludge of Indian wastewater treatment plants, the HQ of UV-328 exceeded 100, showing an extremely high environmental risk. Liu et al.’s assessment revealed that BUVs in Bohai Sea sediments generally pose low overall risks to aquatic organisms, except for UV-P and UV-PS. Specifically, UV-P in 23.0% and 8.2% of Bohai sediment samples may cause moderate risk to algae and low risk to fish, respectively, while UV-PS in one sample showed low risk to both species. In the Yangtze Estuary’s seawater and sediments, the maximum RQ value of UV-329 for fish reached 0.59, whereas other BUVs demonstrated no significant ecological risks. These findings reveal substantial ecological risk variations among BUVs congeners from differing lipophilicity or chemical stability, requiring comprehensive assessment.

Human activities and seasonal variations also warrant significant attention regarding the aquatic ecological risks of BUVs. In the Pearl River Basin adjacent to urban areas, UV-327 and UV-329 have been identified with high ecological risks. During dry seasons, UV-327, UV-328, and UV-326 similarly exhibited high-risk levels, whereas these risks decreased markedly in rainy seasons, indicating direct or indirect dilution effects from precipitation. Furthermore, disparities in aquatic trophic levels lead to differential risk manifestations: algae frequently show higher RQ values than fish due to direct pollutant exposure and high reproductive rates; , however, higher-trophic organisms may accumulate risks through biomagnification in food chains. ,

6.2. Human Exposure Risks

With the deepening research on the biological toxicity of BUVs, their potential human health risks have attracted increasing concern. Methods such as estimated daily intake (EDI), margin of exposure (MOE), and HQ are commonly employed to quantify health risks. ,,

Differences in exposure pathways (dermal contact, dietary intake, and air inhalation) and BUVs concentrations lead to variations in health risk levels. Moualek et al. assessed the risk of BUVs exposure through dietary intake of redfish (Sebastes mentella), revealing an HQ < 2.3 × 10^–2, indicating a low human health risk via this pathway. A Korean study found that the MOE values for all population groups exceeded 100 after consuming various foods, suggesting minimal public health threats from dietary BUVs.

Compared to adults, infants and young children face significantly higher susceptibility due to their lower body weight and weaker metabolic capacity. Notably, the highest EDI of UV-P through breast milk reached 1804 ng/(kg·bw·day), within the same order of magnitude as the reference dose (RfD). Studies showed that Philippine infants’ BUVs EDI from indoor dust ingestion was 5 times higher than that of adults. For infant sunscreen products, the 95th percentile dermal exposure of UV-P approached or exceeded the RfD.

Importantly, inhaled doses of BUVs are typically 2–3 orders of magnitude higher than dermal exposure, indicating inhalation may be a more critical exposure route. Deng et al. evaluated BUVs concentrations in Beijing’s atmosphere and found median EDI values 5–12 orders of magnitude below the RfD, implying negligible health risks. However, adults in plastic greenhouses exhibited EDI levels 2–4 orders of magnitude higher than those exposed to dust in Tianjin, China. In greenhouses covered with biodegradable plastic films, the maximum inhalation HQ for UV-328 reached 0.4, indicating a moderate risk. These findings demonstrate that although current human health risks from BUVs remain manageable, potential risks should not be overlooked due to cumulative or synergistic effects of multiple BUVs and heightened hazards to vulnerable groups (such as infants and occupationally exposed populations). Therefore, rigorous monitoring and restrictions on BUVs usage through regulations are urgently needed.

7. Conclusions and Prospects

BUVs are established global contaminants, ubiquitous in environmental matrices (surface water, wastewater, sediment, air, dust, soil, organism). Their distribution shows significant regional and seasonal variations driven by urbanization, wastewater discharge, and hydrology. Notably, UV-328, UV-329, UV-234, and UV-P persist at high levels, with air and dust as the most contaminated compartments. Human exposure is confirmed via detection in breast milk, urine, and adipose tissue, underscoring pervasive contamination.

