Immunomodulation in adult largemouth bass (Micropterus salmoides) exposed to a model estrogen or mixture of endocrine disrupting contaminants during early gonadal recrudescence

Highlights

• •Environmentally relevant concentrations of estrogenic endocrine-disrupting contaminants can alter immune function in adult LMB, a socioeconomically important fish species.
• •Species-specific baseline information for comparison to fishes collected from impacted environmental settings.
• •These findings warrant further investigation regarding the effects of these contaminants on disease resistance in adult LMB.

Abstract

Disease outbreaks, skin lesions, fish kill events, and reproductive abnormalities have been observed in wild populations of Centrarchids in watersheds throughout the United States. Occurrence of synthetic and natural hormones from wastewater treatment plants and livestock operations, pesticides from agricultural land use, and phytoestrogens have been implicated as potential causes of these adverse effects. Our objective was to investigate possible immunomodulation in adult largemouth bass (Micropterus salmoides) in response to a seasonal exposure to environmentally relevant contaminants in outdoor experimental ponds. Exposures included 17α-ethinylestradiol (EE2; 3.6 ng/L) or a binary mixture of endocrine-active substances commonly detected in surface waters, estrone (E1; 85.6 ng/L) and atrazine (ATR; 5.4 µg/L). The 4-month exposure was conducted from July to November. Functional immune responses of anterior kidney-derived leukocytes were evaluated in December in the week following the end of the dosing period, and in the following April, four months after dosing ended and just prior to spawning. Concentrations of EE2 and E1 in the ponds fell below detectable levels in December, but detectable concentrations of ATR (2.9 µg/L) persisted at least through May. For each sampling time, anterior kidney leukocytes were isolated and grown in primary culture for the assessment of zymosan-stimulated respiratory burst and lectin-stimulated mitogenic responses. We observed seasonal differences in respiratory burst stimulation over time and treatment with a significantly greater response in April relative to December. Respiratory burst activity was also significantly greater in April for fish exposed to the E1+ATR relative to control. In April, prior to spawning, we observed a significantly dampened mitogenic response to PHAP (a T cell mitogen) and LPS (a B cell mitogen) in the EE2 treatment relative to control fish. There were no significant differences in mitogenic responses or respiratory burst between sexes. However, there was significantly higher alternative complement pathway hemolytic activity in males compared to females in both the control and E1+ATR treatment groups. Our results demonstrate that environmentally relevant concentrations of contaminants can alter immune function in a socioeconomically important fish species.

Introduction

Black basses are the prototypical sportfish in North America and are of great ecological and socioeconomic value [1], [2], [3]. While fisheries management research has investigated the impacts of angling and water temperature on black bass wellbeing, less is known about the health consequences of environmental stressors such as exposure to environmental contaminants [4]. Reproductive abnormalities have been observed in wild fish throughout the United States, and a high occurrence of intersex has been reported in largemouth bass (LMB, Micropterus salmoides) and smallmouth bass (Micropterus dolomieu) [5], [6], [7], [8]. In addition to intersex, health concerns have been raised for black bass populations in regions where estrogenic endocrine disrupting compounds (EEDCs) have been reported. A well-studied region in the United States where black basses are purportedly affected by EEDCs is the Chesapeake Bay watershed [9]. Basses in this watershed have a high prevalence of intersex, and mortality events have been observed that are typically associated with skin lesions, opportunistic bacteria, viruses, and parasitic pathogens [10], [11], [12], [13], [14]. The timing of mortality events and opportunistic nature of associated pathogens indicated a role for anthropogenic chemical inputs during high runoff events as potential contributors to immunosuppression [10].

Atrazine (ATR) is used globally as a pre-emergent broadleaf herbicide for the agricultural production of crops. It has been categorized as an endocrine disruptor and is associated with biological effects reported to modulate normal neuroendocrine, reproductive, and immune function [15], [16], [17], [18], [19]. Abnormalities in male bass gonads have been correlated with the presence of ATR in regions where mortality events have also been observed [13]. In areas of high agricultural land use, average ATR concentrations can exceed 5 µg/L, with spikes in the spring potentially above 20 µg/L [14,[20], [21], [22]]. In aquatic environments, organic contaminants such as ATR typically do not occur independently [23]. Estrone (E1) is a naturally occurring estrogen that can enter surface waters through runoff from feedlots or land application of animal wastes [24,25]. Elevated concentrations of E1 commonly co-occur with ATR in agriculturally influenced watersheds and can be found at concentrations that spike above 100 ng/L in the springtime [14,22]. These concentrations of E1 are associated with reproductive abnormalities in fish and are of concern for overall fish reproductive health [24,[26], [27], [28]].

