Abstract
Background:
Epidemiological studies have shown Per- and polyfluoroalkyl substances (PFAS) to be associated with diseases of dysregulated lipid and sterol homeostasis such as steatosis and cardiometabolic disorders. However, the majority of mechanistic studies rely on single chemical exposures instead of identifying mechanisms related to the toxicity of PFAS mixtures.
Objectives:
The goal of the current study is to investigate mechanisms linking exposure to a PFAS mixture with alterations in lipid metabolism, including increased circulating cholesterol and bile acids.
Methods:
Male and female wild-type C57BL/6J mice were fed an atherogenic diet used in previous studies of pollutant-accelerated atherosclerosis and exposed to water containing a mixture of 5 PFAS representing legacy, replacement, and emerging subtypes (i.e., PFOA, PFOS, PFNA, PFHxS, and GenX), each at a concentration of 2 mg/L, for 12 weeks. Changes at the transcriptome and metabolome level were determined by RNA-seq and high-resolution mass spectrometry respectively.
Results:
We observed increased circulating cholesterol, sterol metabolites, and bile acids due to PFAS exposure, with some sexual dimorphic effects observed. PFAS exposure increased hepatic injury, demonstrated by increased liver weight, hepatic inflammation, and plasma alanine aminotransferase levels. Females displayed increased lobular and portal inflammation compared to the male PFAS-exposed mice. Hepatic transcriptomics analysis revealed PFAS exposure modulated multiple metabolic pathways, including those related to sterols, bile acids, and acyl carnitines, with multiple sex-specific differences observed. Finally, we show that hepatic and circulating levels of PFOA were increased in exposed females compared to males, but this sexual dimorphism was not the same for other PFAS examined.
Discussion:
Exposure of mice to a mixture of PFAS results in PFAS-mediated modulation of cholesterol levels, possibly through disruption of enterohepatic circulation.
Keywords: PFAS toxicity, Per- and polyfluoroalkyl substances, liver injury, cardiometabolic disease, hyperlipidemia, cholesterol, bile acids
Introduction
Per- and polyfluoroalkyl substances (PFAS) are commonly used synthetic chemicals that have been manufactured and used in numerous industries around the world since the 1940’s (Buck et al. 2011). Many PFAS are widely utilized for their surfactant properties, leading to their incorporation in to countless industrial and consumer products, such as cookware, food packaging, clothing, and carpets as well as in aqueous film forming foams used by firefighters (Clara et al. 2008; Nickerson et al. 2020; Trier et al. 2011). Over 4000 different versions of PFAS have been identified, with around 256 PFAS being commercially manufactured at globally relevant levels (Buck et al. 2021; OECD 2018). Long-chain “legacy” PFAS, have been used in products for decades, with two of the most well-studied being perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA). Due to the rising concern over the detrimental effects of PFOA and PFOS, these legacy pollutants were phased out 5-10 years ago. However, many legacy PFAS were replaced by alternatives, such as tridecafluorohexane-1-sulfonic (PFHxS), and even more recently by “emerging” PFAS, featuring shorter carbon chains or fluoroether replacements (e.g., hexafluoropropylene oxide dimer acid; GenX) (Brase et al. 2021; Dodds et al. 2020; Lindstrom et al. 2011; Wang et al. 2015).
Although extremely useful in the manufacturing of industrial and consumer products, the strength and stability of PFAS also results in their high resistance to degradation. Consequently, many legacy PFAS (e.g. PFOA and PFOS) have bioaccumulated in the environment, animals, and humans (Calafat et al. 2007; De Silva et al. 2021; Giesy and Kannan 2001). Consequently, mixtures of both legacy as well as alternative and emerging PFAS continue to be ubiquitously found throughout the environment, notably including food and water supplies all over the globe (Domingo and Nadal 2017, 2019). Furthermore, bioaccumulation of PFAS circulating in human blood has been widely reported, including detection in 98% of adult Americans ((CDC) 2005; Calafat et al. 2007). The biological half-lives for legacy PFAS in human serum have been determined to be around 4, 5, and 8.5 years for PFOA, PFOS, and perfluorohexane sulfonate (PFHxS), respectively (Olsen et al. 2007). Although less is known of the halflives of newer PFAS, evidence points to their excretion rates being considerably increased (Gannon et al. 2016).
Given the extensive bioaccumulation and persistence of PFAS, epidemiological studies investigating adverse health effects in humans due to PFAS exposure is a growing area of research, and numerous epidemiological studies have found PFAS exposure to be associated with chronic diseases such as dyslipidemia and cardiometabolic disorders (Blomberg et al. 2021; Darrow et al. 2016; Fragki et al. 2021; Jin et al. 2020; Liu et al. 2018; Winquist and Steenland 2014). Human studies have also established an association between elevated serum PFAS levels and liver injury, such as increased ALT levels, steatosis, and NAFLD severity (Darrow et al. 2016; Gallo et al. 2012; Gleason et al. 2015; Jain and Ducatman 2019a; Jin et al. 2020; Lin et al. 2010). Furthermore, human epidemiological studies focusing on exposure to several model PFAS have reproducibly found PFAS exposure to be associated with hypercholesteremia (Mora et al. 2018; Nelson et al. 2010; Sakr et al. 2007; Steenland et al. 2009). For example, in populations occupationally exposed to PFOA, with an average serum concentration of 1.13ug/ml PFOA, 1 ppm increase in serum PFOA was found to be associated with a 1.06 mg/dl increase in total cholesterol (Sakr et al. 2007). In communities exposed to PFOA via environmental contamination from PFAS manufacturing, people exhibited high mean serum concentrations of 80ng/ml PFOA (Steenland et al. 2009). These communities also demonstrated a positive association between PFAS exposure and total cholesterol, with a predicted increase in 11-12 mg/dl total cholesterol from lowest to highest decile of PFOA. When studying the general U.S. population, data from the 2003-2004 National Health and Nutrition Examination Survey (NHANES) determined a median serum concentration of 3.9 ng/ml PFOA and 21 ng/ml PFOS (Nelson et al. 2010). This study of “background” PFAS levels were found to show trends consistent with the previous human data, including a positive association between PFAS levels and total cholesterol. In fact, multiple studies show the dose-response for cholesterol is not linear, but rather log-linear or asymptotic, which would provide an explanation for rare and seemingly contradictory findings related to enrollment size and statistical power (Frisbee et al. 2010; Li et al. 2020; Steenland et al. 2009).
While there is large body of historical literature linking PFAS exposures to hypocholesterolemia in rodents (Butenhoff et al. 2012; Han et al. 2018; Loveless et al. 2006; Martin et al. 2007; Qazi et al. 2010; Seacat et al. 2002), more recent work has begun to uncover mechanisms linking single PFAS compounds to increased cholesterol levels and cardiovascular disease risk (Fragki et al. 2021; Rebholz et al. 2016; Schlezinger et al. 2020). These more recent studies often utilized models aimed at more closely mirroring human-relevant conditions. For example, a recent study using mice with liver-specific humanized PPARα demonstrated PFOA exposure to be positively associated with increased serum cholesterol levels in males (Schlezinger et al. 2020). Furthermore, there are interspecies differences in lipoprotein metabolism that also contribute to the discrepancies between mouse and human data. For example, cholesteryl ester transfer protein is completely absent in rodents and Apolipoprotein B is cleared much faster from rodent livers compared to humans, resulting in a higher ratio of HDL-C to LDL-C in rodents (Princen et al. 2016). The vast majority of the early rodent PFAS studies were conducted using a standard rodent chow diet containing low amounts of fat and negligible cholesterol. More recent studies have investigated the effects of PFAS in conjunction with diets containing high amounts of fat and/or cholesterol. One such study by Rebholz et al. found that mice fed a western diet and exposed to dietary PFOA resulted in elevated serum cholesterol levels in both male and female mice (Rebholz et al. 2016). Additional variables that may contribute to the apparent discrepancies between animal and human PFAS studies include single chemical exposures versus real world complex mixtures, chronic versus acute exposures, and interactions with other stressors such as diet.
Mechanistic toxicity studies that utilize mixtures of PFAS are understudied and may add additional evidence linking PFAS exposures to adverse health effects. In the present study, male and female C57BL/6J mice were fed an atherogenic diet and exposed to water containing a mixture of 5 environmentally relevant PFAS (PFOA, PFOS, perfluorononanoic acid (PFNA), PFHxS, and GenX) for 12 weeks. Exposure to our PFAS mixture resulted in elevated serum cholesterol levels as early as 5 weeks post exposure, which remained elevated at study conclusion, even with observed strong PPAR alpha activation. We then used unbiased approaches to examine transcriptome and metabolome level changes due to the PFAS mixture focusing on differences between sexes.
