Dietary exposure to pesticides in poultry: From semen quality to embryonic mortality and tissue accumulation

Abstract

The widespread use of pesticides in agriculture increases the risk of chronic dietary exposure in poultry. This study investigated the effects of low-dose exposure to tebuconazole (TEB), imidacloprid (IMI), and glyphosate (GLP), administered individually or in combination at concentrations not exceeding maximum residue limits (MRLs), on male reproductive performance in Gallus gallus. Roosters were assigned to eight groups and exposed for six weeks (Phase I), followed by a four-week pesticide-free recovery period (Phase II).

Sub-MRL pesticide exposure impaired male reproductive function, with the most pronounced effects observed following combined treatments. During Phase I, exposure resulted in reduced semen quality, decreased fertility and hatchability, and increased embryo mortality, particularly in groups receiving IMI alone or in combination. These functional impairments were accompanied by detectable pesticide residues in reproductive tissues and body fluids, as well as modulation of local and systemic immune parameters.

During Phase II, semen parameters showed partial recovery toward control levels; however, fertility and hatchability remained reduced in several groups, coinciding with the persistence of pesticide residues in semen and selected tissues. Combined pesticide exposure consistently produced stronger and more persistent reproductive effects than individual compounds, indicating mixture-specific toxicity and incomplete reversibility.

Collectively, these findings demonstrate that chronic exposure to MRL-compliant pesticide doses can compromise avian reproductive performance, particularly under combined exposure scenarios. The persistence of residues in reproductive compartments and excreta further highlights potential environmental and biological risks, supporting the need to consider reproductive endpoints and chronic mixture exposure in pesticide risk assessment frameworks.

Introduction

In recent years, increasing attention has been given to the adverse effects of high pesticide doses on biological systems, particularly their impact on reproductive health in both mammals (Arıcan et al., 2020; Vasseur et al., 2024) and birds (Hussain et al., 2011; Fréville et al., 2024; Serra et al., 2023). The widespread use of pesticides, including fungicides, insecticides, and herbicides, for crop protection contributes to the contamination of animal feed, thereby exposing poultry to these substances. Numerous pesticides, including those approved for use in organic farming systems, as well as synthetic compounds, leave behind persistent residues (Vicini et al., 2019).

To address concerns about food and feed safety, the European Union has implemented strict regulations on pesticide usage. These include directives and regulations establishing maximum residue levels (MRLs) in cereals, vegetables, fruits, animal-derived products, and other foodstuffs (Directive 2002/32/EC, 2002) (European Parliament and Council, 2002). MRLs, set under Regulation (EC) No. 396/2005, are based on empirical data for each type of crop or animal product and define the legal limits for pesticide residues in food or feed when products are used according to the maximum recommended doses on the label. Complementing this food-safety framework, the EU REACH Regulation (EC) No. 1907/2006 requires the registration, evaluation, and, where necessary, authorization or restriction of chemical substances, including active ingredients and coformulants used in pesticides, to ensure that their broader risks to human health and the environment are thoroughly assessed. However, poultry feed, composed of diverse plant-based ingredients, may still contain measurable levels of pesticide residues.

Reports from the Regional Experimental Station of the Institute of Plant Protection – NRI in Białystok 2018, Regional Experimental Station of the Institute of Plant Protection – NRI in Białystok 2019 and national monitoring programs indicate an increasing frequency of pesticide residue detection in cereals intended for poultry feed, with many detections occurring within established MRLs. While some samples exceed these limits, a considerable proportion fall below regulatory thresholds, which may still exert biological effects under chronic exposure.

In the present study, we focused on three pesticides, tebuconazole, glyphosate and imidacloprid, which were selected on the basis of their extensive agricultural use, environmental persistence, and increasing evidence of adverse effects on reproductive health, even at low exposure levels.

Tebucanozole (TEB) is a triazole fungicide widely used in agriculture and industrial preservation. It inhibits sterol biosynthesis in fungi but is chemically stable and poorly biodegradable, making it a persistent contaminant in soil, water, and food (Bowen et al., 1997; Kahle et al., 2008). TEB is highly lipophilic, which facilitates its absorption into organic matrices and contact materials, potentially increasing its persistence and bioaccumulation (Kahle et al., 2008). Toxicological studies have shown that TEB disrupts endocrine function and leads to reproductive and developmental toxicity. In rats, perinatal exposure to TEB results in masculinization in females and feminization in males, including increased retention of nipple papillae (Taxvig et al., 2007). In vitro studies have demonstrated reduced levels of estradiol and testosterone and increased progesterone concentrations following TEB exposure (Kjærstad et al., 2010). Other observed outcomes include reduced epididymal weight, decreased sperm count and motility, and testicular dysfunction (Taxvig et al., 2007; Jacobsen et al., 2012).

Imidacloprid (IMI) is a neonicotinoid insecticide and chlorinated nicotine derivative that acts on nicotinic acetylcholine receptors (nAChRs) in insects (Hafez et al., 2016). Owing to its high water solubility and environmental persistence, IMI can accumulate in soil, water, and plants, posing exposure risks beyond its application areas (Bonmatin et al., 2005; Rogers et al., 2019). Despite being moderately lipophilic and capable of transient partitioning into fatty tissues or the lipid layer of skin, systemic uptake through intact skin is minimal. After oral exposure, it is rapidly and almost completely absorbed from the gastrointestinal tract and metabolized in the liver to 6-chloronicotinic acid, which is excreted mainly in urine and feces (EFSA, 2013). Although IMI is selective for insect receptors, studies have shown its negative effects on the mammalian reproductive system. It delays testicular development, disrupts spermatogenesis, reduces sperm quality, and alters ovarian morphology (Bal et al., 2012a). Rats exposed to IMI presented reduced sperm count, motility, and viability, as well as lower levels of sex hormones and severe testicular damage (Najafi et al., 2010; Soujanya et al., 2013; Bal et al. 2012b). Even doses below the no observed effect level (NOAEL) of 5–10 mg/kg caused long-term testicular dysfunction in rats during postnatal development and adulthood (Bal et al., 2012a; 2012b). Despite the ban on outdoor agricultural uses of IMI in the EU since 2018, the compound and its metabolites remain frequently detected in soil, surface water, and feed samples, underscoring its environmental persistence and the continued occurrence of limited uses or legacy contamination.

Glyphosate (GLP) is a nonselective herbicide that inhibits EPSP (5-enolpyruvylshikimate-3-phosphate) synthesis in plants (Franz et al., 1997) and is widely used because of its cost-effectiveness and broad-spectrum action against weeds (Coupe and Capel, 2016). It was approved as an active substance in the EU until 15 December 2033 and is subject to specific conditions and restrictions (EU Regulation 2023/2660). Although GLP is primarily hydrophilic, its formulations often contain surfactant adjuvants that confer partial lipophilicity, enhancing penetration into biological membranes and potentially altering absorption profiles. Despite its relatively low environmental persistence (Primost et al., 2017), repeated applications can still pose environmental risks (Mamy et al., 2010). Studies have shown that glyphosate may disrupt hormonal functions. It caused a 35 % reduction in testosterone levels in rats at low nontoxic concentrations of Roundup and glyphosate (1 ppm) (Clair et al., 2012). A decrease in sperm motility and mitochondrial activity in zebrafish at a concentration of 5–10 mg/mL was observed (Lopes et al., 2014). A concentration of 0.36 μg/mL was associated with reduced motility of human sperm (Anifandis et al., 2018). In rats, exposure to Roundup, whose main active ingredient is GLP, resulted in reduced sperm count and motility, as well as testicular damage (Owagboriaye et al., 2017). Low doses of this pesticide are recognized as endocrine disruptors in mammals (Gasnier et al., 2009). Moreover, exposure to GLP has been shown to induce alterations in testicular structure and hormone levels in ducks (de Liz Oliveira Cavalli et al., 2013).