Surface water contamination with 17α-ethinylestradiol (EE2) is often detected in areas with high urban land use and presence of wastewater treatment plants. Concentrations of EE2 in these areas are frequently below 1.0 ng/L but have been observed to reach up to 10 ng/L [29]. Gonad abnormalities, decline in reproductive condition, and altered sex ratios have been observed in feral fish populations downstream from wastewater treatment plants in waters with measured concentrations of <1 ng/L EE2 [30]. Exposures to EEDCs are well-documented to modulate development, reproduction, and immunity in fish [31]. However, it is not known at what times during the reproductive cycle bass are most sensitive to exogenous hormone and/or contaminant exposure, or if effects are reversible when exposure ends. LMB have seasonal, relatively synchronous spawning events that begin in spring once water temperatures reach 17 °C or above [32]. Early gonad recrudescence occurs post-spawning from late summer to fall and is the period when spermatogenesis and oogenesis are beginning in preparation for spawning in the following spring [33,34]. Gonad recrudescence in fish is a time of rapid gamete proliferation and can be a sensitive window of exposure to EEDCs [35,36]. The focus in the current study was to evaluate the potential effects of a potent estrogen and an EEDC mixture on immune function in adult LMB following chronic aqueous exposure during early gonad recrudescence. EE2 was chosen as a positive control because it has been well studied as a model EEDC with strong estrogen receptor agonist activity and is a compound relevant to contamination from wastewater treatment plants [30,37]. Our objective was to assess whether early gonad recrudescence was a period of sensitivity for alterations in immune function associated with exposure to EEDCs. Adult LMB were exposed in outdoor experimental ponds from post-spawning through early gonad recrudescence to either EE2, or a mixture of E1 + ATR. Leukocytes isolated from the anterior kidney were stimulated ex vivo to evaluate functional responses in the week following the end of the dosing period (December) and the following spring, four months after dosing ended and just prior to spawning (April). Alternative complement pathway hemolytic activity was also assessed just prior to spawning (April).

Methods

Exposure and animal care

Pond exposures, sample processing, and data analysis were conducted at the U.S. Geological Survey Columbia Environmental Research Center (CERC, Columbia, Missouri, USA), and the functional immune endpoint analysis was performed at the U.S. Geological Survey Eastern Ecological Science Center (EESC, Kearneysville, West Virginia, USA) after tissue shipment overnight. This study complied with all applicable sections of the Final Rules of the Animal Welfare Act regulations (9 CFR) and all CERC Institutional Animal Care and Use Committee guidelines for the humane treatment of the test organisms during culture and experimentation.

The exposure was conducted in nine 50′ × 72′ (0.08 acre), earthen-lined experimental ponds each containing approximately 110,000 gallons of water. These ponds were filled with well water prior to the study. Conditions were maintained within criteria set forth by the American Standards and Testing Materials [38] for toxicity testing with aquatic organisms. General water quality averaged 200 mg/L hardness, 165 mg/L alkalinity, 8.1 pH, and 0.0300 mg/L ammonia throughout the study. Details about water quality testing and pond descriptions can be found in Leet et al. 2022 [39].

Adult LMB were stocked into the ponds and acclimated for three months prior to the start of chemical exposure. All fish were sexed and distributed in the ponds so that each pond contained 35 fish of similar length and weight, and each pond was randomly assigned one of three treatments (control, EE2, or E1 + ATR) with three replicate ponds per treatment. Ponds were dosed from the end of July through the end of November.