Materials and Methods
Animal Experiments
C57BL/6J male and female mice at 7 weeks of age were purchased from Jackson Laboratories (Bar Harbor, ME, USA) and were allowed to acclimate for 1 week upon arrival. Mice were randomly divided into 4 experimental groups: female+vehicle, female+PFAS, male+vehicle, and male+PFAS. After acclimation, animals were switched to the atherogenic diet for 1 week prior to administration of PFAS. The atherogenic diet (Clinton/Cybulsky low fat diet, 0.15% cholesterol) (Research Diets, New Brunswick, NJ, USA) is described in detail in Supplementary Table 5. and continued on the atherogenic diet throughout the study (Clinton/Cybulsky low fat diet, 0.15% cholesterol) (Research Diets, New Brunswick, NJ, USA). This lowfat atherogenic diet has been used in past studies of persistent organic pollutant toxicity (Petriello et al. 2018). The mice were fed ad libitum. The mice were exposed for 12 weeks via their drinking water to vehicle water alone or to a mixture of 5 PFAS: Perfluorooctane sulfonate (PFOS) (Sigma-Aldrich, St. Louis, MO, USA), perfluorooctanoic acid (PFOA) (Sigma-Aldrich, St. Louis, MO, USA), perfluorononanoic acid (PFNA) (Sigma-Aldrich, St. Louis, MO, USA), tridecafluorohexane-1-sulfonic (PFHxS) (J&K Scientific, Beijing, China) and ammonium perfluoro(2-methyl-3-oxahexanoate) (GenX) (Synquest Laboratories, Alachua, FL, USA). Each of the 5 PFAS was dissolved in water at a concentration of 2 mg/L and water bottles were changed weekly by Division of Laboratory Animal Resources staff. An average adult mouse drinks approximately 4 mL of water a day so this dosing strategy will result in approximately 8 micrograms of individual PFAS per day. For an average mouse of 25 grams, this would approximate to about 0.32 mg/kg per day of individual PFAS. This dosing regimen results in PFAS serum concentrations that mirror high occupational exposure to single PFAS, such as those found in workers at PFAS manufacturing plants, where studies have found average serum PFOA concentrations of 1.3-12.9ug/mL (Costa et al. 2009; Sakr et al. 2007). At study conclusion, mice were fasted overnight and humanely euthanized via carbon dioxide inhalation. The animal protocol was approved by the Wayne State University Institutional Animal Care and Use Committee.
Sample Collection
Weekly measurements were recorded of body weight per mouse and water and food intake per cage (i.e., per 5 mice). After week 6 for the females, and week 5 for the males, COVID-19 restrictions prevented the collection of weekly body weights, food and water intake, and blood collection until end of study. Plasma was collected longitudinally via the submandibular vein at the start of weeks 4, 6, and at study completion via cardiac puncture.
Body composition was measured via an EchoMRI Body Composition Analyzer (Echo Medical System, Houston, TX,) for 5 mice of each group at study conclusion. Tissues were harvested and stored at −80 °C prior to analysis or homogenized in TRI reagent (Thermo-Fisher Scientific). The activity of alanine aminotransferase was measured in the serum by using the Infinity ALT (GPT) Liquid Reagent (Thermo-Fisher Scientific, Waltham, MA). Hepatic triglycerides were determined as described previously. Briefly, lipids were extracted from the tissues using the Folch method as previously described (Folch et al. 1957). In brief, the dried lipids were reconstituted in chloroform and loaded onto a TLC plate. The lipids were separated in a hexane/diethyl ether/acetic acid (80/20/2, v/v/v) solution (Touchstone 1995). Spray the TLC plates lightly with primulin (0.005%) dissolved in acetone/water (80/20, v/v), then visualize bands under UV light. Sections of liver were fixed in 10% neutral-buffered formalin and embedded in paraffin. Sections of liver were analyzed via hematoxylin and eosin staining as well as terminal deoxynucleotidyl transferase dUTP nick end labeling (TUNEL) and examined by a blinded hepatopathologist.
RNA-Sequencing
Total cellular RNA was extracted from frozen liver samples using the TRIzol reagent (Thermo Fisher Scientific). RNA purity and quantity were assessed via spectrophotometer (NanoDrop 2000, Thermo-Fisher Scientific). RNA Sequencing was performed by the Wayne State University Genome Sciences Core. The mRNA-seq library was prepared using the QuantSeq 3’ mRNA-Seq Library Prep Kit FWD (Lexogen). Libraries were assessed by the High Sensitivity D1000 (HS D1000) ScreenTape Assay (Agilent). Samples were sequenced on the NovaSeq system (illumina). Reads were aligned to the mouse genome (Build mm9) and tabulated for each gene region (Anders et al. 2015; Dobin et al. 2013). Differential gene expression analysis was used to compare transcriptome changes between treatment groups and are reported as the fold change between the specified conditions (sex, exposure, and interaction term; exposure x sex), p-value, the Benjamini-Hochberg (BH)-adjusted p-value, and the individual read counts by gene for each sample replicate. Significantly altered genes (|log fold change| ≥ 2; p-value ≤ 0.05) were used to identify affected pathways. Heatmaps were generated using the pheatmap package in R and refined in Adobe Illustrator. Both log2 read counts depicted in red and log2 fold change using a blue and yellow scale are depicted in the heatmaps.
Liver Metabolomic Profiling
Based on the enriched pathways discovered with RNAseq, we then utilized a non-targeted metabolomics approach performed by Metabolon Inc. (Durham, NC) and described previously (Miller et al. 2015). Here we report primarily on a subset of metabolites related to lipid and sterol homeostasis. Seven mice were chosen at random from each group, and corresponding liver and plasma samples from each mouse were shipped to Metabolon. Briefly, samples were prepared using the automated MicroLab STAR® system (Hamilton Company, Salt Lake City, UT) and proteins were precipitated with methanol. Ultrahigh performance liquid chromatography-tandem mass spectroscopy was carried out using a Waters ACQUITY ultra-performance liquid chromatography (UPLC) and a Thermo Scientific Q-Exactive high resolution/accurate mass spectrometer interfaced with a heated electrospray ionization (HESI-II) source and Orbitrap mass analyzer operated at 35,000 mass resolution. Compounds were identified by comparison to library entries of purified standards or recurrent unknown entities. Statistical analysis of fold changes by two-way ANOVA was used to identify metabolites that differed significantly (p<0.05) between experimental groups. Metabolon reported on 879 (liver) and 709 (plasma) compounds of known identity. Q-values for overall PFAS effect and PFAS-sex interaction are also reported. The q-value is calculated as an estimate of the false discovery rate to take into account the multiple comparisons that occur in metabolomic-based studies.
Validation of hepatic gene expression
Total cellular RNA was extracted from frozen liver samples using the TRIzol reagent (Thermo Fisher Scientific). RNA purity and quantity were assessed via spectrophotometer (NanoDrop 2000, Thermo-Fisher Scientific). cDNA was synthesized from the total RNA using the High-Capacity cDNA Reverse Transcription Kit (Thermo Fisher Scientific). Taqman Fast Advanced Master Mix (Thermo Fisher Scientific) was used to perform real-time PCR. The following Taqman probes were used: Cyp7a1 (Mm00484152_m1), Cyp27a1 (Mm00470430_m1), Scl10a1 (Mm00441421_m1), Ephx1 (Mm00468752_m1), Abcc3 (Mm00551550_m1), and GAPDH (Mm99999915_g1) (Thermo Fisher Scientific).
Absolute quantitation of circulating PFAS
Plasma sample extraction was completed using an Agilent EMR-Captiva lipid extraction 96-well plate. The standard curve was extracted on the EMR-Captiva extraction plate using mouse plasma purchased from Sigma Aldrich. The following was done two rows at a time in order to keep the methanol:acetonitrile mixture cold. To begin, 500 μL of cold 15:85 methanol:acetonitrile was added to each well, followed by the PFAS surrogate standard solution to each well in the 96 well extraction plate. A sample size of 50 μL of plasma was added to its specified well, in which it was mixed by plunging up and down with the pipet for complete mixing. The plate was vortexed for 2 minutes at 1100 x g. Using a vacuum apparatus at 2-7 inHg for up to 5 min, the mixture was extracted into the deep well plate. When completely extracted, the plate was brought down to dryness using N2 gas. The plate was capped and stored the 96-well plate in the −80 freezer until ready for analysis. On day of analysis, samples were brought back up to a final volume of 300 μL using 50% methanol with 0.1% acetic acid spike with PFAS internal standard solution. Plate was capped and vortexed for 2 min at 1100 x g.
