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. 2023 Dec 19;58(1):110–120. doi: 10.1021/acs.est.3c06117

Assessing the Toxicity of Benzotriazole Ultraviolet Stabilizers to Fishes: Insights into Aryl Hydrocarbon Receptor-Mediated Effects

Hunter M Johnson †, Justin Dubiel †, Cameron H Collins ‡, Andreas N M Eriksson †, Zhe Lu §, Jon A Doering †,‡,*, Steve Wiseman †,*
PMCID: PMC10785820  PMID: 38112502

Abstract

graphic file with name es3c06117_0003.jpg

Benzotriazole ultraviolet stabilizers (BUVSs) are chemicals used to mitigate UV-induced damage to manufactured goods. Their presence in aquatic environments and biota raises concerns, as certain BUVSs activate the aryl hydrocarbon receptor (AhR), which is linked to adverse effects in fish. However, potencies of BUVSs as AhR agonists and species sensitivities to AhR activation are poorly understood. This study evaluated the toxicity of three BUVSs using embryotoxicity assays. Zebrafish (Danio rerio) embryos exposed to BUVSs by microinjection suffered dose-dependent increases in mortality, with LD50 values of 4772, 11 608, and 56 292 ng/g-egg for UV-P, UV-9, and UV-090, respectively. The potencies and species sensitivities to AhR2 activation by BUVSs were assessed using a luciferase reporter gene assay with COS-7 cells transfected with the AhR2 of zebrafish and eight other fishes. The rank order of potency for activation of the AhR2 from all nine species was UV-P > UV-9 > UV-090. However, AhR2s among species differed in sensitivities to activation by up to 100-fold. An approximate reversed rank order of species sensitivity was observed compared to the rank order of sensitivity to 2,3,7,8-tetrachlorodibenzo[p]dioxin, the prototypical AhR agonist. Despite this, a pre-existing quantitative adverse outcome pathway linking AhR activation to embryo lethality could predict embryotoxicities of BUVSs in zebrafish.

Keywords: species sensitivity, embryotoxicity, quantitative adverse outcome pathway, AhR, CYP1A, microinjection, plastics-associated chemicals

Short abstract

Little is known regarding the potency and species sensitivities to AhR activation by BUVSs in fish. This study shows that UV-P, UV-9, and UV-090 differ in potency, and species differ in their sensitivities to AhR activation by these BUVSs.

1. Introduction

Benzotriazole ultraviolet stabilizers (BUVSs) are a class of chemical contaminants that have recently raised concerns among regulatory bodies due to their persistence, bioaccumulation, biomagnification, and toxic characteristics.1−4 BUVSs are characterized by the presence of a primary benzotriazole moiety attached to a 2-hydroxyphenol group, with significant structural diversity resulting from substituents attached at various positions.5−7 Due to their photostability, BUVSs are additives in industrial and consumer goods, including cosmetics, coatings, adhesives, waxes, paints, motor oils, rubber, and plastics, to protect against degradation and discoloration caused by full-spectrum ultraviolet light (280–400 nm).1,2,7−14 BUVSs are hydrophobic with logarithmic octanol–water partition coefficients (log Kow) greater than 3, providing a potential for environmental accumulation by sorbing to sediment or accumulating in fatty tissues of aquatic organisms, where they can undergo bioaccumulation and biomagnification.11,15−17 BUVSs enter the environment through manufacturing processes, wastewater effluent, sewage, and leaching from waste products.17,18 As a consequence, BUVSs are ubiquitous in the environment, having been detected in surface waters, sediment, and tissues of biota including birds, invertebrates, mammals, and fishes.1−3,9,11,12,17−22 Concentrations of BUVSs in freshwater rivers range from 0.7 to 701 ng/L.19,23,24 Furthermore, sediment concentrations of BUVSs range from 0.16 to 35 μg/g dried weight.3,7,19,25 Specifically, 2-(2-hydroxy-5-methylphenyl)benzotriazole (UV-P), 2-(2H-benzotriazol-2-yl)-4-methyl-6-(2-propenyl)phenol (UV-9), and 2-[3-(2H-benzotriazol-2-yl)-4-hydroxyphenyl]ethyl methacrylate (UV-090) have been detected in sediment at concentrations ranging from 15 to 800 ng/g dried weight and in water at concentrations from 0.9 to 28.1 ng/L at several locations globally.7,13,24 Additionally, concentrations of UV-P and UV-9 in muscle tissues of fish have been reported to range from 2.2 to 9.1 ng/g lipid weight and 1.3–14 ng/g wet weight, respectively.24,26 Fish that are developing in BUVS-contaminated water or sediment, or that are potentially exposed to maternally transferred BUVSs, could experience adverse effects during their early life-stages.2,13,26−28