Dosing and water chemical analysis

Dosing solution concentrations were confirmed by a liquid chromatography/mass spectrophotometer (LC/MS) [39]. Pond concentrations were monitored weekly during the exposure period (end of July through the end of November) and monthly once dosing stopped. The ponds were dosed weekly with either ethanol only for the solvent control ponds and those ponds that did not need additional chemical, or they were dosed with a calculated amount of stock solution (in ethanol) to bring the total pond concentration back to the desired nominal concentration of that treatment. Further details about dosing, chemical analysis via LC/MS, and monitoring of chemical concentrations can be found in Leet et al. 2022 [39]. The test chemicals (17α-ethinylestradiol, CAS no. 57–63–6; estrone, CAS no. 53–16–7; atrazine, CAS no. 1912–24–9; >98 % purity) were purchased from Sigma Aldrich (St. Louis, Missouri, USA). Carbon-13/14 internal standards were acquired from Cambridge Isotope Laboratories in solution (2,3,4–13C3-estrone [13C3-E1], 20,21–13C2-ethynylestradiol [13C2-EE2], and atrazine-ring-UL-13C, >99 % purity) at 100 µg/mL in methanol. The solvents (acetonitrile, methanol, and water) were LC/MS grade, and the ammonium hydroxide (28 % in water) was from Fisher Scientific (Hampton, Virginia, USA).

Pond water samples were collected in 250 mL solvent rinsed, amber glass bottles, and 200 mL of controls or sample were immediately filtered through glass fiber filters (Whatman 0.7 µm GF/F, Fisher Scientific) using a solid phase extraction (SPE) vacuum manifold (Biotage, Uppsala, Sweden). For E1 and EE2 treatments, water samples were then extracted via C18 SPE cartridges (Strata C18-E, 1 g, 20 mL, Phenomenex) and then analyzed. For atrazine analysis, water samples were analyzed directly after filtration.

Fish collection and processing

At each sampling timepoint fish were seined from each pond until three or four of each sex were reached, as we did not want to skew the sex ratio of the fish remaining in each pond. Sample numbers for each endpoint can be found in the figure legends. Fish were euthanized with an overdose (300 mg/L) of buffered MS-222 (Ethyl 3-aminobenzoate methanesulfonate, Sigma Aldrich) for fish metric measurements and tissue sample collection. At the December and April timepoints, anterior kidney was aseptically dissected. At the pre-spawning timepoint in April, venipuncture was performed via manual restraint from the lateral approach to the caudal vein. Heparinized sterile 5-mL syringes with 25-gauge needles were used to collect approximately 5 mL of blood. Blood samples were centrifuged at 2000 × g for 10 min at 4 °C to separate the plasma, then stored in 1 mL aliquots at −20 °C until analysis.

Leukocyte isolation

Unless otherwise noted, all media components were obtained from Sigma Aldrich (St. Louis, Missouri, USA). Fish were aseptically necropsied, and one lobe of the anterior kidney was removed and placed into processing medium (PM; isotonic Leibovitz-15 medium supplemented with 2 % fetal bovine serum (FBS), 100 U/mL penicillin, 100 µg/mL streptomycin, and 10 U/mL sodium heparin) and stored on wet ice. Tissues were shipped overnight to the EESC, Fish Health Laboratory in Leetown, West Virginia, for processing.

Anterior leukocytes were isolated and processed as described previously with minor modification [40]. Tissues were processed within 24 h of excision from experimental fish. Anterior kidney leukocytes were mechanically dissociated from anterior kidney tissue via repeated passage through the barrel of a 3-cc syringe. Tissue preparations were resuspended in PM and allowed to settle for 30 min on wet ice. Supernatants were transferred to new sterile polypropylene conical tubes and centrifuged at 500 g for 10 min at 4 °C. Cells were washed by suspension in PM followed by centrifugation as above and suspended in PM. Cell suspensions were layered onto 32 % Percoll in Hanks Balanced Salt Solution without phenol red, pH 7.2 (HBSS). The cells were centrifuged at 500 g for 40 min at 4 °C. The leukocyte fraction was removed and washed. The number of viable leukocytes was determined by the trypan blue (0.1 % trypan blue in PM) exclusion method using a Countess (ThermoFisher, Waltham, Massachusetts, USA). Leukocytes were suspended at 2 × 10⁷ viable cells/mL in culture medium (CM; l-15 media supplemented with 5 % FBS, 100 U/mL penicillin, and 100 μg/mL streptomycin) or adherence medium (AM; l-15 media supplemented with 0.1 % FBS, 100 U/mL penicillin, and 100 μg/mL streptomycin). All tissues and cell suspensions were maintained at 4 °C in an ice bath during processing.