Samples were analyzed using a Thermo Scientific TSQ Altis with UltiMate 3000 UHPLC. Mobile phases used were A) 20 mM Ammonium acetate in LC-MS Grade Water and B) LC-MS Grade Acetonitrile. Two columns were utilized. A delay column (Thermo Scientific Acclaim RSLC C 18 2.1 x 50 mm, 2.2 μm) was placed after the high-pressure mixer in the system. The analytical column used is a Thermo Scientific Accucore C8 100 x 2.1 mm, 2.6 μm and was held at 40°C. The gradient started at 5% B from 0-1 min, increased to 30% B from 1-3 min, then to 65% B from 3-14 min. The column was rinsed at 98% B from 14-17.5 min and allowed to equilibrate at 5% B from 17.5-20 min. Total run time was 20 minutes. Injection volume was 5 μL. PFAS were detected in negative ion mode using electrospray ionization. Source settings were as follows: negative ion voltage at 1000 V, sheath gas 50 (arbitrary units), aux gas at 12 (arbitrary units), sweep gas at 0.5 (arbitrary units), ion transfer tube temperature at 275°C, and vaporizer temperature at 225°C. Limits of detection for the 5 PFAS were: HFPO-DA; 1 ng/mL, PFOA; 5 ng/mL, L-PFHxS; 0.47 ng/mL, PFNA; 5.0 ng/mL, L-PFOS; 0.48 ng/mL.
All solvents used for LC-MS analysis were LC-MS grade quality. LC-MS grade methanol and acetonitrile, and mouse plasma were purchased from Sigma Aldrich (Burlington, MA, USA). LC-MS grade water, acetic acid (1 mL glass vials), and ammonium acetate were purchased from Fisher Scientific (Waltham, MA, USA). Standards used for analytical determination of PFAS in mouse plasma were Native PFAS linear Primary dilution standard and Method 537 Internal standard mix from Wellington Laboratories, Inc. (Overland Park, KS, USA), and 13C6 PFHxS, 13C9 PFNA, 13C8 PFOS, and 13C8 PFOA were purchased from Cambridge Isotope Laboratories, Inc (Tewksbury, MA, USA).
Data Analysis
Statistical analyses of RNA sequencing gene expression data was performed using R Statistical Software. Statistical analyses of animal measurement data and gene expression data analyzed in Figure 1 were performed using the SigmaPlot version 14 (Systat Software Inc, San Jose, CA, USA). Linear regression analysis was performed using JMP Statistical Software (SAS Institute). Multiple group data were compared using Two Way ANOVA, followed by the Holm-Sidak method post-hoc test for multiple comparisons. Graphed values are expressed as mean ± SEM. A probability value of p ≤ 0.05 was considered statistically significant.
Figure 1. PFAS exposure induces sex-dependent liver injury.
Male and female C57BL/6J mice were exposed to vehicle water or the PFAS mixture for 12 weeks: female+vehicle (n=10), female+PFAS (n=9), male+vehicle (n=10), and male+PFAS (n=10). (A) Male and female liver sections were stained for H&E. Male PFAS mouse: The hepatocytes are swollen with eosinophilic cytoplasm, indicating hepatocyte injury. Scattered acidophil bodies (hepatocyte necrosis) are also present (arrows). Female PFAS mouse: the portal triad demonstrates mild chronic inflammation with lymphocytes. The hepatocytes are swollen with eosinophilic cytoplasm, consistent with injury. In this mouse, hepatocellular and canalicular cholestasis (yellow pigment) is also present. (B) Liver weight given as a percentage of total body weight. (C) Liver sections were analyzed and quantified for number of inflammatory foci or eosinophilic cells per field of view. (D) Percentage of hepatocytes demonstrating eosinophilic cytoplasm (E) Effects of PFAS exposure on hepatic gene expression of liver injury- and toxicant-associated fatty liver disease-related markers. (F) Effects of PFAS exposure on hepatic gene expression of oxidative stress-related markers. (G) Effects of PFAS on 5-Hete metabolite levels. Bars represent mean ± S.E.M. of ten mice (RNA sequencing) or seven mice (Metabolomics Scaled Counts) in each group. Statistical significance for all (p<0.05) was determined by two-way ANOVA analysis and post-hoc comparisons by Holm-sidak method.
Results
3.1. PFAS exposure induces sex-dependent anthropometric changes, oxidative stress, and liver injury
To begin to characterize PFAS mixture toxicity, we measured impacts on body weight and composition as well as glucose and cholesterol. Male mice exposed to PFAS gained ~30% more body weight than male controls during weeks 1-3 of exposure (p=0.039), but this weight gain was not observed in females (p=0.596) (overall exposure effect; p=0.067, interaction; p=0.264) (Supplementary Figure 1D). This was not due to increased food consumption in the mice during these weeks (Supplementary Figure 1C). However, this observation was not retained throughout the duration of the study. Comparing weights at euthanasia to baseline, female control mice body weights increased ~26% and PFAS-exposed females decreased ~18% (p<0.001), while male control mice increased ~37% and PFAS-exposed males decreased ~30% from baseline (p<0.001) (interaction; p<0.001) (Supplementary Figure 1A&D). Additionally, water intake was significantly increased due to PFAS exposure only in female mice during all weeks with available consumption data (i.e., Weeks 1-5; p<0.01), while there was no significant changes in the males (Supplementary Figure 1B). Water intake values were not recorded after Week 4 for males due to COVID-19 restrictions. At euthanasia, body composition was determined by echoMRI. In agreement with the body weight loss, percent body fat was also significantly decreased in the mice exposed to the PFAS mixture for 12 weeks compared to vehicle. The percent body fat of the females decreased by ~45% (p=0.036) after 12 weeks PFAS compared to female controls, while the males decreased by ~72% (p<0.001) compared to male controls (overall exposure effect; p<0.001, interaction; p=0.123) (Supplementary Figure 1E). At the same time, percent lean body mass was not significantly affected in the PFAS-exposed male or female mice compared to their respective control mice, although the overall effect from PFAS exposure regardless of sex was significantly increased (overall exposure effect; p=0.035, interaction; p=0.644) (Supplementary Figure 1F). Fasted glucose levels at study conclusion were significantly decreased by 29% in PFAS-exposed females compared to corresponding vehicle (p<0.001) and decreased by 49% in PFAS-exposed males compared to their corresponding vehicle controls (p<0.001) (interaction; p=0.007) (Supplementary Figure 1G). Total cholesterol levels measured in unfasted PFAS-exposed females at week 4 of the study (3 weeks of PFAS exposure) demonstrated no significant difference compared to vehicle, with cholesterol levels measuring 130 and 126 mg/dl in vehicle and PFAS-exposed female mice, respectively (p=0.68) (Supplementary Figure 1H). At this same timepoint total cholesterol levels in males were also not significantly affected by PFAS exposure, with male total cholesterol levels of 69 mg/dl in male vehicle mice and 78 mg/dl in male PFAS-exposed mice (p=0.29) (Supplementary Figure 1I). However, total cholesterol levels measured in unfasted females after 5 weeks of PFAS exposure demonstrated significantly higher levels in the PFAS-exposed female mice, at 97 mg/dl, compared with vehicle treated female mice, at 54 mg/dl (p<0.001) (Supplementary Figure 1H). Corresponding male samples for this time point were not available.
To characterize the hepatic insult elicited from a mixture of five PFAS (PFOA, PFOS, PFNA, PFHxS, and GenX) in male and female mice after 12 weeks of exposure, liver injury and inflammatory responses were analyzed. Hematoxylin and eosin (H&E) stained liver sections of both male and female mice exposed to either the vehicle water or the PFAS mixture were examined by a hepatopathologist (Figure 1A). Liver weight (given as a percentage of body weight) was significantly increased after 12 weeks of PFAS exposure in both females and males, with females increasing by ~3-fold (p<0.001) and males increasing by ~3.1-fold (p<0.001) over corresponding controls of the same sex (overall exposure effect; p<0.001, interaction; p=0.584) (Figure 1B). Exposure of mice to the PFAS mixture resulted in hepatocyte hypertrophy and injury, as indicated by the enlargement of the hepatocytes, increased eosinophilic cytoplasm, focal cholestasis, and rare acidophil bodies (hepatocyte necrosis) compared to vehicle mice irrespective of sex (Fig.1A&D). In the female PFAS-exposed mice, the portal triad also demonstrates mild chronic inflammation with lymphocytes, as well as hepatocellular and canalicular cholestasis (Fig. 1A). Furthermore, PFAS exposed females appeared to display increased lobular and portal inflammation compared to the male PFAS-exposed mice (Figure 1C). TUNEL staining of liver sections revealed no increase in apoptosis due to PFAS exposure in either sex (results not shown). Finally, circulating plasma levels of alanine aminotransferase (ALT) significantly increased from 5.2U/L to 134.8U/L in PFAS-exposed females compared to vehicle (p<0.001) and 5.2U/L to 100.6U/L in PFAS-exposed males compared to vehicle (p<0.001) (exposure effect; p<0.001, interaction; p=0.788) (Figure 1H).