The current empirical understanding of the adverse effects of exposure to BUVSs on aquatic wildlife, including fishes, is limited to a few studies. Transcriptomic and molecular studies of fish exposed to BUVSs revealed potential dysregulation of the hypothalamic–pituitary–thyroid axis, antiandrogenic activity, disruption of steroidogenesis, and activation of the aryl hydrocarbon receptor (AhR).28−32 The AhR is a ligand-activated transcription factor that regulates the expression of genes involved in a multitude of physiological processes, such as xenobiotic metabolism, cellular growth, and cell migration.13,33−36 Several isoforms of the AhR exist in vertebrates, but in fishes, there is strong evidence that activation of the AhR2 isoform is causative of toxicities in early life-stages, including mortality, spinal and cranial malformations, yolk sac and pericardial edema, and cardiac dysfunction.37−40 As fishes transition out of their early life-stages, numerous toxicities have been associated with activation of the AhR2 including wasting syndrome, hepatotoxicity, and fin necrosis.41−43 Regarding BUVSs, a prior study reported no significant increase in mortality or malformations in zebrafish (Danio rerio) embryos exposed to waterborne UV-P or 2-(3-tert-butyl-2-hydroxy-5-methylphenyl)-5-chlorobenzotriazole (UV-326), but there was increased transcript abundance of cytochrome P450 1A (cyp1a), a widely accepted biomarker for activation of the AhR.28,44 In contrast, another study found no significant increase in the transcript abundance of cyp1a in adult Japanese medaka (Oryzias latipes) exposed to UV-P spiked food.29 The differences in these effects of UV-P on transcript abundance of cyp1a could be due to the variability in the AhR2 structure among fish species which can translate into a vast difference in relative sensitivities (ReSs) among species to the same agonist.45,46 In addition to differences in ReSs, relative potencies (RePs) of AhR agonists can exceed several orders of magnitude within a species when compared to the prototypical AhR agonist 2,3,7,8-tetrachlorodibenzo[p]dioxin (TCDD).47 Lastly, screening of 13 structurally diverse BUVSs for their ability to activate the AhR of mice found that UV-P, UV-9, UV-090, and 2-(5-tert-butyl-2-hydroxyphenyl)benzotriazole (UV-PS) were AhR agonists, indicating that some, but not all, BUVSs can act as agonists of the AhR.6

Potencies of BUVSs as agonists of AhR2 in fishes are unknown. Since other classes of AhR agonists display a wide range of interspecies differences in potencies for activating the AhR, it is likely that a similar diversity exists for BUVSs. As such, investigating interspecies sensitivity to AhR activation by BUVSs will enable a more informed approach to the environmental risk assessment of BUVSs that encompasses a wide array of fish species. Therefore, the primary aim of this study was to determine if BUVSs can exert toxicity in fishes through activation of the AhR2. As an initial step toward accomplishing this, zebrafish embryos, a commonly used model test species, were exposed to three serial doses of UV-P, UV-9, or UV-090 through microinjection and assessed for AhR-mediated toxicities and response of cyp1a. The second goal of this study was to determine whether, and to what extent, there are interspecies differences in the sensitivity to activation of the AhR2 by testing zebrafish and Japanese medaka, two model species that are not native to North America, and seven freshwater species native to North America, namely, brook trout (Salvelinus fontinalis), fathead minnow (Pimephales promelas), lake sturgeon (Acipenser fulvescens), lake trout (Salvelinus namaycush), northern pike (Esox lucius), white sucker (Catostomus commersonii), and white sturgeon (Acipenser transmontanus) by quantifying activation of the AhR2 using a standardized in vitro luciferase reporter gene (LRG) assay of COS-7 cells transfected with the AhR2 of each species. Results of this study will help inform whether activation of the AhR2 by BUVSs could represent a risk to native populations of fishes.

2. Material and Methods

2.1. Embryotoxicity

2.1.1. Animal Care and Embryo Collection

Collection of embryos complied with the University of Lethbridge Animal Welfare Protocol AWP#2114, with full details provided in the Supporting Information. Briefly, embryos were obtained from a breeding culture of zebrafish (Tupfel long-fin strain) approximately 1 h postfertilization (hpf), and viable eggs were pooled for use in microinjections.

2.1.2. Embryo Microinjection

Physiochemical properties of UV-P (purity > 97%), UV-9 (purity > 99%), and UV-090 (purity > 99%) are provided in Table S1. Each BUVS was purchased from Sigma (Mississauga, Ontario, Canada) and prepared at a nominal concentration of 15 mg/mL (high) in DMSO, which is near their maximal solubility. As a first step toward better understanding the toxicities of BUVSs in fish, the nominal concentrations were serially diluted 3-fold to generate additional dosing solutions at 5 mg/mL (medium) and 1.67 mg/mL (low).

Microinjections were performed based on previously described methods48 with modifications49,50 with full details provided in the Supporting Information. Briefly, an IM-400 Electric Microinjector (Narishige Group, Tokyo, Japan) was calibrated to administer approximately 1.5 nL of DMSO or BUVS solution before injection of embryos. Each dose of the BUVSs was injected directly into the yolk sac before the completion of gastrulation (<6 hpf). Three experimental replicates of approximately 100 zebrafish embryos per replicate were injected for every treatment group. Control embryos were injected with 1.5 nL of full-strength DMSO. Each experimental replicate included a DMSO control and three BUVS treatment groups and used embryos from independent breeding events. Lastly, 1 g of embryos (approximately 2500 embryos) was injected per concentration of each BUVS and DMSO control, then immediately frozen at −80 °C for quantification of dose (Section 2.3).