Mitogenesis

Anterior kidney leukocytes suspended in CM were dispensed to 96-well tissue-culture plates (Costar #3396 plates; Corning Inc., Corning, New York, USA) at 1 × 10⁶ cells per well. Anterior kidney leukocytes were plated such that each 96-well plate contained cells from all treatment groups to account for possible plate effect variability. A volume of 50 μL/mL of CM with or without mitogen (control) was added to the wells immediately after cells were plated. Mitogen concentrations used to stimulate the isolated leukocytes were as follows: 10 μg/mL Concanavalin A (CONA, a T cell mitogen), 10 μg/mL Phytohaemagglutinin P (PHAP, a T cell mitogen), or 100 μg/mL lipopolysaccharide (LPS, a B cell mitogen) from Escherichia coli 0111:B4. Mitogen-treated and control wells were replicated in triplicate and plating was performed with the plates on ice. Following plating, leukocytes were incubated in humidified chambers at 10 °C under atmospheric conditions [41]. The mitogenic response was measured on the fourth day (96 h) of incubation post-mitogen stimulation.

The mitogen-induced proliferative response was evaluated using the BrdU-based enzyme-linked immunosorbent assay (ELISA) at room temperature, as described by Gauthier et al. [42] with minor modifications. Briefly, 24 h prior to the ELISA, leukocytes were treated with 25 μL per well of sterile-filtered 65 μM BrdU in l-15 and incubated for an additional 24 h. Plates were centrifuged at 500 g for 10 min at 4 °C, cells were gently washed with 100 μL per well Dulbecco's phosphate-buffered saline (DPBS) and fixed for 15 min with 1 % paraformaldehyde (Sigma Aldrich) in DPBS (pH 7.2). Cells were washed three times as above, and the cell membranes were porated by a 30 s exposure to 50 μL per well of 0.01 % polyoxyethylene-sorbitan monolaurate (Tween-20; Sigma Aldrich) in DPBS. Cells were washed three times, and the wells were then blocked with 250 μL per well of blocking buffer (DPBS containing 1 % bovine serum albumin) for 1 h. Blocking buffer was removed and replaced with 50 μL per well of 0.2 U/mL anti-bromodeoxyuridine Fab fragments labeled with horseradish peroxidase (Roche Chemicals) in blocking buffer and incubated for 1 h. Cells were washed five times with 100 μL per well DPBS, then 50 μL of an enzyme development solution (EDS; 10 mM sodium citrate buffer, 728 μM 2,2′-azino-bis (3-ethylbenzthiazoline-6-sulphonic acid), 0.03 % H₂O₂ and pH 4.0) were added to each well. The optical density (405 nm) of the solution in each well was determined at 5 min increments over a period of 20 min using an SpectraMax M4 (Molecular Devices Corporation, Sunnyvale, California, USA).

Stimulation index (SI) values for individual fish were calculated as the replicate mean optical density for a given set of mitogen-treated leukocytes divided by the replicate mean optical density of the associated mitogen free (control) leukocytes. SI values were calculated for all time points, and the maximum SI (mSI) value for a given pair of mitogen-treated and control leukocytes was defined as the highest SI value determined for the four time points within a treatment group.

Respiratory burst

The production of extracellular reactive oxygen species (ROS) was determined using the nitroblue tetrazolium (NBT) reduction assay [43]. Briefly, 1 × 10⁶ leukocytes suspended in AM were added to the wells of a 96-well plate and incubated at 10 °C for 2 h. The assay was run at 10 °C as a projected mean temperature between winter and spring pond water temperatures. Media was then removed from all wells, replaced with CM, and incubated in a humidified chamber at 10 °C for 36 h to allow cells to attain a resting state. On the day of the assay, culture medium was removed, and cells were gently washed with unsupplemented l-15. Cells were then treated with zymosan (1 mg/mL) for 1 h prior to determination of NBT reduction. In short, NBT was dissolved in dimethyl sulfoxide; DMSO, 20 %, w/v) and unsupplemented l-15 prewarmed to 50 °C to completely dissolve NBT to a final concentration of 2 mg/mL. This NBT solution was heated for an additional 10 min at 50 °C and filter sterilized. Zymosan A was added to the NBT solution at a final concentration of 1 mg/mL. An NBT solution without zymosan A was prepared to assess basal ROS production and discriminate between induced versus basal production for all experimental treatments. All solutions were cooled to 10 °C prior to the assay. The NBT/zymosan solution (100 µL) was added to the adherent cell cultures and were incubated for 1 h, followed by removal of the supernatant. Cells were fixed with the addition of 70 % methanol and room temperature incubation for 5 min. Unreduced NBT was removed by washing cells five times with 70 % methanol. Reduced NBT was dissolved with the addition of 60 µL of 2 M KOH to each well followed by vigorous pipetting. Seventy microliters of DMSO was added to all wells, and the optical densities were read at 630 nm (OD₆₃₀) on a SpectraMax M4 multimode microtiter plate reader (Molecular Devices). SI values for individual fish were calculated as the replicate mean optical density for a given set of zymosan-treated leukocytes divided by the replicate mean OD₆₃₀ of the unstimulated control.