Toxicological studies (predominantly freshwater vertebrates) reveal that BUVs exert multifaceted organ-specific effects. Specifically, endocrine disruption occurs via ER/ERR and AHR activation. BUVs lead to developmental defects, including cardiac abnormalities and embryo mortality. On the immunological front, BUVs elicit inflammatory responses and dysregulation of immune genes. Hepatotoxicity is mediated by oxidative stress and apoptosis. BUVs induce neurotoxicity by causing mitochondrial dysfunction and altering locomotor behavior. Notably, endocrine toxicity is currently the most frequently and thoroughly studied toxic effect. Additionally, BUVs demonstrably harm human health by disrupting cholesterol metabolism, epidermal signaling, metabolic homeostasis, and cell cycle regulation. In plants, BUVs impede photosynthesis and accelerate carbon metabolism, thus posing a threat to ecosystem productivity.

BUVs exert toxicity through interconnected signaling networks, with UV-P, UV-234, UV-329, and UV-326 being the most studied. These exhibit diverse toxicological profiles and mechanisms, including AHR activation leading to disrupted xenobiotic metabolism and endocrine signaling. Endocrine disruption arises from ER/ERR agonism and antiandrogenic activity, notably potent estrogenic effects from UV-P. Oxidative stress underlies tissue damage and inflammation. Additional mechanisms involve disrupted lipid metabolism, mitochondrial dysfunction, and caspase-mediated apoptosis.

Current environmental risks of BUVs are assessed as low-to-moderate and manageable. However, intensifying anthropogenic pressures, coupled with trophic transfer and bioaccumulation in food webs, may precipitate future ecological disturbances. Proactive regulatory intervention is thus warranted. Human health impacts demand urgent attention, especially given infants’ heightened vulnerability to bioaccumulative and synergistic effects.

Significant knowledge gaps persist in BUVs research. Occurrence data beyond sediments and air remain scarce, geographically limited, and temporally sparse. Toxicity mechanisms remain poorly understood, many effects are uncharacterized, and underlying molecular pathways require elucidation. Future priorities should: (1) Expand monitoring to understudied regions and matrices; (2) Assess long-term effects at environmentally relevant concentrations; (3) Extend biotoxicity assessment beyond freshwater vertebrates; (4) Apply multiomics to decipher species-specific mechanisms. Addressing these gaps will elucidate BUVs’ ecological and health risks, informing mitigation strategies.

Glossary

Abbreviations

BUVs

benzotriazole ultraviolet stabilizers

DOM

dissolved organic matter

dw

dry weight

BCF

bioconcentration factor

WWTPs

wastewater treatment plants

ND

not detected

ww

wet weight

AHR

aryl hydrocarbon receptor

ER

estrogen receptor

ERR

estrogen-related receptor

AR

androgen receptor

IL

interleukin

ROS

reactive oxygen species

MDA

malondialdehyde

SOD

superoxide dismutase

CAT

catalase

GSH-PX

glutathione peroxidase

EGFR

epidermal growth factor receptor

HepG2

human hepatocellular carcinoma cells

MAPK

mitogen-activated protein kinase

Nf-κB

nuclear factor-κB

TNF-α

tumor necrosis factor-α

RQ

risk quotient

HQ

hazard quotient

EDI

estimated daily intake

MOE

margin of exposure

RfD

reference dose

Data will be made available on request.

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/envhealth.5c00139.

• Data associated with this article including Tables S1–S8 (PDF)

T.L. and Y.F.: data curation, investigation, visualization, and writing-original draft; L.Z.: investigation, visualization, and writing-review and editing; Z.T.: data curation, investigation, and writing-original draft; K.F.: visualization and writing-review and editing; D.C.: funding acquisition and writing-review and editing; P.J.: supervision and writing-review and editing; and M.G.: conceptualization, supervision, funding acquisition, writing-original draft, and writing-review and editing.

The authors declare no competing financial interest.

Published as part of Environment & Health special issue “New Pollutants: Challenges and Prospects”.