2.1.3. Embryo Rearing and Assessment

Full details regarding embryo rearing and assessment are provided in the Supporting Information. Following injection, embryos were incubated for 24 h in plastic Petri dishes containing dechlorinated City of Lethbridge municipal water at 26 ± 1 °C. After 24 h, any embryos that were nonviable or dead were discarded and not included in the final mortality data. Twenty-four viable embryos were randomly selected and transferred into independent wells of a 24-well plate (Eppendorf Canada, Mississauga, ON, Canada) containing 2 mL of water at 26 ± 1 °C and were reared until complete yolk sac utilization at 15 days postfertilization (dpf).49 Daily water renewals of 50% were performed. Remaining embryos from the high dose of each of the BUVSs were reared until 5 dpf, the completion of hatch, and then allocated into groups of 10 and immediately frozen at −80 °C for quantitative real-time polymerase chain reaction (qPCR). For embryos transferred to the 24-well plates, daily assessments of pericardial and yolk sac edema, spinal curvature, and mortality were performed for each embryo by use of a Zeiss Discovery.V12 stereo microscope (Carl Zeiss Canada, Toronto, ON, Canada). Images were captured using a Zeiss Axiocam 105 Color (Carl Zeiss Canada, Toronto, ON, Canada) and ZEN lite imaging software (Carl Zeiss Microscopy, Oberkochen, Germany).

2.1.4. Quantitative Real-Time PCR

Transcript abundance of cyp1a was quantified in 5 dpf embryos exposed to the highest dose of each BUVS to determine activation of the AhR. Details are provided in the Supporting Information.

2.2. Luciferase Reporter Gene Assay

The LRG assay followed a previously described protocol51 with modifications.52 In short, immortalized COS-7 cells that lack an endogenous AhR pathway were chemically transfected with 8 ng of species-specific AhR2,38,45,52,53 1.55 ng of white sturgeon ARNT2,52 20 ng of rat CYP1A reporter construct,54,55 and 0.75 ng of renilla luciferase vector to assess transfection efficiency (Promega, Madison, Wisconsin, USA) per well. In previous studies, the use of the ARNT of a single species did not affect species sensitivities to TCDD so it is unlikely to affect the results in the present study.38 Additionally, several isoforms of the AhR (AhR1 & AhR2) exist in vertebrates because of gene duplication and diversification events.35,56,57 The AhR2 was chosen because activation of this isoform has been linked with early life stage mortality and malformations in fish.37,38,40 After transfection, cells were exposed to 9 nominal concentrations of UV-P, UV-9, or UV-090, ranging from 0.3 to 30 000 nM, which were prepared in 100% DMSO from the 15 mg/mL stock solutions described in Section 2.1.1. The concentration ranges of BUVSs were chosen with the goal of covering the entire concentration–response curve. However, due to the low potency of BUVSs, the greatest tested concentration of 30 000 nM was selected based on concerns about cytotoxicity and solubility at greater concentrations as observed during initial range finding studies (data not shown). Each chemical-specific assay was completed in triplicate with each replicate consisting of 4 independent wells per concentration. A SpectraMax i3x plate reader (Molecular Devices, San Jose, California, USA) was used to measure the luminescence of the luciferase reporter. Maintenance, transfection, dosing of COS-7 cells, and plate reading were performed at the SynBridge core facility at the University of Lethbridge.

2.3. Analysis of BUVSs in Eggs

Chemical analysis was performed to determine doses of each BUVS in zebrafish eggs following microinjections, as previously described.29 Full details are provided in the Supporting Information.

2.4. Statistics and Data Analysis

Measured doses were used for all analyses of the in vivo data. Differences in the cyp1a transcript abundance between treatments and controls were determined using IBM SPSS Statistics 20 software. Data was assessed for normality using the Shapiro–Wilk test and homogeneity of variance using a Levene test. A Kruskal–Wallis test was used based on outputs from normality and homogeneity of variance tests. GraphPad Prism 9 software v.9.4.1 for windows (GraphPad Software, San Diego, California, USA) was used to generate in vivo dose–response curves. Curves were fit to a four-parametric logistic model, and doses that caused 0, 10, and 50% lethality (LD0, LD10, and LD50) were calculated for zebrafish embryos exposed to each BUVS. Control background mortality was normalized to 0 using Min–Max Scaling under the “Normalize” function in GraphPad Prism. A Fisher exact test, conducted using GraphPad Prism, was used to identify statistically significant differences in mortality (p < 0.05) between DMSO-injected and BUVS-injected embryos. Figures were generated using GraphPad Prism.