Alternative complement pathway analysis

The alternative complement pathway (ACP) activity was assayed using rabbit red blood cells (RRBC; Innovative Research, Novi, Michigan, USA; IC10–0510 - 10 mL) as targets (adapted from Sunyer and Tort [44]). RRBC suspension (3 %) diluted in veronal buffer (Boston BioProducts, Milford, Massachusetts, USA; IBB-310X-250 mL) were mixed with serial dilutions of plasma (20 µl RRBC + 120 µL plasma dilution). The spontaneous hemolysis (negative control) was obtained by adding 20 µL of RRBC to 120 µL of veronal buffer. The total lysis (positive control) was obtained by adding 40 µL of RRBC to 240 µL of distilled water. After incubation for 100 min at 27 °C, the samples were centrifuged at 2000 g for 10 min at 4 °C. Supernatant from each sample (40 µL) was added to 40 µL of distilled water in a 96-well plate. Samples were run in duplicate. The absorbance (Abs) was measured at 405 nm on a BioTek Synergy 4 (Winooski, Vermont, USA), and the mean absorbance for replicate samples was used for analysis. Percent lysis was determined by% lysis = ((Abs405 test - Abs405 blank)/ (Abs405 total lysis - Abs405 blank)). The reciprocal of the plasma dilution causing 50 % lysis (ACH50) was calculated as an estimate ACP hemolytic activity using a probit analysis [44]. For each sample a four-point probit curve was graphed using JMP 14.2.0, and the inverse prediction was used to calculate the plasma dilution corresponding to ACH50. A quality control (QC) sample with a known ACH50 was added to each plate. Data from a plate was not used if the QC sample, positive control, or negative control absorbance readings were not within two standard deviation units of mean plate to plate variation (mean and deviation based on intra-laboratory variation data).

Statistical analysis

Data were tested for normality using the Shapiro–Wilks W-test and homogeneity of variance via the Brown–Forsythe test of homogeneity of variances. Mitogen and respiratory burst data were analyzed using a Kruskal–Wallis test, and significant differences between group means were determined using the Duncan multiple range test using GraphPad PRISM. ACP data were analyzed using one-way ANOVA, and significant differences between group means were determined using a Tukey's test using JMP 14.2.0. Differences were considered statistically significant when p ≤ 0.05. Raw data can be found through the U.S. Geological Survey and is publicly available at https://doi.org/10.5066/P9U2U3A1, https://doi.org/10.5066/P9LDPEJF, https://doi.org/10.5066/P91G7KMO.

Results

Exposure

Mean ± standard error (SE) water concentrations in the treatment ponds were 85.6 ± 6.9 ng/L E1, 5.4 ± 0.1 µg/L ATR, and 3.6 ± 0.4 ng/L EE2 over the dosing period (end of July through the end of November) [39]. Within a month after ceasing dosing on the first of December, the mean ± SE concentrations of E1 and EE2 were 2.9 ± 0.9 ng/L and 0.1 ± 0.04 ng/L, respectively. However, ATR continued to persist at a mean of 2.9 ± 0.6 µg/L throughout the remainder of the study (December to May). The mean temperature across the ponds was 5 ± 1 °C during December and 13 ± 3 °C during April. There were no differences observed between treatments for any water quality parameters assessed. All raw water quality, water chemistry, and fish metric data for this experiment can be found at https://doi.org/10.5066/P9U2U3A1 (accessed on 28 August 2023).