We then utilized results from RNA-seq analyses to examine transcriptional changes due to PFAS exposure focusing on a subset of genes related to toxicant-associated fatty liver disease (TAFLD) (Baselli et al. 2020; Cave et al. 2016; Coilly et al. 2019; He et al. 2021). Hepatic mRNA levels of keratin 18 (Krt18) were upregulated 2.2-fold in PFAS-exposed females compared to vehicle (p<0.001) and 1.8-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; p<0.001, interaction; p=0.269) (Figure 1E). Hepatic mRNA levels of fatty acid binding protein 4 (Fabp4) were upregulated 19.8-fold in PFAS-exposed females compared to vehicle (p<0.001) and 15-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; p<0.001, interaction; p=0.696) (Figure 1E). Hepatic mRNA levels of fibroblast growth factor 21 (Fgf21) were upregulated 13.5-fold in PFAS-exposed females compared to vehicle (p<0.001) and 15.1-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; p<0.001, interaction; p=0.464) (Figure 1E). Hepatic mRNA levels of albumin (Alb) were downregulated 3.78-fold in PFAS-exposed females compared to vehicle (p<0.001) and 4-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; p<0.001, interaction; p=0.686) (Figure 1E). In addition, results from our RNA-seq analyses revealed transcriptional changes due to PFAS exposure in a subset of genes related to oxidative stress, many of which displayed a significant interaction between sex and PFAS exposure (Figure 1F). For example, hepatic mRNA levels of glutathione s-transferase mu 1 (Gstm1) were upregulated by 2.7-fold in PFAS-exposed females compared to vehicle (p<0.001) and 3.9-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; p=0.049) (Figure 1F). Hepatic mRNA levels of NAD(P)H quinone dehydrogenase 1 (Nqo1) were upregulated by 3.8-fold in females compared to vehicle (p<0.001) and 11.3-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; p=0.012) (Figure 1F). Hepatic mRNA levels of superoxide dismutase 2 (Sod2) were upregulated by 1.5-fold in PFAS exposed females compared to corresponding vehicle (p<0.001), while there was no significant change in males due to PFAS exposure (interaction; p=0.007) (Figure 1F). Finally, to investigate if the upregulated oxidative stress-induced genes correspond to increased biomarkers of oxidative stress, we examined the hepatic and circulating metabolome and determined that hepatic levels of 5-hydroxyeicosatetraenoic acid (5-HETE) were increased by ~3.9-fold (p<0.001) in PFAS exposed females, but no significant differences in 5-HETE were observed between male groups (interaction; p=0.017) (Figure 1G).
3.2. PFAS exposure modulates hepatic and circulating bile acid and sterol levels
To begin to investigate the observation that PFAS may impact on cholesterol homeostasis and flux, we completed metabolomic analysis on hepatic tissue and plasma for both sexes. Clear differences were observed between PFAS-exposed and vehicle mice in both the liver and plasma, indicating substantial alteration in the metabolic profile in the mice in both liver and plasma after PFAS exposure (Table 1). Except for taurohyocholate, all measured bile acids (primary, conjugated, and secondary) which had significant fold changes were decreased in the livers of PFAS-exposed mice, while all the primary bile acids showed a significant increase in the plasma. The full list of sterols, primary bile acids, and secondary bile acid metabolites measured in liver tissue and plasma are included in Supplementary Tables 1 and 2, respectively. Table 1 shows relative fold change increases or decreases in several common sterols, primary bile acids, and secondary bile acids, both in the liver tissue and in the circulating plasma. At euthanasia, fasted plasma cholesterol levels in PFAS-treated mice were increased 1.62-fold in females (p=0.0017) and 1.33-fold in males (p=0.048) compared to their controls (overall exposure effect; q<0.01, interaction; n.s.) (Table 1). Hepatic cholesterol levels, however, were not significantly altered females (p=0.78) or males (p=0.51) (overall exposure effect; q=0.14, interaction; q=0.43). Hepatic triglyceride levels were also measured using thin-layer chromatography and both sexes exhibited decreased total triglyceride levels (Supplementary Figure 6). Notably, hepatic levels of the sterol metabolite 4-cholesten-3-one was increased by 5.23-fold in PFAS-exposed females compared to vehicle (p<0.001) and 15.41-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; q=0.01) (Table 1). Correspondingly, plasma levels of 4-cholesten-3-one increased by 6.06-fold in PFAS-exposed females compared to vehicle (p<0.001) and 16.51-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; q<0.01) (Table 1). Hepatic levels of the sterol metabolite cholesterol sulfate were reduced 0.26-fold in PFAS-exposed females compared to vehicle (p<0.001) and 0.28-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.48). Plasma levels of the sterol metabolite cholesterol sulfate were not significantly altered in females (p=0.13) or males (p=0.62) (overall exposure effect; q<0.05, interaction; q=0.70). Hepatic levels of the sterol metabolite 7-alpha-hydroxy-3-oxo-4-cholestenoate (7-Hoca) were reduced 0.51-fold (p=0.003) in PFAS-exposed females compared to vehicle, while the effect on males was not significant (p=0.27) (overall exposure effect; q<0.01, interaction; p=0.20). However, plasma levels of 7-Hoca were increased 2.22-fold (p=0.002) in PFAS-exposed males compared to vehicle, while the effect on females was not significant (p=0.72) (overall exposure effect; q<0.01, interaction; p=0.18) (Table 1).
Table 1. PFAS exposure modulates hepatic and circulating sterol and bile acid levels.
Hepatic protein and plasma samples from n=7 mice from each treatment group were analyzed by a non-targeted metabolomics approach using LC-MS by Metabolon Inc. Relative fold change between various treatment groups is given for hepatic sterols and bile acid metabolites. Statistical significance for the interaction and main effects of PFAS exposure and gender was determined by two-way ANOVA analysis. Red indicates a significant relative fold change increase (p<0.05). Blue indicates a significant relative fold change reduction (p<0.05). q-values for overall PFAS treatment and PFAS-sex interaction are displayed in the right columns, with significant (q<0.05) q-values indicated in bold.