Statistical analysis for in vitro AhR activation was performed in R v.4.2.2 (The R Foundation for Statistical Computing, 2022) coupled to RStudio v.2022.12.0.353 (Rstudio Team 2022) as described in Dubiel et al.58 Outliers were determined by preforming a Chi-squared test (statistical cutoff: p ≤ 0.05) using the Tests for Outliers package and then removed from the finalized data set.59 Dose–response curves were generated in Rstudio and fit using a four-parameter log–logistic function using the Analysis of Dose–response Curves package,60 and effective concentration threshold (ECthreshold) of activation of the AhR2 in COS-7 cells was established. The ECthreshold was identified as the first chemical concentration that caused a significant increase in response (p ≤ 0.05) from the DMSO control, with all subsequent doses also having significant increases in response. The ECthreshold was selected because it is independent of the concentration–response curve reaching a maximal response. Half maximal effective concentration (EC50) was not calculated because BUVSs did not reach a clear maximal response in most, if not all, species. Due to few replicates (n = 3), normality was not assessed and significant differences between treatments and controls were determined using the Kruskal–Wallis H test combined with Dunn’s post hoc test not adjusted for multiple comparisons when assessing ECthreshold. Relative potencies and ReSs were calculated by dividing the ECthreshold of the reference chemical or species by the ECthreshold of the chemical or species of interest. Providing the ECthreshold generated by the LRG assay along with the molar mass of each chemical as input variables into the qAOP model provided in Doering et al.61 gives the predicted LD10, LD50, and LD100 as output variables. The accuracy of the qAOP for predicting dose–response curves for BUVSs from the ECthreshold was evaluated using mean absolute error (MAE), mean absolute percentage error (MAPE), and fold difference. Concentrations used for the LRG assay were carefully designed as 3-fold dilutions, a choice made to minimize the intervals between each concentration. This enhances the precision of the ECthreshold measurement, ensuring it remains within a range less than 3-fold, all while guaranteeing the capture of the entire concentration–response curve.

3. Results

3.1. Embryotoxicity

Exposure of zebrafish embryos to UV-P, UV-9, or UV-090 caused a dose-dependent increase in mortality with a rank order of potency of UV-P > UV-9 > UV-090 (Table 1). The LD50 values were 4772, 11 608, and 56 292 ng/g-egg for UV-P, UV-9, and UV-090, respectively (Table 1). In addition to mortality, the proportion of embryos exhibiting malformations (yolk sac edema, pericardial edema, and spinal curvature) increased following exposure to each chemical (Table 2). However, the occurrence of malformations did not increase significantly with increasing doses of BUVSs. Representative images of the observed malformations are provided (Figure S1). Transcript abundance of cyp1a was assessed to confirm in vivo activation of zebrafish AhR2. Zebrafish embryos exposed to the maximal dose of UV-P and UV-9, which caused 51 and 73% embryo mortality, respectively, showed a 5.9 and 42.2-fold significant increase in cyp1a transcript abundance, respectively (Figure S2). There was no significant increase in the transcript abundance of cyp1a in embryos exposed to the maximal dose of UV-090, which caused 48% embryo mortality, compared to the control group (Figure S2).

Table 1. Calculated and Predicted Lethal Doses (LDs) (ng/g-egg) That Caused 10 and 50% Mortality of Zebrafish Embryos Exposed to Doses of UV-P, UV-9, or UV-090a,b,c.

Chemical Calculated LD0 Calculated LD10 Calculated LD50 Predicted LD0 Predicted LD10 Predicted LD50
UV-P 112 (0.28–822) 854 (10–2630) 4772 (3064–8313) 71 148 426
UV-9 1323 (270–6493) 1724 (489–6073) 11 608 (8687–16747) 187 401 1212
UV-090 634 (55–7367) 7190 (1327–38956) 56 292 (38656–123551) NAd NA NA
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a

The predicted LD0, LD10, and LD50 were established using a quantitative adverse outcome pathway for low-potency agonists of the AhR.61 The range of values presented in brackets represent the 95% confidence interval.

b

LD values are calculated based on measured doses of each BUVS.

c

Predicted values are calculated on metrics outlined in Table S4.

d

NA = Not applicable.

Table 2. Measurement Parameters of Zebrafish Embryos Used in Early Life Stage Toxicity Testing That Were Exposed to a DMSO Control and Each of the BUVSs (UV-P, UV-9, and UV-090)a,b.

Treatment Nominal dosing solution concentration (ng/nL) Nominal embryo concentration (ng/g-egg) Measured embryo concentration (ng/g-egg) Malformations (%) Mortality (%) Mortality normalized to control
DMSO control NDc ND ND 10 (5) 32 (8) 0 (0)
  1.67 4167 3090 13 (9) 51 (19)d 31 (20)
UV-P 5.00 12 500 2700 14 (8) 56 (17)d 36 (19)
  15.00 37 500 5210 18 (8) 65 (14)d 51 (13)
  1.67 4167 5910 22 (7) 56 (2)d 24 (14)
UV-9 5.00 12 500 10 500 22 (12) 74 (5)d 56 (8)
  15.00 37 500 37 500 22 (14) 84 (5)d 73 (9)
  1.67 4167 10 400 14 (5) 46 (4) 12 (11)
UV-090 5.00 12 500 18 800 6 (5) 46 (12)d 23 (13)
  15.00 37 500 63 400 19 (2) 58 (19)d 48 (18)
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a

The values presented in brackets are mean ± standard deviation (±SD).

b

Malformations: spinal curvature, yolk sac edema, and pericardial edema.

c

ND = Not Detected.

d

Significant difference between control mortality and treatment mortality (p < 0.05, Fisher’s exact test).