Pesticides easily penetrate both the blood (Rudzi et al., 2022) and semen (Tan et al., 2016), compromising their quality (Mehrpour et al., 2014), as well as tissues, including testicular tissue, leading to dysfunction (Ham et al., 2023). Studies have shown that long-term pesticide exposure adversely affects multiple systems, with most reports focusing on its impact on the nervous system (Ghasemnejad-Berenji et al., 2021) and the reproductive system (Sharma et al., 2022), as well as its detrimental effects on embryonic development (Baldacci et al., 2018). Research on pesticide exposure through eggs has shown that these compounds increase embryonic mortality and reduce hatching rates (Gildersleeve et al., 1987) while simultaneously compromising semen quality and testicular function.

Although some studies have investigated the effects of low-dose pesticide exposure (Shi et al., 2011; Kaczyński et al., 2017), these studies typically involve short-term models, and most toxicological studies continue to focus on high-dose exposure. As a result, the potential risks associated with repeated exposure to low, environmentally relevant concentrations remain insufficiently understood.

Therefore, the aim of this study was to evaluate the effects of six weeks of dietary exposure to low, sublethal doses of tebuconazole, imidacloprid, and glyphosate—administered individually or in combination, corresponding to daily intakes of 0.02087 mg/kg BW/day (TEB), 0.00522 mg/kg BW/day (IMI), and 0.529 mg/kg BW/day (GLP) on reproductive performance in Gallus gallus roosters. Specifically, we assessed semen quality; fertilization and hatchability rates; embryonic mortality; and pesticide residues in the blood, semen, testicles, breast muscles, liver, and manure of roosters in vivo. In addition, we analyzed immune-related tissues (cecal tonsils and the spleen) to determine whether low-dose pesticide exposure modulates mucosal and systemic immune responses in poultry. These daily intake levels were derived from analytically confirmed pesticide concentrations in feed prepared at EU MRLs, reflecting environmentally realistic exposure scenarios.

Materials and methods

Chemicals

The substances used in the experiment were obtained from commercial suppliers. PNA Alexa Fluor 488 (a lectin derived from Arachis hypogaea) was purchased from Merck (St. Louis, MO, USA). Fluorescent markers, including propidium iodide (PI), JC-1, SYBR-14, Fluo-3 AM, C₁₁-BODIPY^581/591, acridine orange (AO), and M540 (Merocyanine), were obtained from Thermo Fisher Scientific (Waltham, MA, USA).

The pesticides used in the experiment were commercially available plant protection products: glyphosate (Roundup 360 Plus), tebuconazole (Orius Extra 250 EW), and imidacloprid (KOHINOR 200 SL).

For immunohistochemistry procedures, the following materials and instruments were used: a CryoStar NX50 cryostat (Epredia, Kalamazoo, MI, USA), background reducing component (Agilent, Santa Clara, CA, USA), monoclonal mouse antibodies (Southern Biotech, Birmingham, AL, USA), and Euparal mounting medium (Roth GmbH, Karlsruhe, Germany).

Preparation of the feed

The pesticide levels applied to the feed were not selected as predefined toxicological doses, but were derived from the MRLs under Regulation (EC) No. 396/2005 for cereals and feed ingredients. Feed mixtures were therefore prepared to contain residues corresponding to their respective MRL values. The resulting daily exposure (mg/kg body weight/day) was calculated post hoc based on the analytically confirmed feed concentrations, average daily feed intake, and body weight of the roosters.

Stock solutions were prepared by dissolving the commercial formulations in distilled water according to the recipe below. Immediately before application, 100 mL of each stock solution was diluted to 1.00 L with water (1:10, v/v) to obtain the working solution applied to the feed. The amounts of formulation used to prepare the stock solutions were as follows: Orius Extra 250 EW (tebuconazole, 250 g/L) 1.4 mL; Kohinor 200 SL (imidacloprid, 200 g/L) 0.42 mL; and Roundup 360 Plus (glyphosate, 360 g/L acid equivalent) 24.2 mL. Combinations: TEB+IMI: 1.4 mL + 0.42 mL; TEB+GLP: 1.4 mL + 24.2 mL; IMI+GLP: 0.42 mL + 24.2 mL; TEB+IMI+GLP: 1.4 mL + 0.42 mL + 24.2 mL.

A mechanical mixer was loaded with 75 kg of poultry feed. While mixing, the 1.00 L working solution was sprayed evenly with a handheld garden sprayer; mixing continued for 2 h to ensure homogeneity and minimize residual moisture. Five representative samples per treatment were collected 24 h after mixing and reanalyzed after 7 days to assess stability. Analyses were performed in an ISO 17025-accredited laboratory via LC‒MS with validated methods for GLP (Kaczyński et al., 2017), TEB (Hrynko et al., 2023) and IMI (Mojsak et al., 2018). Because small application losses (sprayer hold-up, deposition on mixer walls) are expected, analytically measured concentrations were taken as nominal doses. The mean concentrations at 24 h were 0.40 mg/kg TEB, 0.10 mg/kg IMI, and 10.15 mg/kg GLP, which remained stable in the combination groups after 7 days.

Animals

Roosters

The study utilized 80 twenty-one-week-old Green-legged Partridge roosters (Gallus gallus domesticus), each of which was housed individually in cages measuring 70 cm × 95 cm × 85 cm to ensure controlled monitoring of their behavior and physiological parameters. These roosters were reared from the first day of life under uniform conditions, including a controlled feeding regimen consisting of specially formulated feed tailored to their nutritional needs at each stage of growth and free of pesticides. During the study period, the birds were maintained at temperatures ranging from 18°C to 20°C, with a 14-hour light and 10-hour dark photoperiod to simulate optimal environmental conditions.

Hens

The laying hens, which were Hy-Line Brown at the age of 26 weeks, were placed in individual cages with a slated floor; the width and depth of the cages were 60 cm, and the height was 50 cm. The floor had a 6° slope, which allowed the eggs to slowly roll down into a trough, from which they were subsequently collected and transported to the egg storage area. The cage was equipped with a 60 cm feeder, two automatic nipple drinkers, a 60 cm perch, a designated nesting area, and a mat on which loose bedding material was scattered. The hens were fed a complete feed mixture purchased from a feed mill, which was free of pesticides. The rooms where the birds were kept were equipped with systems to maintain and adjust the required environmental parameters: temperature of 16–19°C, air humidity of 60–70 %, and daylight duration of 16 hours.

Experimental design

The roosters were divided into 8 groups, each consisting of 10 roosters (Fig. 1). They were fed specially prepared feed. All the feed was prepared to contain pesticide residues at levels corresponding to EU MRL-based targets, and the final concentrations were confirmed analytically (see the Preparation of the Feed section) and then administered as a feed mixture. Group 1 served as the control without pesticide residues in their diet. Group 2 received feed with the addition of TEB, group 3 with the addition of IMI, group 4 with the addition of GLP, group 5 with the addition of TEB + IMI, group 6 with the addition of TEB + GLP, group 7 with the addition of IMI + GLP, and group 8 with the addition of TEB + IMI + GLP.