Respiratory burst

There were no significant differences observed in respiratory burst activity between sexes. Therefore, data for both sexes were pooled for seasonal analysis. Seasonal differences in respiratory burst activity were observed for all treatments (p < 0.0001), so seasons were analyzed independently. During December there were no differences in respiratory burst across treatments. In general, the SI response for respiratory burst was low and ranged between 0.7 and 2.9 across all samples collected in December. During April, the respiratory burst response of anterior kidney leukocytes from the E1 + ATR exposed fish was significantly greater (p = 0.0045) than that from the control group (Fig. 1). The range of SI for respiratory response across all samples collected in April was 0.9 – 7.3. Unnormalized cell yields across treatments during December ranged from 2.2 – 4.6 × 10⁷ cells/ anterior kidney; mean = 3.31 × 10⁷ cells/ anterior kidney. Cell yields were significantly higher in April and ranged from 3.4 – 7.7 × 10⁷ cells/ anterior kidney; mean = 5.59 × 10⁷ cells/ anterior kidney.

Fig. 1.

Respiratory burst stimulation index in adult largemouth bass at the December timepoint in the week following the end of the dosing period (Control n = 15, EE2 n = 16, E1 + ATR n = 17), and the April timepoint four months following the end of the dosing period and just prior to spawning (Control n = 20, EE2 n = 21, E1 + ATR n = 19). Asterisks denote significant differences compared to control (p < 0.005). Data presented as violin plots where the dashed line represents the mean and dotted lines represent the quartiles.

Mitogenesis

There were no significant differences in mitogenic stimulation indices observed between sexes. Therefore, data for both sexes were pooled for seasonal analysis. Seasonal differences were observed in mitogenesis activity. No differences were observed in mitogenesis during December just after the end of the dosing period, and in general, stimulation indices were low (not shown). A combination of low cell yields and the unavailability of immunophenotyping reagents precluded cell population analysis. Based on semi-quantitative measures, cell yields were significantly lower in fish sampled during December. In April, four months after the end of dosing, the EE2-exposed fish had significantly lower mitogenic responses to PHAP (a T cell mitogen; p = 0.028) and LPS (a B cell mitogen; p = 0.031) relative to control fish (Fig. 2).

Fig. 2.

Mitogenesis stimulation in adult largemouth bass at the April time point just prior to spawning (Control n = 20, EE2 n = 21, E1 + ATR n = 19). Asterisks denote significant differences compared to control (p < 0.05). Data presented as violin plots where the dashed line represents the mean and dotted lines represent the quartiles.

Alternative complement pathway

Sex differences were observed in the ACP activity in plasma from LMB at the April sampling, four months after exposures ended and just prior to spawning. This difference was statistically significant in the control fish (p = 0.012) and E1+ATR exposed fish (p = 0.015), with a trend of lower ACP activity in the females than males across all treatment groups (p = 0.0004). The difference between sexes was not statistically significant in the EE2 exposure group. There were no significant changes in ACP activity with either EE2 or E1 + ATR exposure compared to controls.

Discussion

Immune response activities in fishes and other poikilothermic organisms are modulated by seasonal changes in environmental conditions, including temperature [45,46]. Seasonal differences in functional immune responses and cell yields of anterior kidney leukocytes were observed in LMB in the current study. In general, functional immune responses were low in all treatment groups, including controls, during the December sampling compared to April. It is likely that treatment differences in the measured endpoints were not observed in December due to holistic changes in cell metabolism and response kinetics associated with lower temperatures and winter conditions [47,48].

Previously, estrogen receptor agonists have been shown to modulate respiratory burst responses in fishes [49]. ATR has also been shown to affect the production of ROS, as well as induce inflammation and glycometabolism in carp lymphocytes [50,51]. The respiratory response in the current study was significantly greater in fish from the E1+ATR mixture treatment relative to controls in April.

In previous studies, mitogenesis activity in both T- and B-cells have been shown to be dampened in response to E2 exposure [52]. This was also observed in the lymphocyte response to PHAP and LPS of EE2-treated fish in the current study; both responses were significantly reduced compared to the control group. Differences in mitogen responses may be the result of shifts in cell populations associated with treatments or effects on immune physiology; however, these measures were not evaluated here given the unavailability of immunophenotyping reagents. Exposure of juvenile LMB to EE2 has been shown to affect immunometabolism. This shift in metabolism involves increased glycolysis that could be advantageous for disease resistance in the short-term, but may lead to depleted glycogen stores in the long-term [52].