3.2. AhR2 Transactivation

The AhR2 of zebrafish was activated in a concentration-dependent manner by UV-P and UV-9 but not UV-090 (Figure 2 and Table 3). The rank order of potency for activation of the zebrafish AhR2 was UV-P > UV-9 > UV-090 (Table 3), and the ECthreshold to activate the zebrafish AhR2 was 1000 nM for UV-P and 3000 nM for UV-9. UV-P activated the AhR2s of each of the other eight species in a concentration-dependent manner (Figure 2); the ECthreshold ranged from 300 nM in white sucker to 30 000 nM in Japanese medaka (Table 3 and Figure 2). UV-9 only activated the AhR2s of zebrafish, lake sturgeon, and northern pike in a concentration-dependent manner, and each had an ECthreshold of 3000 nM (Figure 2 and Table 3). Lastly, UV-090 did not activate the AhR2 of any of the tested species (Figure 2 and Table 3). The rank order of potencies for AhR2 activation was UV-P > UV-9 > UV-090 for each of the species tested, except for lake sturgeon and northern pike, where UV-P and UV-9 were equipotent (Table 3).

Figure 2.

Figure 2

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Dose–response curves of COS-7 cells transfected with the AhR2 of (A) white sucker, (B) zebrafish, (C) lake sturgeon, (D) northern pike, (E) brook trout, (F) fathead minnow, (G) lake trout, (H) white sturgeon, and (I) Japanese Medaka following the exposure to UV-P, UV-9, or UV-090. Data is presented as mean ± standard error (±SE) based on three independent assays, each performed with four replicates per chemical concentration. 95% confidence intervals are represented by gray shaded area and vertical dotted lines represent ECthreshold.

Table 3. Calculated Effective Concentrations That Pass the Minimum Threshold for Activation (ECthreshold) of Each Investigated Species-Specific AhR2 by UV-P, UV-9, or UV-090abc.

Sspecies Chemical ECthreshold (nM)
White Sucker UV-P 300
UV-9 NAd
UV-090 NA
TCDD 1.0d
Zebrafish UV-P 1000
UV-9 3000
UV-090 NA
TCDD 1.0
Lake Sturgeon UV-P 3000
UV-9 3000
UV-090 NA
TCDD 0.1d
Northern Pike UV-P 3000
UV-9 3000
UV-090 NA
TCDD 1.00d
Brook Trout UV-P 10 000
UV-9 NA
UV-090 NA
TCDD 0.10d
Fathead Minnow UV-P 10 000
UV-9 NA
UV-090 NA
TCDD 0.3d
Lake Trout UV-P 10 000
UV-9 NA
UV-090 NA
TCDD 0.03d
White Sturgeon UV-P 10 000
UV-9 NA
UV-090 NA
TCDD 0.03d
Japanese Medaka UV-P 30 000
UV-9 NA
UV-090 NA
TCDD 0.3d
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a

NA = Not applicable.

b

Data obtained from Doering et al.38

c

Data obtained from Doering et al.45

d

Data obtained from Doering et al.52

Species sensitivity to AhR2 activation by UV-P ranged 100-fold based on the ECthreshold, with a rank order of sensitivity of white sucker > zebrafish > lake sturgeon = northern pike > brook trout = fathead minnow = lake trout = white sturgeon > Japanese medaka (Table 3). No such difference in species sensitivity for AhR2 activation by UV-9 or UV-090 was observed. The predicted LD50 values of UV-P, derived from the qAOP for low-potency agonists of the AhR,61 were 162, 426, 1028, 1028, 2701, 2701, 2701, 2701, and 6522 ng/g-egg for white sucker, zebrafish, lake sturgeon, northern pike, brook trout, fathead minnow, lake trout, white sturgeon, and Japanese medaka, respectively (Table S3). Zebrafish, lake sturgeon, and northern pike had predicted LD50 values of 1212 ng/g-egg for UV-9 (Table S3).

4. Discussion

BUVSs are an environmentally relevant class of persistent contaminants, but current understanding of their toxicity to fishes is limited. There is a growing body of evidence that some BUVSs, including UV-P, can act as agonists of the AhR based on studies of the human AhR in yeast reporter assays and increase of cyp1a expression in zebrafish.13,28,63 In contrast, studies of Japanese medaka show no evidence of AhR activation by UV-P.29 This suggests that BUVSs have the potential to cause adverse effects via activation of the AhR, but that potency might differ among species. These differences in potency could result from species-specific differences in sensitivities to activation of the AhR by BUVSs that are driven by differences in the structure of the AhR protein. However, whether potencies of BUVSs for activation of the AhR differ among fish species was unknown. Therefore, the present study explored the toxicities of three commonly detected BUVSs to zebrafish embryos and their ability to activate the AhR of zebrafish and eight other species as a mode of toxicity. Results suggest that UV-P and UV-9, but not UV-090, are low-potency agonists of the AhR2 of zebrafish and cause AhR-mediated toxicity in embryos, but there likely are substantial differences in sensitivities to BUVSs across distinct species of fish.