Fig. 1.

Overview of the experimental design and collected material. Roosters were randomly divided into eight treatment groups (n = 10), including control, single pesticide (TEB, IMI, GLP), and their combinations. During phase I (weeks 1–6), birds received pesticide-supplemented feed, and semen was collected twice per week. Two artificial inseminations were performed. Following a four-week washout period (weeks 7–10), confirmed through feed residue analysis, phase II (weeks 11–13) included three semen collections and inseminations to evaluate recovery. Biological samples (semen, serum, organs, and manure) were collected after both phases for analyses of fertility, immune markers, and embryo development.

The daily intake of pesticides in the feed of the roosters was as follows: TEB at 0.02087 mg/kg of body weight per day, IMI at 0.00522 mg/kg of body weight per day and GLP at 0.529 mg/kg of body weight per day. Water was available ad libitum. Average daily feed intake and body weight were recorded throughout phase I and used to calculate the daily pesticide intake (mg/kg BW/day).

During the six-week experimental period (phase I), the birds received their assigned diets, and semen was collected twice weekly. Collection started in the first week of pesticide-containing feed administration and continued through week six. All roosters subsequently underwent a four-week washout period on a pesticide-free control diet. In phase II (weeks 11–13), semen samples were collected three times per group to determine whether previously observed effects on semen quality persisted or had further evolved.

To assess the actual fertility of the roosters and the fertilization capability of their sperm in vivo, 40 hens (5 females per group) were inseminated with semen collected from each study group. The hens were inseminated four times: twice during weeks 5 and 6 and again during weeks 12 and 13 of the experiment. This approach aimed to determine whether any changes in rooster fertility were reversible and whether reproductive performance improved after a 4-week pesticide-free diet.

After 6 and 13 weeks of the experiment, five males from each group were euthanized, and samples of blood, tissues (testes, liver, and breast muscle), and manure were collected to assess pesticide residue levels. Moreover, after 6 weeks, the spleen and cecal tonsils were collected for immunohistochemical evaluation. Immunological analyses were restricted to phase I, corresponding to the period of active dietary pesticide exposure, during which potential immunomodulatory effects were expected. Phase II was designed to assess the reversibility of reproductive parameters following a pesticide-free withdrawal period; therefore, immune endpoints were not repeated.

Ethical approval for the study was obtained from the Local Animal Welfare Committee in Wroclaw (Approval number: 063/2022/P1).

Semen collection

For semen collection, the same individuals were subjected to the dorso-abdominal massage method under consistent conditions (Burrows and Quinn, 1937). All the collected ejaculates were deemed clean and suitable for analysis. The number of semen collections in phase I was 12 (6 × twice a week), whereas in phase II was 3.

Assessment of sperm quality

Comprehensive semen analysis was carried out after collection; sperm motility parameters were evaluated with the CASA system, and flow cytometry was used to assess membrane integrity, mitochondrial activity, acrosome integrity, chromatin structure, the intracellular calcium level and lipid peroxidation.

The assessment of sperm motility parameters was conducted via computer-assisted semen analysis (CASA Ceros II, Hamilton Thorne). The analysis was conducted at 37°C, with 5 fields evaluated per sample and a minimum of 200 sperm cells analyzed. The system settings were as follows: frame rate of 60 Hz, recording time of 1.0–1.5 s per field, and cell detection thresholds corresponding to a sperm head area of 5–15 µm² and pixel intensity of 70–75. Semen samples were diluted 50 times in Dulbecco’s modified medium with low glucose (DMEM) to achieve a sperm concentration of approximately 50 × 10⁶/mL. A 3 μL portion of the diluted semen was placed in an analysis chamber (Leja, Nieuw-Vennep, Netherlands) maintained at 37°C, and five random fields were examined for each sample. The parameters assessed included the percentage of motile sperm (MOT), the percentage of progressively motile spermatozoa (PROG), path velocity (VAP), progressive velocity (VSL), and curvilinear line velocity (VCL).

A Guava EasyCyte 5 Merck Millipore flow cytometer was used to assess various sperm characteristics. A 488 nm argon ion laser was used to excite the fluorescent probes, and data acquisition was performed with GuavaSoft 3.1.1 software. Each sample was analyzed for 5000 events, with nonsperm events excluded on the basis of scatter properties.

The plasma membrane integrity was evaluated via the Live/Dead Sperm Viability Kit, with staining performed via SYBR-14 (at a final concentration of 333 nM) and propidium iodide (PI, 40 μM), following the protocol described by (Partyka et al., 2010). A 300 µL aliquot of the diluted suspension was transferred into cytometric tubes, followed by the addition of 5 µL of the SYBR-14 working solution. The working solution was prepared by diluting the stock SYBR-14 reagent with distilled water at a 1:49 ratio. The samples were first stained with SYBR-14 and incubated in the dark at room temperature for 10 minutes. Three minutes before analysis, the cells were counterstained with PI. Positive SYBR-14 staining (green fluorescence) and negative PI staining were identified as having intact plasma membranes (PMI).

The mitochondrial activity in the sperm was evaluated via JC-1 and PI staining. A 3 mM JC-1 stock solution was prepared. For each sample, 0.67 µL of JC-1 was added to 500 µL of semen diluted in BPSE. The samples were incubated in darkness at 37°C for 20 minutes. Subsequently, 3 µL of PI was added, and samples were incubated for an additional 2–3 min prior to analysis. Sperm showing orange fluorescence were classified as having high mitochondrial membrane potential (HMMP), whereas those displaying both green and orange fluorescence were considered to have moderate mitochondrial activity. The sperm stained with PI were identified as nonviable cells.

Acrosome integrity was evaluated via conjugation with the lectin PNA Alexa Fluor 488. The sperm samples were incubated with a PNA working solution (1 μg/mL) in darkness at room temperature for 5 minutes, followed by the addition of PI prior to analysis. PNA staining allowed for the identification of cells with altered beta-galactosyl residues, whereas PI enabled differentiation between viable and nonviable cells, aiding in the assessment of acrosomal integrity. This dual-marker method distinguished specific subpopulations, including live cells with intact acrosomes (PI-PNA-), live cells with compromised acrosomes (PI-PNA+), dead cells with intact acrosomes (PI+PNA-), and dead cells with damaged acrosomes (PI+PNA+).

Chromatin status was assessed via acridine orange (AO) staining, where green fluorescence indicated normal double-stranded DNA, and increased red fluorescence signified denatured DNA (DFI). Semen samples were diluted with BPSE to a final concentration of 1 × 10⁶ spermatozoa per mL. Then, 100 µL of this suspension was mixed with 200 µL of lysis solution (0.1 % Triton X-100, 0.15 M NaCl, 0.08 M HCl, pH 1.4) for brief acid denaturation, which was maintained for 30 seconds. This mixture was then combined with 600 µL of acridine orange solution (6 µg AO/mL in a buffer containing 0.1 M citric acid, 0.2 M Na₂HPO₄, 1 mM EDTA, and 0.15 M NaCl, pH 6). After a 3-minute incubation, the samples were introduced into a flow cytometer for analysis.