| Biochemical Name | Fold of Change PFAS Female vs. PFAS Male | Fold of Change PFAS Female vs. Vehicle Female | Fold of Change PFAS Male vs. Vehicle Male | Overall PFAS Treatment q-value | Sex-PFAS Treatment Interaction q-value |
|---|---|---|---|---|---|
| Hepatic Sterols and Bile Acids | |||||
| Sterols | |||||
| cholesterol | 1.01 | 1.02 | 0.95 | q=0.14 | q=0.43 |
| cholesterol sulfate | 1.29 | 0.26 | 0.28 | q<0.01 | q=0.48 |
| 7-alpha-hydroxy-3-oxo-4-cholestenoate (7-Hoca) | 1.10 | 0.51 | 0.81 | q<0.01 | q=0.20 |
| 4-cholesten-3-one | 1.02 | 5.23 | 15.41 | q<0.01 | q=0.01 |
| Primary Bile Acid Metabolism | |||||
| cholate | 0.82 | 0.37 | 0.49 | q=0.01 | q=0.49 |
| chenodeoxycholate | 0.97 | 0.34 | 0.21 | q<0.01 | q=0.37 |
| beta-muricholate | 1.09 | 0.23 | 0.17 | q<0.01 | q=0.33 |
| alpha-muricholate | 1.60 | 0.12 | 0.12 | q<0.01 | q=0.30 |
| tauro-beta-muricholate | 1.24 | 0.52 | 0.95 | q<0.05 | q=0.33 |
| Secondary Bile Acid Metabolism | |||||
| deoxycholate | 2.30 | 0.17 | 0.16 | q<0.01 | q=0.58 |
| taurodeoxycholate | 0.93 | 0.04 | 0.14 | q<0.01 | q=0.01 |
| ursodeoxycholate | 1.13 | 0.11 | 0.08 | q<0.01 | q=0.32 |
| tauroursodeoxycholate | 1.25 | 0.23 | 0.43 | q<0.01 | q=0.28 |
| taurohyocholate | 0.91 | 1.07 | 2.82 | q<0.01 | q=0.15 |
| taurochenodeoxycholic acid (7 or 27)-sulfate | 1.00 | 0.21 | 0.03 | q<0.01 | q<0.01 |
| Plasma Sterols and Bile Acids | |||||
| Sterols | |||||
| cholesterol | 0.75 | 1.62 | 1.33 | q<0.01 | q=0.60 |
| cholesterol sulfate | 0.80 | 0.86 | 0.96 | q<0.05 | q=0.70 |
| 7-alpha-hydroxy-3-oxo-4-cholestenoate (7-Hoca) | 0.79 | 1.11 | 2.22 | q<0.01 | q=0.18 |
| 4-cholesten-3-one | 0.72 | 6.06 | 16.51 | q<0.01 | q<0.01 |
| Primary Bile Acid Metabolism | |||||
| cholate | 0.93 | 2.86 | 12.43 | q<0.01 | q=0.21 |
| chenodeoxycholate | 1.39 | 2.44 | 2.75 | q<0.01 | q=0.89 |
| beta-muricholate | 1.01 | 2.32 | 6.92 | q<0.01 | q=0.24 |
| Secondary Bile Acid Metabolism | |||||
| deoxycholate | 0.84 | 0.37 | 2.66 | q=0.12 | q<0.01 |
| taurodeoxycholate | 0.95 | 0.14 | 1.27 | q<0.01 | q<0.01 |
| ursodeoxycholate | 0.96 | 0.70 | 3.02 | q<0.05 | q=0.34 |
| tauroursodeoxycholate | 0.83 | 3.36 | 11.48 | q<0.01 | q=0.10 |
| hyocholate | 0.83 | 4.85 | 20.98 | q<0.01 | q=0.14 |
As the main route of cholesterol metabolism and turnover is through conversion to bile acids (di Gregorio et al. 2021), we next report impacts of PFAS exposure on bile acid profiles. Hepatic metabolomics revealed that the majority of primary bile acids were significantly reduced by PFAS exposure in both males and females, indicating possible changes in bile acid synthesis, excretion, export, or reuptake (Table 1). Hepatic levels of chenodeoxycholate were reduced by 0.34-fold in PFAS-exposed females compared to vehicle (p=0.009) and reduced by 0.21-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.37). Hepatic levels of β-muricholate were reduced by 0.23-fold in PFAS-exposed females compared to vehicle (p<0.001) and reduced by 0.17-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; p<0.01, interaction; p=0.33). Hepatic levels of α-muricholate were reduced by 0.12-fold in PFAS-exposed females compared to vehicle (p<0.001) and reduced by 0.12-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.30). However, hepatic levels of the primary bile acid cholate was not significantly affected in males (p=0.18) or females (p=0.054) (overall exposure effect; q<0.01, interaction; q=0.49).
Contrary to the liver, metabolomic analysis of mouse plasma revealed that circulating levels of the majority of primary bile acids were significantly increased by PFAS exposure in both males and females compared to vehicle. Plasma levels of cholate were elevated 2.86-fold in PFAS-exposed female compared to vehicle (p=0.025), while levels in PFAS-exposed males were elevated 12.43-fold compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.21). Plasma levels of β-muricholate were increased by 2.32-fold in PFAS-exposed females compared to vehicle (p=0.021) and 6.92-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.24). Although the measured levels of secondary bile acids reported in Table 1 represent hepatic tissue levels as opposed to those of the small intestine, some insight can be gained by analysis of the measured levels in the liver and plasma. Hepatic levels of deoxycholate, ursodeoxycholate, and tauroursodeoxycholate were all significantly decreased, ranging from 0.08-fold to 0.43-fold reduced (p<0.01), in PFAS-exposed males or females compared to their respective vehicles (overall exposure effect; q<0.01 for all). Hepatic levels of taurodeoxycholate decreased by 0.04-fold in PFAS-exposed females compared to vehicle (p<0.001) and decreased by 0.14-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; q=0.01). Hepatic levels of taurochenodeoxycholic acid (7 or 27)-sulfate decreased by 0.21-fold in PFAS-exposed females compared to vehicle (p=0.37) and decreased by 0.03-fold in PFAS-exposed males compared to vehicle (p=0.12) (interaction; q<0.01). Interestingly, hepatic levels of the secondary bile acid taurohyocholate, produced by taurine conjugation of hyocholic acid, showed a 2.82-fold increase in PFAS-exposed males compared to vehicle (p=0.006) and no significant change in PFAS-exposed females compared to vehicle (p=0.63) (overall exposure effect; q<0.01, interaction; q=0.15). As opposed to the liver, plasma levels of tauroursodeoxycholate were increased 3.36-fold in PFAS-exposed females compared to vehicle (p=0.002) and 11.48-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.10). Plasma levels of hyocholate were increased 4.85-fold in PFAS-exposed females compared to vehicle (p=0.003) and 20.98-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.14). Notably, levels of taurodeoxycholate were significantly reduced 0.14-fold only in PFAS-exposed females compared to vehicle (p<0.001), while levels in males were not significantly affected (p=0.91) (interaction; q<0.01). Deoxycholate was significantly reduced by 0.37-fold in PFAS-exposed females compared to vehicle (p=0.003), while levels in PFAS-exposed males were increased by 2.66-fold compared to vehicle (p=0.002) (interaction; q<0.01).
3.4. Acyl carnitine levels are dysregulated due to PFAS in sex specific ways
Given the increased markers of oxidative stress we observed in the livers of PFAS-exposed mice, we next examined whether PFAS exposure increased mitochondrial fatty acid oxidation, which has been proposed as a pathway leading to reactive oxygen species production (Zorov et al. 2014). Our metabolomics results revealed relative changes in the levels of ketone bodies and several acyl carnitines in the liver and plasma, as shown in Supplementary Table 3. At euthanasia, relative hepatic levels of acetylcarnitine were not significantly altered in PFAS-exposed females compared to vehicle (p=0.053), but were increased 8.21-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.05). Hepatic levels of hexanoylcarnitine were increased by 1.58-fold in PFAS-exposed females compared to vehicle (p<0.05) and 11.54-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; q=0.041). Hepatic levels of octanoylcarnitine were increased by 7.33-fold in PFAS-exposed males compared to vehicle (p<0.001), while females showed no differences (p=0.21) (interaction; q=0.006). Hepatic levels of decanoylcarnitine were increased by 1.38-fold in PFAS-exposed females compared to vehicle (p=0.035) and 7.49-fold in PFAS-exposed males compared to corresponding vehicle (p<0.001) (interaction; q=0.001). Hepatic levels of laurylcarnitine were increased by 1.37-fold in PFAS-exposed females compared to vehicle (p=0.011) and 8.97-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; q=0.032). Hepatic levels of myristoylcarnitine were increased by 2.04-fold in PFAS-exposed males compared to vehicle (p=0.004), while females showed no significant differences (p=0.96) (overall exposure effect; q<0.01, interaction; q=0.085). Hepatic levels of stearoylcarnitine were increased by 1.7-fold in PFAS-exposed females compared to vehicle (p=0.001) and 5.43-fold in PFAS-exposed males compared to vehicle (p<0.001) (interaction; q=0.006). Hepatic levels of arachidoylcarnitine were decreased by 0.19-fold in PFAS-exposed females compared to vehicle (p<0.001), while males were not significantly affected (p=0.36) (interaction; q<0.001). Hepatic levels of behenoylcarnitine were decreased by 0.35-fold in PFAS-exposed males compared to vehicle (p<0.001), while females were not significantly affected (p=0.17) (overall exposure effect; q<0.01, interaction; q=0.065). Hepatic levels of lignoceroylcarnitine were decreased by 0.58-fold in PFAS-exposed females compared to vehicle (p<0.001) and 0.50-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.35).