Exposure to UV-P, UV-9, and UV-090 caused dose-dependent increases in mortality among zebrafish embryos, with potencies relative to TCDD that are comparable to low-potency agonists of the AhR, such as coplanar polychlorinated biphenyls (PCBs; Figure 1 and Table 4).64 The most potent BUVS, UV-P, had an LD50 of 4772 ng/g-egg, followed by UV-9 with an LD50 of 11 608 ng/g-egg, and the least potent compound, UV-090, with an LD50 of 56 292 ng/g-egg (Figure 1 and Table 1). While derived from just three serial concentrations, these LD50 values serve as an initial foundation for understanding the potencies of BUVSs. Furthermore, this is the first study to demonstrate mortality following exposure to BUVSs, but prior studies used waterborne or dietary exposure which likely resulted in a lesser effective dose relative to microinjection.28,29,63,65 In addition to mortality, zebrafish embryos exposed to UV-P, UV-9, and UV-090 in the present study showed increased incidences of AhR-mediated toxicity related to blue sac disease, which encompasses yolk sac and pericardial edema, and spinal curvature.37,40 Further, transcript abundance of cyp1a increased in zebrafish embryos exposed to UV-P and UV-9, which aligns with previous findings and further suggests activation of the AhR (Figure S2).13,28 In contrast, zebrafish embryos exposed to UV-090 did not show an increased transcript abundance of cyp1a (Figure S2). Results from the in vitro AhR transactivation assay show activation of the zebrafish AhR2 by UV-P and UV-9 with ECthreshold values of 1000 and 3000 nM, respectively, but zebrafish AhR2 was not activated by UV-090 up to the greatest tested concentration of 30 000 nM (Table 3 and Figure 2). Activation of the zebrafish AhR2 by UV-P and UV-9, but absence of any activation by UV-090, is consistent with the cyp1a response in embryos and suggests that the embryo mortality following exposure to UV-P and UV-9 is, at least partially, mediated by the AhR. Considering the absence of increased cyp1a transcript abundance in vivo and the lack of AhR activation in the LRG assay, these findings strongly suggest that mortality of zebrafish embryos exposed to UV-090 occurs by a mode of toxicity that is independent of AhR activation. The in vitro AhR transactivation assay utilized COS-7 cells that lack an endogenous AhR pathway and have essentially no intrinsic expression of cyp1a that is responsible for phase I biotransformation of xenobiotics.66 Therefore, it is unlikely that UV-090 is being hydroxylated through phase I metabolism in vitro. Although the mechanism by which UV-090 caused embryotoxicity is not currently known and is beyond the scope of the present investigation, toxicities related to blue sac disease can occur in fish through AhR-independent mechanisms and should be further investigated.67−69

Figure 1.

Figure 1

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Dose–response curves generated from early life stage mortality of zebrafish embryos (15 dpf) that were exposed to (A) UV-P, (B) UV-9, or (C) UV-090 (green triangles and dotted line). Data is presented as mean ± standard error (±SE) based on three replicate studies, each with n = 24 ± 4 embryos per dose of chemical. Data is normalized for the control background mortality to equal 0. Intersect of long dashed lines indicates calculated lethal doses causing 50% mortality (LD50) for each BUVS. Red circle and solid line dose–response curve is a prediction of the LD50 generated by using the quantitative adverse outcome pathway of low-potency AhR agonists based on ECthreshold.61

Table 4. Potencies of BUVSs (UV-P, UV-9, UV-090) Relative to TCDD for Each Species Investigated in This Study Based on the Calculated ECthreshold and EC50 Values.

Species Chemical ReP to TCDD Order of potency
White Sucker UV-P 0.003 UV-P > UV-9 & UV-090
UV-9 NAa
UV-090 NA
Zebrafish UV-P 0.001 UV-P > UV-9 > UV-090
UV-9 0.0003
UV-090 NA
Lake sturgeon UV-P 0.00003 UV-P & UV-9 > UV-090
UV-9 0.00003
UV-090 NA
Northern pike UV-P 0.0003 UV-P & UV-9 > UV-090
UV-9 0.0003
UV-090 NA
Brook trout UV-P 0.00001 UV-P > UV-9 & UV-090
UV-9 NA
UV-090 NA
Fathead Minnow UV-P 0.00003 UV-P > UV-9 & UV-090
UV-9 NA
UV-090 NA
Lake trout UV-P 0.000003 UV-P > UV-9 & UV-090
UV-9 NA
UV-090 NA
White sturgeon UV-P 0.000003 UV-P > UV-9 & UV-090
UV-9 NA
UV-090 NA
Japanese Medaka UV-P 0.00001 UV-P > UV-9 & UV-090
UV-9 NA
UV-090 NA
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a

NA = Not applicable.