There was no significant effect of treatment on ACP activity; however, there were significant sex differences in the control and E1+ATR treated fish at the end of the study (April sampling, Fig. 3). This sex-specific difference did not occur in the EE2-treated fish, possibly due to a shift in the hemolytic activity in the males treated with EE2. Although alternative complement activity was not significantly different in the EE2-treatment group compared to the control group during the April collections, the lack of sex-specific differences in the EE2 treatment group is notable considering the exposure to EE2 was only during early recrudescence.

Fig. 3.

Alternative complement pathway hemolytic activity in largemouth bass collected at the April time point just prior to spawning (Females: Control n = 9, EE2 n = 8, E1 + ATR n = 6; Males: Control n = 9, EE2 n = 11, E1 + ATR n = 5). Mean (± SE) reciprocal of the plasma dilution causing 50 % lysis of rabbit red blood cells (ACH50). Error bars represent one standard error from the mean. Asterisk denotes significant difference between female and male within treatment group (p < 0.05).

The current state of knowledge does not allow for direct assessment of immunotoxic potential of contaminants to whole fish or populations due to the limited assay standardization, knowledge of assay performance to different toxicant groups, and reference data on the varied immune functions [53,54]. However, several endpoints commonly used in field assessments of LMB health were assessed in the current study [40], and these data can help fill data gaps in our knowledge of how representative EEDCs can modulate immune function under controlled conditions. Our results indicated that early recrudescence is a sensitive window of exposure to EEDCs and may lead to latent effects on LMB immune function. No significant treatment effects were observed just after the end of the dosing period in December, but several endpoints were altered in those previously exposed groups in April, four months post-dosing. Effects observed in the E1+ATR exposure group could be caused by the persistence of ATR, albeit at half the original dosed concentration through the post-dosing period. However, because EE2 quickly degraded in the study ponds, effects observed in the EE2 exposure group in April are considered latent effects. Latent effects of EE2 exposure during early recrudescence were also seen in male reproductive condition, including reduced gonadosomatic index and sperm motility [39]. In addition, EE2 exposure has also been shown to induce transcriptional effects on complement pathway genes in juvenile LMB and to modulate disease susceptibility [39].

In the present study, the effects of exposure to EE2 and the effects of exposure to the E1+ATR mixture were distinct. Fish in the EE2 treatment group exhibited lower mitogenic responses to PHAP and LPS relative to control fish, indicating effects on T-cell and B-cell responses, respectively, and fish in the EE2 treatment group did not exhibit significant differences in ACP activity between males and females that was observed in the control group. In contrast, fish in the E1+ATR mixture treatment group showed increased respiratory burst activity and significantly higher ACP activity in males compared to females. These differences in responses between the treatment groups are consistent with previously reported effects of ATR exposure on reproduction, without induction of biomarkers of estrogenic responses, in medaka (Oryzias latipes) and fathead minnow (Pimephales promelas) [55].

Conclusion

Results here demonstrate that environmentally relevant concentrations of endocrine-disrupting contaminants alter immune function during the spring just prior to spawning in an economically important fish species. These data provide species-specific baseline information for comparison to fish collected from impacted environmental settings. The present study indicates immunomodulation, but extrapolation beyond measured endpoints to disease-resistance and population-level effects is not possible from this study alone. Further investigation on effects of these contaminants on disease resistance in adult LMB would be warranted to gain a more complete assessment of potential impacts on fish health.

CRediT authorship contribution statement

Jessica K. Leet: Writing – review & editing, Writing – original draft, Visualization, Validation, Methodology, Formal analysis, Data curation, Conceptualization. Catherine A. Richter: Writing – review & editing, Visualization, Methodology, Conceptualization. Rachel A. Claunch: Writing – review & editing, Methodology, Investigation, Formal analysis, Data curation. Robert W. Gale: Writing – review & editing, Formal analysis, Data curation. Donald E. Tillitt: Writing – review & editing, Funding acquisition, Conceptualization. Luke R. Iwanowicz: Writing – review & editing, Methodology, Formal analysis, Data curation, Conceptualization.

Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Data availability

• Raw data can be found through the U.S. Geological Survey and is publicly available at https://doi.org/10.5066/P9U2U3A1, https://doi.org/10.5066/P9LDPEJF, https://doi.org/10.5066/P91G7KMO.