Conversely, the relative circulating levels of acylcarnitines were largely decreased in the plasma of PFAS-exposed mice (Supplementary Table 3). Plasma levels of hexanoylcarnitine were decreased by 0.48-fold in PFAS-exposed females compared to vehicle (p=0.023), while males showed no significant differences (p=0.27) (overall exposure effect; q<0.01, interaction; q=0.64). Plasma levels of octanoylcarnitine were decreased by 0.39-fold in PFAS-exposed females compared to vehicle (p=0.002) and 0.51-fold in PFAS-exposed males compared to vehicle (p=0.013) (overall exposure effect; q<0.01, interaction; q=0.78). Plasma levels of laurylcarnitine were decreased by 0.45-fold in PFAS-exposed females compared to vehicle (p=0.007) and 0.57-fold in PFAS-exposed males compared to vehicle (p=0.022) (overall exposure effect; q<0.01, interaction; q=0.87). Plasma levels of myristoylcarnitine were decreased by 0.37-fold in PFAS-exposed females compared to vehicle (p<0.001) and 0.62-fold in PFAS-exposed males compared to vehicle (p=0.040) (overall exposure effect; q<0.01, interaction; q=0.52). Plasma levels of palmitoylcarnitine were decreased by 0.50-fold in PFAS-exposed females compared to vehicle (p<0.001) and 0.67-fold in PFAS-exposed males compared to vehicle (p=0.016) (overall exposure effect; q<0.01, interaction; q=0.66). Plasma levels of stearoylcarnitine were decreased by 0.37-fold in PFAS-exposed females compared to vehicle (p<0.001) and 0.43-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.81). Plasma levels of arachidoylcarnitine were decreased by 0.69-fold in PFAS-exposed males compared to vehicle (p=0.005), while females showed no significant differences (p=0.83) (overall exposure effect; q<0.01, interaction; q=0.24).
Hepatic levels of the ketone body 3-hydroxybutyrate (BHBA), which is produced via hepatic fatty acid oxidation and circulate to other organs during gluconeogenesis, was increased in the liver by 3.35-fold in PFAS-exposed females compared to vehicle (p<0.001) and increased by 3.50-fold in PFAS-exposed males compared to vehicle (p<0.001) (overall exposure effect; q<0.01, interaction; q=0.58). In the plasma, relative levels of BHBA were also increased 3.94-fold in PFAS-exposed females compared to vehicle (p=0.008) and increased by 4.62-fold in PFAS-exposed males compared to vehicle (p=0.005) (overall exposure effect; q<0.01, interaction; q=0.91).
3.3. PFAS effects on expression of genes related to metabolism of bile acids and fatty acids
To identify possible transcriptional mechanisms that may explain the observed metabolomics results we performed hepatic RNA sequencing. Gene ontology enrichment analyses were performed using the Database for Annotation, Visualization, and Integrated Discovery (DAVID) software, and heat maps were created for genes enriched for multiple pathways related to lipid and sterol homeostasis. Figure 2A depicts the heatmap generated using genes identified as part of the sterol or bile acid biosynthetic, metabolic, or catabolic pathways. We examined the following 6 genes related to sterol metabolic processes (GO:0016125~sterol metabolic process): Nsdhl, Apoa4, Fgf1, Ebp, Cebpa, and Vldlr. mRNA levels of Nsdhl, Fgf1, Ebp, and Cebpa all markedly decreased by 0.18-0.42 fold change in the PFAS-exposed mice compared to vehicle (overall exposure effect; p<0.01). The mRNA of Apoa4 markedly increased by 3.83-fold change in the PFAS-exposed mice compared to vehicle (overall exposure effect; p<0.01). mRNA levels of Vldlr demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels increased by 4.57-fold in PFAS-exposed females compared to vehicle (p<0.05) and increased by 57.2-fold in PFAS-exposed males compared to corresponding vehicle treated controls (p<0.05) (interaction; p<0.01) (Figure 2A and Table 2).
Figure 2. PFAS exposure modulates genes related to metabolism of bile acids and fatty acids.
Hepatic mRNA was isolated and analyzed via RNA sequencing. Mice in each treatment group: female+vehicle (n=10), female+PFAS (n=9), male+vehicle (n=10), and male+PFAS (n=10). (A) Heatmap of PFAS-induced effects and sex-related effects on hepatic gene expression pattern for bile acid metabolism-related genes. (B) Heatmap of PFAS and sex-related effects on hepatic gene expression pattern for fatty acid metabolism-related genes. Red = indicates scaled raw counts. Yellow/Blue = indicates fold-change (FC) of the interaction between PFAS exposure and sex. #Statistically significant interaction (p<0.05) between PFAS exposure and sex was determined by Two-Way ANOVA. *Statistically significant effect from PFAS exposure.
Table 2. PFAS exposure modulates genes related to metabolism of bile acids and fatty acids.
Liver samples from each treatment group were analyzed by RNA sequencing. Mice in each treatment group: female+vehicle (n=10), female+PFAS (n=9), male+vehicle (n=10), and male+PFAS (n=10). List of hepatic bile acid metabolism- and fatty acid metabolism-associated genes and their fold change due to overall PFAS exposure. Significant p-values and significant Benjamini-Hochberg (BH)-adjusted p-values are denoted in bold (p<0.05) for overall PFAS exposure and PFAS-sex interaction.
| Gene name | Overall PFAS Exposure Fold Change | Overall PFAS Exposure p-value | Overall PFAS Exposure BH-adjusted p-value | PFAS-Sex Interaction p-value | PFAS-Sex Interaction BH-adjusted p-value |
|---|---|---|---|---|---|
| GO:0016125~sterol metabolic process | |||||
| Nsdhl | 0.41 | p<0.001 | p<0.001 | p=0.003 | p=0.42 |
| Apoa4 | 3.83 | p<0.001 | p<0.001 | p=0.26 | p=0.72 |
| Fgf1 | 0.18 | p<0.001 | p<0.001 | p=0.40 | p=0.91 |
| Ebp | 0.38 | p<0.001 | p<0.001 | p=0.27 | p=0.74 |
| Cebpa | 0.42 | p<0.001 | p<0.001 | p=0.14 | p=0.52 |
| Vldlr | 4.57 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| GO:0008206~bile acid metabolic process | |||||
| Cyp46a1 | 0.22 | p<0.001 | p=0.002 | p=1.0 | p=1.0 |
| Cyp7b1 | 0.33 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Slc27a5 | 0.36 | p<0.001 | p<0.001 | p=0.53 | p=1.0 |
| Cyp7a1 | 0.14 | p<0.001 | p<0.001 | p=0.01 | p=0.09 |
| Cyp27a1 | 0.56 | p<0.001 | p<0.001 | p=0.25 | p=0.70 |
| Akr1d1 | 0.06 | p<0.001 | p<0.001 | p=0.14 | p=0.52 |
| Cyp39a1 | 0.40 | p<0.001 | p=0.009 | p<0.001 | p<0.001 |
| Hsd3b7 | 0.79 | p=0.005 | p=0.06 | p=0.01 | p=0.10 |
| Cyp4a12a | 1762.4 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Cyp4a12b | 2454.0 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Cyp8b1 | 1.24 | p=0.15 | p=0.54 | p<0.001 | p=0.005 |
| GO:0015721~bile acid and bile salt transport | |||||
| Atp8b1 | 0.47 | p<0.001 | p<0.001 | p=0.57 | p=1.0 |
| Slco1a4 | 0.05 | p<0.001 | p<0.001 | p<0.001 | p=0.005 |
| Slc10a1 | 0.15 | p<0.001 | p<0.001 | p=0.002 | p=0.03 |
| Slc10a2 | 0.05 | p<0.001 | p<0.001 | p=0.34 | p=0.84 |
| Ephx1 | 6.14 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Abcc2 | 2.18 | p<0.001 | p<0.001 | p=0.12 | p=0.48 |
| Abcc3 | 3.85 | p<0.001 | p<0.001 | p=0.07 | p=0.32 |
| Cyp2c70 | 0.27 | p<0.001 | p<0.001 | p=0.002 | p=0.02 |
| Abcb11 | 0.35 | p<0.001 | p<0.001 | p=0.59 | p=1.0 |
| GO:0019395~fatty acid oxidation | |||||
| Bdh2 | 0.13 | p<0.001 | p<0.001 | p=0.33 | p=0.81 |
| Acacb | 0.46 | p<0.001 | p<0.001 | p=0.001 | p=0.02 |
| Ech1 | 3.66 | p<0.001 | p<0.001 | p=0.23 | p=0.67 |
| Acoxl | 2.32 | p=0.18 | p=0.59 | p=0.30 | p=0.78 |