In fishes, interspecies differences in sensitivities to activation of the AhR2 can span several orders of magnitude between the most sensitive and least-sensitive species.38,58 Specifically, TCDD and other chlorinated dioxin-like compounds show clear differences in species sensitivities to AhR2 activation and associated early life toxicities, with the zebrafish AhR2 being among the least sensitive to activation by TCDD.38,70 As such, it is likely that zebrafish do not provide the optimal representation of sensitivities for native species of environmental and regulatory concern. Therefore, activation of the AhR2 of Japanese medaka, another model species, and seven native fish species, brook trout, fathead minnow, lake sturgeon, lake trout, northern pike, white sucker, and white sturgeon, by UV-P, UV-9, and UV-090 was determined in the in vitro AhR transactivation assay. Across all species there was a 100-fold range of sensitivity to activation of the AhR2 by UV-P, with ECthreshold values ranging from 300 to 30 000 nM (Table 3). The rank order of species sensitivities was white sucker > zebrafish > lake sturgeon = northern pike > brook trout = fathead minnow = lake trout = white sturgeon > Japanese medaka (Table 3). In a previous study, exposure of Japanese medaka, the species with the least-sensitive AhR2 to activation by BUVSs, to UV-P showed no evidence of AhR activation in vivo based on abundance of cyp1a transcript and AhR-mediated embryotoxicities.29 In studies with zebrafish, which possess one of the most sensitive AhR2s to activation by BUVSs in the transactivation assay, there was evidence of activation of the AhR2 based on increased transcript abundance of cyp1a.(13,28) These results suggest that the differences in species sensitivities observed in the in vitro transactivation assay translate to differences in sensitivity to AhR activation observed in vivo. TCDD has a similar range of sensitivity to activation of the AhR2 among species, being 33-fold based on the ECthreshold (Table 3). However, the rank order of sensitivity to TCDD—lake trout > white sturgeon > brook trout > fathead minnow > lake sturgeon > Japanese medaka > white sucker > zebrafish > northern pike—is almost a reversed rank order of sensitivity to UV-P.38,52 Similarly, UV-9 activated the AhR2 of the fishes that were among the least sensitive to TCDD (zebrafish, lake sturgeon, northern pike) but not the fishes that were among the most sensitive to TCDD (white sturgeon, brook trout, lake trout) (Table 3). However, there was no difference in the sensitivities to activation of the AhR2 among zebrafish, lake sturgeon, and northern pike which all had an ECthreshold of 3000 nM (Table 3). Due to the constraints of the LRG assay, it is possible for the ECthreshold values to exhibit an error of up to 3-fold. In terms of risk assessment, when considering that the maximal disparity in sensitivity to AhR activation is also 3-fold, it can be inferred that the species sensitivity to activation of the AhR by UV-9 remains essentially consistent between these three species. In contrast, UV-090 did not activate the AhR2 of any of the species at any concentration tested (Figure 2). Taken together and based on the fishes tested here, results for UV-P and UV-9 are unique in that the rank order of species sensitivities to activation of the AhR2 appears to be approximately reversed when compared to TCDD. Therefore, species considered at least risk to chlorinated dioxin-like compounds could be at greatest risk to BUVSs that are agonists of the AhR. Several species are known to have AhRs that are substantially less sensitive to activation by TCDD relative to the fishes studied here and could be at elevated risk, including Atlantic cod (Gadus morhua), seabirds, and amphibians.71−74

Challenges related to assessing the risk of chemical pollutants based on measurement of early molecular changes, such as in vitro AhR transactivation, led to the development of adverse outcome pathways (AOPs) which provide a systematic way to understand and describe a sequence of events that lead from molecular-level occurrences, termed the molecular initiating event (MIE), to an adverse outcome (AO) of regulatory concern.38,75 A mechanism-based biological model, termed a quantitative adverse outcome pathway (qAOP), quantifies the magnitude of the MIE to predict the likelihood or severity of subsequent AOs. As a result of variations in sensitivities among species to AhR activation, a qAOP has been developed that utilizes ECthreshold from the standardized in vitro AhR transactivation assay utilized in the present study, which represents the MIE in this qAOP, to predict full dose–response curves of early life stage mortality among fishes and other taxa, which denotes the AO of this qAOP, to within an order of magnitude, on average.61 This qAOP has been validated for a suite of chlorinated dioxin-like compounds that range in potency by more than 30 000-fold.61 Recently, it was demonstrated that this qAOP cannot predict early life toxicities for PAHs, likely resulting from complexity related to biotransformation.49 However, whether this qAOP could accurately predict the dose–response curves of BUVSs was unknown. By the use of ECthreshold values from the present study as an input into the qAOP model, the predicted LD50 values for UV-P and UV-9 in zebrafish were 426 and 1212 ng/g-egg (Table S3) while the calculated LD50 values were 4772 and 11 608 ng/g-egg, respectively (Figure 1 and Table 1). The fold difference between the predicted and experimental data for the average of LD0, LD10, and LD50 values was 6.2 lower for UV-P and 7.0 lower for UV-9 (Table S4). The MAEs for BUVSs are slightly higher than what had been previously reported for low-potency agonists of the AhR; however, the MAPE and fold difference are within the range of accuracy determined for other low-potency agonists using this model and represents an order of magnitude estimate commonly employed in risk assessments.61 Therefore, these findings provide the first support for the qAOP61 having a chemical applicability domain beyond chlorinated dioxin-like chemicals and suggest that this model could be used to generate order of magnitude estimates of the toxicity of BUVSs to other species using ECthreshold from the in vitro AhR transactivation assay.