Calcium ion levels were detected via Fluo 3 AM dye. The procedure involved incubating diluted rooster sperm in BPSE with Fluo 3 AM at a final concentration of 0.5 mM for 15 minutes at 37°C. After incubation, the sample was centrifuged, and the supernatant was discarded. Then PI was added to distinguish between live and dead cells, allowing for the analysis of cell populations with varying calcium ion levels via a flow cytometer (Harrison et al., 1993).

Lipid peroxidation (LPO) in the sperm plasma membrane was assessed via the fluorescent lipid probe C₁₁-BODIPY^581/591, which contains two double bonds that are oxidized upon reaction with reactive oxygen species (ROS), resulting in a color change (red when unoxidized and green when oxidized). Semen samples were diluted in BPSE (500 µL) and mixed with 1 µL of 2 mM C₁₁-BODIPY^581/591 in ethanol, followed by a 30-minute incubation at 37°C in darkness, as described previously (Partyka et al., 2011). After incubation, the samples were centrifuged at 500 × g for 3 minutes, and the sperm pellet was resuspended in 500 µL of BPSE. For viability assessment, sperm were stained with PI and incubated for 5 minutes at room temperature before cytometric analysis.

Insemination of laying hens

Hens from all groups were inseminated with a dose of 200 × 10⁶ spermatozoa. Within 50 minutes after collection, semen was deposited intravaginally via plastic pipettes (IMV, U 212, L'aigle, France) and an inseminating pistol (IMV, U 695, L'aigle, France). In each group, four inseminations were performed.

Egg incubation

The eggs from each group were collected daily, placed weekly in a JARTOM incubator (Jartom, PL) and incubated according to the instructions for chicken eggs (temperature 37.7°C, humidity 57 %, egg turning every hour). Fecundity results (fertility, hatchability of the set and fertile eggs and embryo mortality) were determined for the eggs collected from Day 2 after the first insemination to day 6 after the last insemination. Fertility was estimated on the basis of egg candling on day 7 of incubation, all fertile eggs were incubated until hatching, and hatchability was determined. Infertile eggs and those with dead embryos were broken and evaluated macroscopically to determine the day of death. Embryo mortality was assessed as total embryo mortality, without differentiation into early or late embryonic death. All non-hatched eggs containing non-viable embryos detected at the end of incubation were included together in the embryo mortality parameter.

Euthanasia

Males from each group were euthanized through cervical dislocation, preceded by the administration of analgesics and sedatives via intramuscular injection (butorphanol at 0.1 mg/kg body weight and medetomidine at 0.1 mg/kg body weight). The death of the animals after euthanasia was confirmed through clinical examination and observation. The characteristics confirming death included the cessation of heart and respiratory activity, absence of reflexes, and a decrease in body temperature.

Assessment of changes in the spleen and cecal tonsils

Sample collection and processing

Sections of the spleen and cecal tonsils (CT) were collected, fixed in 4 % phosphate-buffered paraformaldehyde (pH 7.4) for 1 h, washed in 0.1 M phosphate buffer and infiltrated with buffered 30 % sucrose. The samples were then frozen via a cryostat, cut into 10 μm serial sections, air-dried overnight and kept frozen until further immunohistochemical staining.

Immunohistochemical staining

After endogenous peroxidase activity was quenched with a 3 % hydrogen peroxide solution, the cryosections were preincubated with antibody diluent with background reducing component for 20 min to prevent nonspecific binding. Monoclonal mouse antibodies targeting chicken antigens, namely, Bu-1 (also known as the chicken B-cell marker chB6) (clone AV20, 1:500), CD4 (clone CT-4, 1:200), and CD8α (clone CT-8, 1:200), were then applied to the serial sections, with PBS used as a control. The slides were incubated for 1 h at room temperature. Antibody visualization was achieved via EnVision FLEX+ (Dako/Agilent) with 3,3ʹ-diaminobenzidine chromogen (DAB) according to the manufacturer’s instructions. Finally, the sections were dehydrated in a series of alcohols and embedded in Euparal (Roth GmbH, Karlsruhe, Germany).

Histometric analysis

For the spleen and CT images, the percentages of Bu-1+, CD4+ and CD8+ cells were estimated on immunohistochemically stained slides that were examined and photographed under a Nikon Eclipse Ni light microscope (Nikon, Melville, NY, USA) equipped with a video camera. The brown-colored area occupied by antigen-positive cells was estimated on a 0.344 mm² area (200 × magnification) via the NIS-Elements BR 5.41 program and expressed as a percentage of the field of view. In the spleen, the CD4+/CD8+ cell ratio was calculated for each individual, followed by estimation of the average for the group. Moreover, in the spleen, the area of the germinal centers was calculated for a 1.376 mm² area (100 × magnification) and expressed as a percentage of the field of view (Madej and Bednarczyk, 2016). On CT, the fields of view were consistently selected starting from the lamina propria mucosae and extending toward the lumen of the organ (Madej et al., 2020). In each case, a histologist ensured the elimination of all artifacts.

Methods of pesticide residue analysis

The studies of TEB, IMI, GLP content in blood serum, seminal plasma, breast muscle, liver and testicular tissues were performed in a laboratory certified in accordance with ISO 17025. The analysis was carried out via validated methods involving the use of liquid chromatography coupled with mass spectrometry. A detailed description of the methods used and their validation is included in the Supplementary materials (Additional file 1).

Statistical analysis

The following statistical methods were applied in the analysis. For data expressed as percentages, assumptions of normality of residuals were first verified via the Shapiro–Wilk test, and homogeneity of variances was assessed via Levene’s test, both of which were performed on transformed data. If these assumptions were met, ANOVA was conducted on the transformed data, followed by Tukey’s post hoc test for pairwise comparisons. For data not expressed as percentages, only the normality of residuals was tested via the Shapiro–Wilk test. If the normality assumption was not satisfied, we performed the Kruskal–Wallis test, followed by Dunn’s post hoc test for detailed group comparisons.

In addition, fertility rates, hatchability, and embryo mortality rates were analyzed as binomial data (success/failure). Overall differences among experimental groups were assessed via the chi-square test, and pairwise group comparisons were performed via Fisher’s exact test without multiple comparison correction. Differences between the experimental phases (Phase I and Phase II) within each treatment group were also evaluated via Fisher’s exact test.

The data were analyzed via Statistica 13.3 software (StatSoft Polska Sp. z o.o., Cracow, Poland) and GraphPad Prism version 10.4.0 (GraphPad Software, San Diego, CA, USA). Differences were considered statistically significant at P < 0.05 and highly significant at P < 0.01. The results are presented as the means ± SEMs.

Results

Effects of pesticides on semen characteristics (phase I and II)

Sperm motility

Feeding roosters pesticide-supplemented diets for six weeks (Phase I) altered several motility parameters (Fig. 2A–E). The IMI+GLP combination produced the strongest negative effect, significantly reducing progressive motility (P < 0.01, Fig. 2A), while overall motility remained unchanged across groups (Fig. 2B). Velocity parameters (VSL, VCL, VAP) varied depending on the pesticide mixture: the TEB+IMI+GLP and TEB+GLP groups showed significantly increased VSL, VCL, and VAP values compared with IMI+GLP (P < 0.01, Fig. 2C–E).

Fig. 2.