| Hsd17b4 | 3.57 | p<0.001 | p<0.001 | p=0.01 | p=0.11 |
| Sesn2 | 2.08 | p=0.03 | p=0.18 | p=0.07 | p=0.33 |
| Acox2 | 3.08 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Cpt1b | 26.5 | p<0.001 | p<0.001 | p=0.81 | p=1.0 |
| Acad10 | 2.30 | p<0.001 | p<0.001 | p=0.93 | p=1.0 |
| Acaa1b | 5.80 | p<0.001 | p<0.001 | p=0.96 | p=1.0 |
| GO:0046486~glycerolipid metabolic process | |||||
| Dgkh | 29.2 | p<0.001 | p<0.001 | p=0.79 | p=1.0 |
| Fitm2 | 2.92 | p<0.001 | p<0.001 | p=0.49 | p=1.0 |
| Mogat2 | 2.23 | p=0.20 | p=0.64 | p=0.83 | p=1.0 |
| Lpin3 | 2.44 | p=0.003 | p=0.04 | p=0.47 | p=0.98 |
| Acsl4 | 3.85 | p<0.001 | p<0.001 | p=0.22 | p=0.65 |
| Pnpla1 | 0.42 | p=0.11 | p=0.45 | p=1.0 | p=1.0 |
| Lpl | 18.2 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Tnxb | 0.38 | p=0.01 | p=0.09 | p=0.008 | p=0.07 |
| Gpld1 | 0.12 | p<0.001 | p<0.001 | p=0.79 | p=1.0 |
| Slc27a5 | 0.36 | p<0.001 | p<0.001 | p=0.53 | p=1.0 |
| Cps1 | 0.22 | p<0.001 | p<0.001 | p=0.93 | p=1.0 |
| Pck1 | 0.36 | p<0.001 | p<0.001 | p=0.01 | p=0.10 |
| Lpin1 | 0.20 | p<0.001 | p<0.001 | p=0.25 | p=0.71 |
| Ddhd2 | 4.94 | p<0.001 | p<0.001 | p=0.01 | p=0.09 |
| Dagla | 3.45 | p=0.03 | p=0.21 | p=0.07 | p=0.34 |
| Esr1 | 0.17 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
We next examined the following 11 genes related to bile acid biosynthetic processes (GO:0008206~bile acid metabolic process): Cyp46a1, Cyp7b1, Slc27a5, Cyp7a1, Cyp27a1, Akr1d1, Cyp39a1, Hsd3b7, Cyp4a12a, Cyp4a12b, and Cyp8b1 (Figure 2 and Table 2). The mRNA of Cyp46a1, Slc27a5, Cyp7a1, Cyp27a1, and Akr1d1 all markedly decreased by 0.06-0.56 fold change in the PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). mRNA levels of Cyp7b1 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.32-fold in PFAS-exposed females compared to vehicle (p<0.05) and decreased by 0.03-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01). mRNA levels of Cyp39a1 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.38-fold in PFAS-exposed females compared to vehicle (p<0.05), while males were not significantly affected (interaction; p<0.01). Furthermore, mRNA levels of Cyp4a12a and Cyp4a12b, both involved in the synthesis of hyocholic acid, also demonstrated a significant interaction between PFAS exposure and sex. The hepatic mRNA levels of Cyp4a12a increased by 2158.5-fold in PFAS-exposed females compared to vehicle (p<0.05) and increased by 1.77-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01). The hepatic mRNA levels of Cyp4a12b increased by 5897.8-fold in PFAS-exposed females compared to vehicle (p<0.05) and increased by 1.94-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01). mRNA levels of Cyp8b1 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.47-fold in PFAS-exposed males compared to vehicle (p<0.05), while females showed no significant differences (interaction; p<0.01).
We subsequently examined the following 9 genes related to bile acid transport (GO:0015721~bile acid and bile salt transport): Atp8b1, Slco1a4, Slc10a1, Slc10a2, Ephx1, Abcc2, Abcc3, Cyp2c70, and Abcb11 (Figure 2A and Figure 3). mRNA levels of Abcc2 increased by 2.18-fold change and Abcc3 increased by 3.85-fold change in the PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). Hepatic mRNA levels of Atp8b1, Slc10a2, and Abcb11 all decreased by 0.05-0.47-fold change in PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). mRNA levels of Slco1a4 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.05-fold in PFAS-exposed females compared to vehicle (p<0.05) and decreased by 0.12-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01). Hepatic mRNA levels levels of Slc10a1 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.15-fold in PFAS-exposed females compared to vehicle (p<0.05) and 0.08-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.05). Hepatic mRNA levels of Cyp2c70 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.26-fold in PFAS-exposed females compared to vehicle (p<0.05) and 0.16-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.05). mRNA levels of Ephx1 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels increased by 6.04-fold in PFAS-exposed females compared to vehicle (p<0.05) and 3.22-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01).
Figure 3. Effects of PFAS exposure and sex on the enterohepatic circulation.
This schematic depicts pathways related to cholesterol and bile acid metabolism and transport. Perturbations in these pathways due to PFAS exposure, or the interaction of PFAS exposure and gender, are indicated by color. Red: Significant PFAS induction. Blue: Significant PFAS reduction. Bolded Black*: Significant PFAS x Sex interaction. Black: No significant effect. 1Not measured. Changes in gene expression patterns analyzed by RNA sequencing are shown in italics. Significance is determined at p-values adjusted for multiple comparisons (q<0.05). Changes in metabolic signature are shown outlined in black. Genes and metabolites shown in the Gut portion of the schematic are inferred from hepatic results.
Another heatmap was generated using genes identified as part of the fatty acid biosynthetic, metabolic, or transport pathways (Figure 2B). Of the many genes included, 10 genes related to fatty acid beta-oxidation processes were analyzed (GO:0019395~fatty acid oxidation), as follows: Bdh2, Acacb, Ech1, Acoxl, Hsd17b4, Sesn2, Acox2, Cpt1b, Acad10, and Acaa1b (Figure 2B). Hepatic mRNA levels of Ech1, Acoxl, Sesn2, Cpt1b, Acad10, Hsd17b4, and Acaa1b all increased by 2.08-26.5-fold change in PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). The mRNA of Bdh2 decreased by 0.13-fold change in PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). mRNA levels of Acacb demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.44-fold in PFAS-exposed females compared to vehicle (p<0.05), while males showed no significant differences (interaction; p<0.05). mRNA levels of Acox2 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels increased by 2.99-fold in PFAS-exposed females compared to vehicle (p<0.05) and 1.52-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01).
Another 17 genes related to acylglycerol metabolic processes (GO:0046486~glycerolipid metabolic process) were analyzed, as follows: Dgkh, Fitm2, Mogat2, Lpin3, Acsl4, Pnpla1, Lpl, Tnxb, Gpld1, Slc27a5, Apoa4, Cps1, Pck1, Lpin1, Ddhd2, Dagla, and Esr1 (Figure 2B). Hepatic mRNA of Dgkh, Fitm2, Lpin3, Acsl4, Apoa4, Mogat2, and Ddhd2 were all increased by 2.23-29.2 fold change in PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). mRNA levels of Gpld1, Slc27a5, Cps1, Pck1, and Lpin1 were all decreased by 0.12-0.36-fold change in PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.01). Hepatic mRNA levels of Lpl demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels increased by 17.3-fold in PFAS-exposed females compared to vehicle (p<0.05) and 8.2-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01). mRNA levels of Esr1 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels decreased by 0.17-fold in PFAS-exposed females compared to vehicle (p<0.05) and 0.42-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.01) (Table 2 and Figure 2).
Real-time PCR was used for the validation of expression levels of several key genes: Cyp7a1, Cyp27a1, Slc10a1, Ephx1, and Abcc3 (Supplementary Figure 5).