Concentrations of BUVSs in freshwater ecosystems can range from 0.7 to 701 ng/L, and dried-weight sediment concentrations can range from 0.16 to 35 μg/g.3,7,13,19,23−25 Concentrations of certain BUVSs (UV-P, UV-329, UV-326, UV-234, UV-328, and UV-327), as high as 377 ng/g lipid weight have been detected in the belly fat of fishes and approximately 0.9, 0.64, and 0.64 ng/g median wet weight in the blood plasma of bottlenose dolphins (Tursiops truncatus), snapping turtles (Chelydra serpentina), and double-crested cormorant (Phalacrocorax auritus), respectively.11,62,76 Results from the present study suggest that in vivo toxicity of BUVSs begins to occur at nearly 1500 ng/g for zebrafish embryos, which were one of the most sensitive species to AhR2 activation by BUVSs (Table 1). Therefore, based on the current reported environmental concentrations of BUVSs and the results of the present study, it is unlikely that BUVSs pose a significant threat to fishes through AhR-mediated pathways. A previous study found that several BUVSs, including UV-P, accumulated in tissues, including ovaries, of zebrafish with a bioconcentration factor exceeding 10.65 Given the lipophilicity of BUVSs, if the widespread use of BUVSs continues to increase globally, bioaccumulation and biomagnification of BUVSs could result in elevated exposure to BUVSs which, in turn, could lead to AhR-mediated effects. Japanese medaka, the least-sensitive species tested in the AhR transactivation assay had no evidence of AhR activation in in vivo studies, suggesting that other species found to have AhRs that are insensitive, such as brook trout, fathead minnow, lake trout, and white sturgeon, are similarly unlikely to suffer from AhR-mediated toxicities from BUVSs. However, the evidence of zebrafish having sensitivities similar to PCBs suggests that other sensitive species, such as white sucker, northern pike, and lake sturgeon, might suffer AhR-mediated toxicities to BUVSs at toxicity thresholds similar to PCBs. If the reverse rank order of species sensitivity to AhR2 activation by BUVSs holds true in vivo, this becomes more problematic for other species that possess AhRs that are substantially less sensitive to TCDD than the species tested in the present study, such as Atlantic cod, seabirds, and amphibians.71−74 Therefore, sensitivity of AhR activation in these species should be determined, as they could be at a greater risk for BUVS-induced AhR-mediated toxicities. To the best of our knowledge, this study presents the first evidence that AhR agonists beyond chlorinated dioxin-like chemicals are applicable to the previously developed qAOP.61 Furthermore, this study shows promise that the qAOP61 can be used to pragmatically assess sensitivities to BUVSs to a diversity of species from data generated from a standardized in vitro AhR transactivation assay. Utilizing this model could be essential for efficiently assessing potential risks posed by BUVSs to other species, including those species that could be at the greatest risk.

Acknowledgments

S.W. was supported by a Tier II Canada Research Chair in Aquatic and Mechanistic Toxicology, infrastructure grants from the Canadian Foundation for Innovation John R. Evans Leaders Fund (CFI-JELF; Project #35224), and Government of Alberta Research Capacity Program. S. Wiseman (Grant #: RGPIN 2017-04474) and Z. Lu (Grant #: RGPIN-2019-05761, DGECR-2019-00268) were supported by Discovery Grants from the Natural Sciences and Engineering Research Council of Canada and the Natural Sciences and Engineering Research Council of Canada’s Plastics Science for a Cleaner Future program (Grant #: ALLRP-558410-20). JAD received support from the NSERC Postdoctoral Fellowships program (PDF-546121-2020). Support from staff in the Aquatic Research Facility at the University of Lethbridge is acknowledged and appreciated. Thanks to M. Denison (University of California, Davis, CA, USA) for donation of rat CYP1A1 reporter construct and to R. Tanguay (Oregon State University, Corvallis, OR, USA) for donation of the zebrafish AhR2 expression construct.

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.3c06117.

  • Animal care and embryo collection; embryo microinjections; embryo rearing and assessment; quantitative real-time PCR; analysis of BUVSs in eggs; chemical properties of BUVSs (Table S1); primer sequence, efficiency, target mRNA, and accession number for oligonucleotideprimers (Table S2); predicted 20, 50, and 100% lethal doses for UV-P and UV-9 (Table S3); metrics used to calculate the predicted lethal doses causing 0, 10, and 50% mortality for both UV-P and UV-9 (Table S4); images of control larvae and larvae exposed to BUVSs (Figure S1); effect of exposure to maximal concentration of BUVSs (Figure S2); dose–response curve of activation of the AhR2 of zebrafish by TCDD in COS-7 cells transfected with the AhR2 (Figure S3) (PDF)

The authors declare no competing financial interest.

Supplementary Material

es3c06117_si_001.pdf (285.3KB, pdf)

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