The effects of different combinations of pesticides on motility parameters of chicken spermatozoa, including the percentage of progressively motile spermatozoa (PROG) - A, the percentage of motile sperm (MOT) - B, the progressive velocity (VSL) - C, the curvilinear line velocity (VCL) - D, and the path velocity (VAP) - E in phases I and II. The data are expressed as the means ± SEM.

Different letters above the bars indicate significant differences AB (P < 0.01) between each treatment group and the control within the same phase. Asterisks indicate significant differences between the same treatment group in phase I vs. phase II (*P < 0.05, **P < 0.01, ***P < 0.001).

After withdrawal of pesticides (Phase II), velocity parameters generally returned to control levels. Cessation of IMI+GLP exposure improved progressive motility and restored VSL, VCL, and VAP (P < 0.01), while TEB withdrawal enhanced VCL and VAP (Fig. 2D–E). Total motility remained unaffected (Fig. 2B).

Flow cytometry parameters

In Phase I, IMI reduced sperm membrane integrity (P < 0.05) (Fig. 3A) and lowered HMMP compared with TEB and TEB-containing mixtures (P < 0.05) (Fig. 3B). The spermatozoa from the control group presented a significantly (P < 0.01) greater percentage of live cells with acrosome damage compared to sperm exposed to pesticides, including TEB, GLP, and their combination with IMI (Fig. 3C). IMI resulted in the highest (P < 0.01) DFI values, indicating its particularly harmful effect on DNA integrity. In contrast, TEB+IMI+GLP significantly decreased (P < 0.01) DNA fragmentation in chicken spermatozoa (Fig. 3D). Pesticides had no effect on intracellular calcium levels (both high- and low-calcium conditions; Fig. 3E–F) or lipid peroxidation (Fig. 3G). No significant differences in any of the evaluated parameters between the groups were observed in phase II. The cessation of pesticides decreased sperm membrane integrity in the TEB+IMI+GLP group (Fig. 3A).

Fig. 3.

The effects of different combinations of pesticides on flow cytometry parameters: plasma membrane integrity (PMI) - A, high mitochondrial potential (HMMP) - B, live cells with damaged acrosome – C, DNA fragmentation index (DFI) – D, high intercellular Ca²⁺ in live cells – E, high intercellular Ca²⁺ in live cells – F, live cells with LPO - G in phases I and II. The data are expressed as the means ± SEM.

Different letters above the bars indicate significant differences AB (P < 0.01) between each treatment group and the control within the same phase. Asterisks indicate significant differences between the same treatment group in phase I vs. phase II (**P < 0.01).

Fertility, hatchability, and embryo mortality

Six-week exposure to TEB, IMI, TEB+GLP, IMI+GLP, and TEB+IMI+GLP pesticides resulted in a significant reduction in fertility and hatchability compared with the control group (P < 0.01; Table 1). In addition, exposure to IMI alone led to a significant increase in embryonic mortality during phase I (P < 0.01).

Table 1.

Effects of different pesticide combinations on fertility parameters, including the fertility rate, hatchability, and embryo mortality, in phases I and II.

	Phase I			Phase II
Para- meters Groups	Fertility rate [%]	Hatchability of incubated eggs [%]	Embryo mortality [%]	Fertility rate [%]	Hatchability of incubated eggs [%]	Embryo mortality [%]
CTR	91.9 ± 5.4 ᴬ	78.2 ± 8.5 ᴬ	15.0 ± 5.5 ᴬ	98.5 ± 1.7 ᴬ	84.6 ± 10.5 ᴬ	14.1 ± 9.2
TEB	74.2 ± 8.2 ᴮ*	59.6 ± 10.5 ᴮ*	22.1 ± 3.3 ᴬᴮ	96.4 ± 3.1 ᴬᴮ*	89.3 ± 2.1 ᴬᴮ*	7.4 ± 0.8
IMI	64.6 ± 5.6 ᴮ*	46.3 ± 9.0 ᴮ*	28.3 ± 8.6 ᴮ*	82.0 ± 0.9 ᴮ*	76.0 ± 0.2 ᴬᴮ*	7.3 ± 1.2 *
GLP	92.5 ± 1.8 ᴬᴮ	78.8 ± 1.4 ᴬᴮ	15.1 ± 3.0 ᴬᴮ	79.0 ± 7.7 ᴮ	69.4 ± 8.9 ᴮ	11.5 ± 1.8
TEB + IMI	94.6 ± 4.8 ᴬᴮ*	77.0 ± 8.5 ᴬᴮ	18.6 ± 6.8 ᴬᴮ	73.0 ± 18.6 ᴮ*	63.5 ± 17.6 ᴮ	13.0 ± 2.5
TEB + GLP	77.9 ± 8.2 ᴮ	60.3 ± 10.6 ᴮ	23.5 ± 5.5 ᴬᴮ	83.0 ± 9.3 ᴮ	66.0 ± 9.6 ᴮ	20.5 ± 2.5
IMI + GLP	58.6 ± 4.5 ᴮ*	46.0 ± 9.2 ᴮ*	21.6 ± 11.0 ᴬᴮ	86.0 ± 3.3 ᴮ*	71.9 ± 6.5 ᴬᴮ*	16.3 ± 4.4
TEB+IMI+GLP	57.0 ± 4.3 ᴮ	44.2 ± 4.1 ᴮ	22.5 ± 3.6 ᴬᴮ	66.7 ± 7.8 ᴮ	56.7 ± 1.4 ᴮ	15.0 ± 7.9

Abbreviations: CTR – control; TEB – tebuconazole; IMI – imidacloprid; GLP – glyphosate; Phase I – exposure period; Phase II – post-exposure period. Different superscript letters within the same column and phase indicate statistically significant differences between groups based on pairwise comparisons using Fisher’s exact test (P < 0.01). Groups sharing at least one common letter do not differ significantly. Asterisks (*) indicate statistically significant differences between Phase I and Phase II within the same treatment group (Fisher’s exact test, P < 0.05). In Phase I, the number of eggs (n) per group ranged from approximately 68–89, and in Phase II from approximately 60–93.

In phase II, most pesticides, except of TEB, caused a significant decrease (P < 0.01) in the fertility rate. GLP, TEB+IMI, TEB+GLP and TEB+IMI+GLP significantly reduced (P < 0.01) hatchability. No statistically significant differences in embryo mortality in phase II were observed.

The cessation of pesticides revealed that the fertility rate significantly increased (P < 0.05) in the TEB, IMI, and IMI+GLP groups; however, in the TEB+IMI group, it significantly decreased (P < 0.05). Similarly, hatchability increased significantly (P < 0.05) in phase II in the TEB, IMI and IMI+GLP groups. Moreover, embryonic mortality decreased in phase II, in the IMI group (p < 0.05).

Pesticide levels in tissues, body fluids and manure

TEB

During phase I, TEB accumulated significantly in all examined matrices compared with the control group (P < 0.01) (Fig. 4A–F). Among solid tissues, TEB concentrations followed a clear descending pattern, with the highest levels detected in manure, followed by liver, breast muscle, and testes.

Fig. 4.

The level of TEB in the testes (A), breast muscle (B), liver (C), blood serum (D), semen (E) and manure (F) in phases I and II. Different letters above the bars indicate significant differences AB (P < 0.01) between each treatment group and the control within the same phase. Asterisks indicate significant differences between the same treatment group in phase I vs. phase II (*P < 0.05, **P < 0.01, ***P < 0.001).

Following pesticide withdrawal (phase II), TEB concentrations declined markedly and were close to or below the detection limit in most analyzed tissues and body fluids. A significant reduction was observed across all groups (P < 0.05–0.001). Detectable residues persisted only in semen, particularly in the TEB+IMI group, where concentrations remained comparable to those observed in phase I.

IMI

In phase I, IMI exposure resulted in significant accumulation of this pesticide in the testes, liver, blood serum, semen, and manure (P < 0.01) (Fig. 5A–F), while concentrations in breast muscle remained below the detection limit. Moderate levels were detected in the liver, whereas the highest accumulation occurred in manure, especially when IMI was administered in combination with other pesticides.

Fig. 5.

The level of IMI in the testes (A), breast muscle (B), liver (C), blood serum (D), semen (E) and manure (F) in phases I and II. Different letters above the bars indicate significant differences AB (P < 0.01) between each treatment group and the control within the same phase. Asterisks indicate significant differences between the same treatment group in phase I vs. phase II (*P < 0.05, **P < 0.01, ***P < 0.001).

In phase II, IMI residues decreased substantially across all analyzed materials. Following the withdrawal period, a significant reduction in IMI concentrations was observed in the testes, liver, blood serum, and manure in all IMI-exposed groups, with levels remaining low or near the detection limit.

GLP

During phase I, GLP accumulated significantly in all analyzed tissues, body fluids, and manure compared with the control group (P < 0.01) (Fig. 6A–F). The highest concentrations were observed in manure and liver, whereas lower levels were detected in breast muscle. In semen, the greatest GLP accumulation occurred in the GLP-only group.

Fig. 6.

The levels of GLP in the testes (A), breast muscle (B), liver (C), blood serum (D), semen (E) and manure (F) in phases I and II. Different letters above the bars indicate significant differences AB (P < 0.01) between each treatment group and the control within the same phase. Asterisks indicate significant differences between the same treatment group in phase I vs. phase II (*P < 0.05, **P < 0.01, ***P < 0.001).

After pesticide withdrawal (Phase II), GLP concentrations declined in most tissues; however, significant residues persisted in the testes, breast muscle, and manure. In particular, in the semen the GLP-only group showed significantly higher GLP levels compared with the control (P < 0.01). Overall, although a significant decrease was noted following the break, GLP remained detectable in selected matrices, especially manure.

Immune cell profiles

Immunohistochemical analysis of the cecal tonsil showed a significantly higher proportion of CD8⁺ cells in the IMI+GLP group compared with the control (P < 0.05; Fig. 7A, Fig. 8). In the spleen, neither the immune cell populations nor the CD4⁺/CD8⁺ ratio differed among groups (Fig. 7B–C). The only alteration observed in the spleen was a significant increase in germinal center area in the TEB+GLP group relative to the control (P < 0.05; Fig. 7D).

Fig. 7.

Impact of pesticides on the immune response in the cecal tonsils and spleen of chickens in phase I (mean ± SEM). The area occupied by antigen-positive cells in the cecal tonsils (A) and spleen (B) of chickens from the control group and groups receiving pesticides. The CD4+/CD8+ cell ratio (C) and germinal center area (D) in the spleen of chickens from the control group and groups receiving pesticides. Different letters above the bars indicate significant differences ab (P < 0.05) between each treatment group and the control.

Fig. 8.

Distribution of Bu-1+ (A), CD4+ (B), and CD8+ (C) cells in the cecal tonsils of chickens in the control group. A: The germinal centers (arrowheads) are clearly visible. C: CD8+ cells are mainly distributed in the subepithelial areas (arrowheads). D: increase in CD8+ cell number in the group receiving imidacloprid and glyphosate (IMI+GLP). Immunohistochemical staining. Scalebar = 100 μm.

Discussion

Although all pesticide treatments in our study remained within the maximum residue limits (MRLs) set by current EU legislation, clear adverse effects on key reproductive endpoints were observed. Our findings demonstrate that chronic exposure to sub-MRL levels of pesticides can compromise avian reproductive performance, as reflected by impaired sperm functionality, reduced fertilization success, altered embryo development, and decreased fertility and hatchability. Importantly, these reproductive impairments were accompanied by measurable pesticide residues in tissues and body fluids, along with alterations in peripheral immune organs. Notably, pesticide concentrations within the MRLs were sufficient to alter sperm motility parameters; however, some of these effects diminished following the discontinuation of pesticide administration, suggesting at least partial reversibility. Together, these observations indicate that current regulatory thresholds may not fully account for the biological consequences of long-term or combined pesticide exposure.

Effects of individual pesticides on reproductive traits

The individual pesticides are discussed below according to the strength and biological relevance of their effects on reproductive traits. In our study, TEB, IMI, and GLP as individual substances, did not affect sperm kinematic parameters, membrane integrity, nor lipid peroxidation. The results for plasma membrane integrity are consistent with findings regarding bovine sperm viability (Kabakci et al., 2021), rooster sperm in an in vitro experiment (Napierkowska et al., 2024), and glyphosate effects in zebrafish under in vivo conditions (Lopes et al., 2014). An intriguing phenomenon we observed was that acrosome damage in rooster spermatozoa was lower in almost all pesticide-treated groups. This may be related to alterations in sperm calcium homeostasis induced by pesticides, as demonstrated in a feeding experiment with GLP (Serra et al., 2021). The literature also reports that neonicotinoids can affect sperm ion channels, including calcium channels regulating hyperactivation (Xu et al., 2023). Although no significant differences in Ca²⁺ concentration mentioned authors detected, the downward trend observed may indicate subtle disturbances in calcium homeostasis. However, in our study, we observed only a trend toward decreased calcium levels in rooster sperm following pesticide exposure, but the differences were not statistically significant.

Our study revealed that IMI was the only pesticide that did not affect acrosome structure, similar to the findings of an in vitro experiment (Napierkowska et al., 2024), and to what was observed in sperm from mice fed atrazine at doses of 0.1 and 1 mg/kg (Saalfeld et al., 2018). In vivo studies in mice have demonstrated that triazoles, such as TEB, induce mitochondrial dysfunction in female germ cells, resulting in ROS accumulation, apoptosis, and oocyte DNA damage (Zhang et al., 2019). Whereas in our study, within individual pesticides, IMI decreased the mitochondrial potential of the chicken spermatozoa. It seems that these pesticide types may also exert genotoxic effects. This hypothesis is further supported by one of the parameters we examined related to chromatin structure. The insecticide IMI negatively affected sperm DNA, increasing chromatin damage in rooster sperm. This effect likely contributed to the fertilization process and embryo development in the exposed group, which will be discussed in detail in the following section. Similar observations have been made in previous studies, where other organophosphorus insecticides had the most detrimental effect on DNA structure (Piña-Guzmán et al., 2006). In our case, this did not lead to increased lipid peroxidation, as has been reported in zebrafish exposed to herbicides (Mesmar et al., 2024) and in mice exposed to insecticides (Piña-Guzmán et al., 2006). The proposed mechanisms include increased ROS production, impaired mitochondrial function, and interference with DNA repair. Among all tested pesticides, exposure to IMI resulted in a decreased fertility rate, reduced hatchability, and, notably, the highest embryo mortality. This outcome may be largely associated with the elevated DNA fragmentation index observed in spermatozoa. Our findings are consistent with earlier studies demonstrating that IMI can markedly disrupt avian development (Kammon et al., 2012; Balani, et al., 2011). In the present study, the administered IMI dose was 0.00522 mg/kg body weight/day, whereas a substantially higher dose of 5 mg/kg body weight was applied in a previous report (Kammon et al., 2012); nevertheless, developmental impairments in chicks were observed in both cases. Furthermore, IMI exposure has been reported to significantly decrease hatchability and survival during early developmental stages in fish, such as Clarias gariepinus, and to induce morphological abnormalities in embryos and larvae (Erhunmwunse et al., 2023). Similarly, another insecticide, sumithion, has been reported to reduce hatchability and increase embryo mortality in zebrafish, causing various deformities in embryos (Rahman et al., 2020). Our findings of increased embryo mortality after IMI exposure are also supported by other studies (Gao et al., 2016).

TEB, as triazole fungicide, also significantly affected chicken reproduction, particularly during phase I of our experiment. In house sparrows (Passer domesticus), exposure to TEB results in its presence in eggs, which can disrupt embryonic development and reduce reproductive success (Bellot et al., 2024). This observation aligns with the marked lipophilicity of the compound (Płatkiewicz et al., 2025), thereby favoring deposition in lipid-rich yolk and vitelline membranes and prolonging embryonic exposure. Tebuconazole has been shown to disrupt reproductive function in zebrafish by altering hormone levels and gene expression in the hypothalamic–pituitary–gonadal (HPG) axis, leading to reduced egg production and fertilization success (Yan et al., 2023). Consistently, in our experiment on roosters, prolonged exposure to TEB resulted in a decline in fertility and hatchability rates, highlighting its negative impact on male reproductive performance. A similar trend was observed in other studies, where it was demonstrated that exposure to fungicides in grey partridges can decrease fertility, hatchability, and chick survival (Bro et al., 2015). Additional evidence of the negative impact of TEB on reproductive capacity and hatchability in birds has been provided by research conducted on other avian species, reinforcing concerns about its broader effects (Bellot et al., 2022; Lopez-Antia et al., 2021 ). Importantly, our TEB dose was 0.02087 mg/kg body weight/day, whereas in one of the referenced studies, it was 0.165 mg/kg bw/day (Bellot et al., 2022). In both cases, a reduction in the hatching rate was observed. Similar findings were also reported at doses of 0.2 and 1.1 mg/kg body weight/day, where the hatching rate decreased by 23 %, and the brood size was 1.5 times smaller in the high-dose group than in the control (Lopez-Antia et al., 2021).

The prolonged exposure to GLP in our study revealed a negative effect only during the second phase of the experiment, which may indicate a delayed action of GLP or its ability to accumulate in tissues and be gradually released. Long-term exposure to GLP in broiler chickens led to a reversible (after two weeks) increase in early embryonic mortality and delays in the development of surviving embryos (Estienne et al., 2022). Even at a GLP dose below the NOAEL, this pesticide accumulates in the egg yolk, causing significant embryo mortality at an early stage of development (Estienne et al., 2022). Notably, the GLP dose in our study was 0.529 mg/kg body weight/day, whereas in the cited study, it was significantly greater at 47 mg/kg body weight/day. Although GLP alone did not significantly affect fertility in phase I of our experiment, previous findings in poultry suggest that even low levels of glyphosate residues in feed can reduce egg hatchability (Foldager et al., 2021).

Combined pesticide exposure and interaction effects

Negative effects on sperm traits became apparent only when pesticides were applied in combination. IMI combined with GLP, even at permissible levels, reduced motility parameters similarly to effects reported for insecticides in feeding experiments on rats (Perobelli et al., 2010) and in the in vitro study, where GLP significantly reduced motility parameters in men (Nerozzi et al., 2020). Disturbances in progressive motility may result from pesticide interference with spermatogenesis, sperm transport through the reproductive duct, or mitochondrial function after ejaculation. IMI and GLP are also well known to disrupt mitochondrial activity by inhibiting the respiratory chain and inducing ROS production (Mesnage et al., 2015; Karami-Mohajeri and Abdollahi, 2013), which may explain the decline in mitochondrial membrane potential observed in this group. Moreover, it should be emphasized that after the interruption in pesticide exposure, the group receiving the combination of all three pesticides showed a decrease in sperm membrane integrity, indicating a persistent synergistic effect of their combined action. Therefore, our study broadens the toxicological perspective by including other pesticide classes, specifically insecticides and herbicides, and highlights their comparable potential to impair male reproductive function.

The strongest reproductive impairments in fertility and hatchability rates were observed in the IMI+GLP and TEB+IMI+GLP groups. Synergistic toxicity of these mixtures has previously been reported in invertebrates (Man et al., 2023) and other animal models, suggesting that even sub-MRL doses may interact to potentiate the effects of individual compounds. In phase II, the combined pesticide-treated groups also continued to exhibit reduced fertility and hatchability, although some parameters improved relative to those in phase I. This finding suggests that while certain effects may be partially reversible after exposure ceases, the negative impact on reproductive outcomes persists in several groups, particularly those exposed to combined pesticides. In the postexposure phase, reduced fertility and hatchability were particularly evident in the group exposed to the mixture of all three pesticides.

Pesticide residues and persistence in reproductive tissues

In our study, IMI, TEB, and GLP effectively penetrated body fluids, tissues and organs. The presence of these pesticides in manure provides clear evidence of their systemic absorption followed by fecal elimination. In particular, the high levels of GLP detected in rooster feces confirm concerns regarding the unintentional introduction of glyphosate-based herbicide (GBH) residues into the environment through the use of organic fertilizers such as poultry manure (Muola et al., 2021), highlighting manure as a relevant environmental exposure pathway.

After oral administration, TEB is rapidly distributed throughout the body, and its metabolites are primarily excreted in urine (Oerlemans et al., 2019). In our experiment, the highest concentrations of TEB in blood serum were observed in the groups coexposed to IMI, both with and without GLP, suggesting that IMI may delay TEB elimination from the bloodstream. TEB showed strong affinity for lipid-rich tissues, consistent with its physicochemical properties and previous findings in mammals (Oerlemans et al., 2019; Jeong et al., 2022). In our study, TEB was detectable in the liver at a daily dose of 0.02256 mg/kg bw/day, whereas much higher doses were required in pigs. Notably, we observed lower tissue concentrations of TEB in individuals exposed to mixtures containing IMI, similar to findings reported by Man et al. (2023). This may suggest toxicokinetic interactions, such as competition for P450 metabolism (Willow et al., 2019) or altered hepatic transport. However, further confirmation of these mechanisms is required, as available data on such interactions remain limited.

IMI was detected in serum, tissues, and feces, confirming its rapid absorption and active systemic distribution. Previous studies have reported higher IMI concentrations in seminal fluid than in gonadal tissues, likely due to differences in blood-testis barrier permeability and limited metabolic activity in seminal plasma (Mikolić and Karačonji, 2018). Our results support this observation, as IMI persisted in semen even after exposure cessation, which may explain the sustained reduction in fertility observed during phase II. Moreover, the persistence of IMI in feces after the exposure period further supports manure as a reliable nonlethal matrix for monitoring recent pesticide exposure.

Among the tested compounds, GLP displayed the greatest persistence, consistent with reports describing its prolonged elimination phase in birds (Fréville et al., 2023) and its relatively low bioavailability and limited metabolism (Peillex and Pelletier, 2020). In semen, GLP levels remained detectable even after the withdrawal period (phase II), whereas they were almost completely eliminated from serum. This pattern may reflect reduced local blood flow, limited active clearance mechanisms, and possible binding to tissue components, collectively reducing elimination efficiency in reproductive compartments. A similar tendency for GLP to reach relatively high concentrations in seminal plasma compared with blood has been reported previously (Vasseur et al., 2024). The absence of active metabolism of GLP to its primary metabolite, AMPA in seminal fluid further suggests limited metabolic conversion in this compartment, potentiallypromoting its accumulation. Together, these findings indicate that GLP tissue persistence may be influenced by tissue-specific properties and coexposure conditions, supporting the relevance of mixture-specific toxicokinetic interactions.

Immunological modulation in the spleen and cecal tonsils

Previous toxicological studies have demonstrated that tebuconazole (TEB) can affect immune function by inducing cytochrome P450–dependent detoxification pathways and oxidative stress, ultimately promoting apoptosis via mitochondrial and endoplasmic reticulum stress pathways (Li et al., 2022; Othmène et al., 2022). Azole-type fungicides have also been shown to interfere with immune cell survival, proliferation, and differentiation, potentially impairing immune development and responses (Mokarizadeh et al., 2015). However, evidence from avian models remains limited, and chronic dietary exposure to TEB did not significantly alter immune biomarkers in house sparrows (Passer domesticus) receiving 0.164 mg/kg/day for 11 weeks (Bellot et al., 2024).

In the present study, significant immunological alterations were observed primarily in the gut-associated lymphoid tissue. Specifically, the percentage of CD8+ cells in the cecal tonsils was significantly higher in the IMI+GLP group than in the control. In addition, birds exposed to the TEB+IMI+GLP mixture showed a tendency toward increased proportions of both CD4⁺ and CD8⁺ cells in the cecal tonsils (P = 0.066 and P = 0.067, respectively). These findings indicate that dietary exposure to IMI and GLP, particularly in combination, is associated with activation of the cytotoxic cellular immune response in the intestinal immune compartment.

In contrast to the local intestinal response, systemic immune modulation was primarily reflected in splenic B-cell activity. The formation and expansion of germinal centers in the spleen represent a hallmark of T-dependent B-cell responses and typically occur within weeks of antigenic stimulation (MacLennan, 1994). Recent studies have shown that germinal center positioning and responsiveness are tightly regulated, with immune activation driving the migration and proliferation of antigen-committed B and T cells within predefined splenic niches (Avancena et al., 2021; Victora and Nussenzweig, 2022). In our study, chickens exposed to the TEB+GLP combination exhibited significantly larger germinal center areas in the spleen compared with controls, suggesting modulation of the splenic B-cell response. Although the underlying mechanism was not directly investigated, this effect may reflect low-grade immune activation driven by pesticide-induced oxidative stress or cytokine signaling rather than an overall increase in B-cell numbers.

The increased proportion of cytotoxic cells observed in the cecal tonsils of the IMI+GLP group is consistent with previous reports describing immunotoxic effects of imidacloprid (IMI) in poultry. Earlier studies demonstrated that IMI exposure can impair immune organ function and alter cellular immune responses in broiler chickens (Kammon et al., 2012). In red-legged partridge (Alectoris rufa), both high (53.4 mg/kg/day) and low (8.8 mg/kg/day) IMI doses were shown to suppress T-cell responses, as assessed by the phytohemagglutinin test (Lopez-Antia et al., 2013, 2015). However, other studies reported no clear immunosuppressive or immunostimulatory effects following oral IMI exposure in White Leghorn chickens at doses up to 15.5 mg/kg/day (Franzen-Klein et al., 2020), indicating that IMI-induced immunomodulation may depend on dose, exposure duration, and species.

Intestinal immune responses are closely linked to the composition and activity of the gut microbiota (Broom and Kogut, 2018; Kayama, et al., 2020; Lukáčová et al., 2023). In vitro studies have shown that glyphosate-based herbicides can negatively affect the growth and viability of beneficial poultry-associated microbiota, including Enterococcus spp., Bacillus badius, Bifidobacterium adolescentis, and Lactobacillus spp. (Shehata et al., 2013). Therefore, it is plausible that GLP-induced alterations of the intestinal microbiota may indirectly contribute to the observed modulation of local immune responses in the cecal tonsils.

Integrated mechanistic interpretation and conclusions

Collectively, our findings indicate that chronic exposure to sub-MRL levels of TEB, IMI, and GLP results in adverse reproductive outcomes through the convergence of toxicokinetic persistence, mixture-specific interactions, and functionally relevant biological responses. The detection of all compounds in reproductive tissues and body fluids, together with their persistence in semen and partial post-exposure reversibility of selected sperm parameters, provides a mechanistic context linking internal exposure to impaired semen quality, reduced fertility, and increased embryo mortality.

Notably, the strongest reproductive impairments were observed following combined pesticide exposure, highlighting the importance of mixture-specific effects that are not addressed by current regulatory frameworks. The observed modulation of local and systemic immune responses further suggests that immune alterations may represent early and sensitive indicators of chronic pesticide exposure. Taken together, these results indicate that reproductive toxicity can occur at exposure levels far below established NOAELs, emphasizing the relevance of incorporating reproductive endpoints and to account for chronic and combined exposures when defining protective thresholds.

These findings further indicate that pesticide residues not only traverse the avian digestive system but also persist in excreta, thereby contributing to environmental circulation and potential long-term exposure pathways. In this context, current MRLs, largely derived from dietary risk assessments, may not fully capture risks to avian reproductive performance under chronic and combined exposures. Our results support the relevance of incorporating reproductive endpoints, such as semen quality, fertility, and hatchability, into future risk assessment frameworks, including benchmark dose–based approaches.

More broadly, these observations highlight the importance of reassessing safety margins for pesticides with documented endocrine-disrupting or mitochondria-related toxic effects when evaluating long-term reproductive outcomes.

CRediT authorship contribution statement

Skarlet Napierkowska: Writing – review & editing, Writing – original draft, Visualization, Investigation, Formal analysis, Data curation. Pascal Froment: Writing – review & editing, Writing – original draft, Supervision, Methodology, Investigation, Funding acquisition, Conceptualization. Artur Kowalczyk: Writing – review & editing, Methodology, Investigation, Conceptualization. Joanna Rosenberger: Writing – review & editing, Methodology, Investigation. Piotr Kaczyński: Writing – review & editing, Writing – original draft, Methodology, Investigation, Conceptualization. Bożena Łozowicka: Writing – review & editing, Writing – original draft, Methodology, Investigation, Conceptualization. Jan P. Madej: Writing – review & editing, Writing – original draft, Methodology, Investigation. Joelle Dupont: Writing – review & editing, Writing – original draft, Supervision, Methodology, Conceptualization. Mariusz Birger: Writing – review & editing, Investigation. Łukasz Smaga: Writing – review & editing, Formal analysis. Wojciech Niżański: Writing – review & editing, Conceptualization. Agnieszka Partyka: Writing – review & editing, Writing – original draft, Validation, Supervision, Resources, Project administration, Methodology, Funding acquisition, Data curation, Conceptualization.

Disclosures

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.

Appendix. Supplementary materials

Supplementary Materials

• Additional file 1 (PDF): Methods of pesticide residue analysis. Detailed description of the methodology used for the detection and quantification of pesticide residues in biological samples.