3.5. Increases in circulating and hepatic relative PFAS levels after PFAS exposure and activation of PPARα target genes
Because of the increase in circulating cholesterol levels seen in both our male and female mice exposed to our PFAS mixture, and the effect sex had on the relative cholesterol levels, we next analyzed the relative hepatic and plasma PFAS levels. The metabolomic platform leveraged for our analyses included three of the five PFAS within the mixture. PFAS were observed in all mice, and interestingly, female controls displayed significantly lower PFAS compared to male controls (~0.3-fold) (p<0.05) (Table 3A). As expected, 12 weeks of PFAS exposure led to increased hepatic and circulating PFAS levels in mice, but there were distinct compound-specific differences. Hepatic levels of PFOS increased by 386.3-fold in PFAS-exposed females compared to vehicle (p<0.05) and 127.2-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.05) (Table 3A). Hepatic levels of PFOA increased by 4693.4-fold in PFAS-exposed females compared to vehicle (p<0.05) and 571.9-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.05). Hepatic levels of PFHxS increased by 2035.7-fold in PFAS-exposed females compared to vehicle (p<0.05) and 387.1-fold in PFAS-exposed males compared to vehicle (p<0.05) (overall exposure effect; p<0.05, interaction; 0.05<p<0.10). No significant correlations were observed between hepatic PFAS levels and circulating cholesterol levels (data not shown). As opposed to hepatic levels, circulating plasma levels of PFOS increased by 545.2-fold in PFAS-exposed females compared to vehicle (p<0.05) and 708.9-fold in PFAS-exposed males compared to vehicle (p<0.05) (overall exposure effect; p<0.05, interaction; 0.05<p<0.10) (Table 3A). Plasma levels of PFOA increased by 10883.6-fold in PFAS-exposed females compared to vehicle (p<0.05) and 6936.8-fold in PFAS-exposed males compared to vehicle (p<0.05) (overall exposure effect; p<0.05). Plasma levels of PFHxS increased by 6491-fold in PFAS-exposed females compared to vehicle (p<0.05) and 5414.2-fold in PFAS-exposed males compared to vehicle (p<0.05) (overall exposure effect; p<0.05, interaction; n.s.). Significant positive correlations were observed in the female PFAS-exposed mice between circulating cholesterol levels and plasma levels of PFOA and PFHxS (Supplementary Figure 3). Only plasma levels of PFOA displayed a positive significant correlation with circulating cholesterol levels in the male PFAS-exposed mice. Finally, in remaining available plasma samples we measured absolute concentrations of all 5 PFAS by electrospray ionization mass spectrometry (Table 3B). Although the exposure mixture was made up of equal concentrations of the 5 PFAS, circulating levels differed with PFNA>PFHxS>PFOS>PFOA>GenX. HFPO-DA (GenX) levels were a magnitude lower than the more persistent PFAS with observed concentrations that averaged ~100 ppb. As observed with the untargeted metabolomics, PFOA was significantly higher in exposed females and PFOS was significantly higher in exposed males.
Table 3. Gender specific changes in PFAS levels.
A. Liver and circulating plasma samples from n=7 mice from each treatment group was analyzed by a non-targeted metabolomics approach by Metabolon Inc. Three of the five PFAS included in the PFAS mixture were analyzed using LC-MS. Relative fold change between various treatment groups is given for hepatic and plasma levels of PFOS, PFOA, and PFHxS. Statistical significance for the interaction and main effects of PFAS exposure and gender was determined by two-way ANOVA analysis. B. Absolute plasma concentrations (ug/mL) of 5 PFAS included in the exposure mixture. Plasma was extracted from male and female mice (n=3 per group) exposed to PFOS, PFOA, PFNA, PFHxS, and GenX and levels were quantitated by electrospray ionization mass spectrometry. Red indicates a significant relative fold change increase (p<0.05); Blue indicates a significant relative fold change reduction (p<0.05).
| A. | |||||
|---|---|---|---|---|---|
| Relative PFAS Levels | |||||
| Biochemical Name | Fold Change Vehicle Female vs. Vehicle Male | Fold Change PFAS Female vs. PFAS Male | Fold Change PFAS vs. Vehicle | Overall PFAS Treatment q-value | Sex-PFAS Treatment Interaction q-value |
| Relative Hepatic PFAS Levels | |||||
| Perfluorooctanesulfonate (PFOS) | 0.30 | 0.92 | 188.03 | q<0.01 | q<0.01 |
| Perfluorooctanoate (PFOA) | 0.30 | 2.43 | 1513.94 | q<0.01 | q<0.01 |
| Perfluorohexanesulfonic acid (PFHxS) | 0.25 | 1.30 | 705.46 | q<0.01 | q=0.12 |
| Relative Plasma PFAS Levels | |||||
| Perfluorooctanesulfonate (PFOS) | 1.00 | 0.77 | 627.07 | q<0.01 | q=0.28 |
| Perfluorooctanoate (PFOA) | 1.25 | 1.96 | 10270.75 | q<0.01 | q=0.40 |
| Perfluorohexanesulfonic acid (PFHxS) | 0.80 | 0.96 | 6629.5 | q<0.01 | q=0.88 |
| B. | |||||
| Absolute Plasma PFAS Levels | |||||
| Biochemical Name | PFAS Female Average Concentration (ug/mL) | PFAS male Average Concentration (ug/mL) | Overall Average Concentration (ug/mL) | Fold Change PFAS Female vs. PFAS Male | p-value |
| PFOA | 23.4 +/− 9.0 | 7.3 +/− 1.8 | 15.4 +/− 10.5 | 3.2 | p=0.04 |
| L-PFOS | 13.8 +/− 2.3 | 22.5 +/− 4.1 | 18.1 +/− 5.6 | 0.6 | p=0.03 |
| L-PFHxS | 22.0 +/−8.3 | 25.6 +/− 5.8 | 23.8 +/−6.7 | 0.9 | p=0.57 |
| PFNA | 33.7 +/− 7.6 | 40.9 +/− 10.6 | 37.3 +/− 9.1 | 0.8 | p=0.40 |
| HFPO-DA (GenX) | 0.19 +/− 0.11 | 0.10 +/− 0.04 | 0.14 +/− 0.09 | 1.8 | p=0.27 |
As PFAS are known agonists of PPARα, our RNA sequencing analysis revealed significant alterations in the regulation of numerous PPARα target genes (Table 4). The mRNA of Fgf21 and Hmgcs2 were increased by 13.62- and 2.71-fold change, respectively, in PFAS-exposed mice compared to vehicle overall (overall exposure effect; p<0.05). The mRNA levels of Cd36 demonstrated a significant interaction between PFAS exposure and sex, with hepatic levels increased by 6.58-fold in PFAS-exposed females compared to vehicle (p<0.05) and 9.18-fold in PFAS-exposed males compared to vehicle (p<0.05) (interaction; p<0.02). In addition to being a PPARα target gene, Hmgcs2 is also the first enzyme in the cholesterol biosynthesis pathway from acetoacetyl-CoA. Although Hmgcs2 is upregulated by PFAS exposure, the subsequent cholesterol biosynthesis pathway enzymes, including Hmgcr, Mvk, Pmvk, and Mvd, were shown to be downregulated (Supplementary Table 4).
Table 4. PFAS exposure modulates hepatic PPARα target gene expression levels.
Liver samples from each treatment group were analyzed by RNA sequencing. Mice in each treatment group: female+vehicle (n=10), female+PFAS (n=9), male+vehicle (n=10), and male+PFAS (n=10). List of hepatic Pparα target genes and their fold change due to overall PFAS exposure. Significant p-values and significant Benjamini-Hochberg p-values are denoted in bold (p<0.05) for overall PFAS exposure and PFAS-sex interaction.
| Hepatic PPARα Target Gene | Overall PFAS Exposure Fold Change | Overall PFAS Exposure p-value | Overall PFAS Exposure Benjamini-Hochberg adjusted p-value | PFAS-Sex Interaction p-value | PFAS-Sex Interaction Benjamini-Hochberg-adjusted p-value |
|---|---|---|---|---|---|
| Pparα | 0.78 | p=0.002 | p=0.03 | p=0.64 | p=1.00 |
| Lipid/Hormone Transport | |||||
| Cd36 | 6.71 | p<0.001 | p=0.02 | p=0.001 | p=0.02 |
| Slc27a1 | 17.07 | p<0.001 | p<0.001 | p=0.41 | p=0.92 |
| Acyl-CoA Formation/Hydrolysis/Binding | |||||
| Acsl1 | 2.79 | p<0.001 | p<0.001 | p=0.50 | p=1.00 |
| Fabp1 | 2.18 | p<0.001 | p<0.001 | p=0.01 | p=0.09 |
| Mitochondrial-Oxidation/Oxidative Phosphorylation | |||||
| Acaa2 | 1.54 | p<0.001 | p<0.001 | p=0.09 | p=0.39 |
| Acadm | 3.78 | p<0.001 | p<0.001 | p=0.53 | p=1.00 |
| Cpt1a | 1.67 | p<0.001 | p<0.001 | p=0.48 | p=1.00 |
| Cpt2 | 3.11 | p<0.001 | p<0.001 | p=0.004 | p=0.04 |
| Slc25a20 | 3.96 | p<0.001 | p<0.001 | p=0.28 | p=0.74 |
| Ketogenesis/Ketolysis | |||||
| Fgf21 | 13.62 | p<0.001 | p<0.001 | p=0.65 | p=1.00 |
| Hmgcs2 | 2.71 | p<0.001 | p<0.001 | p=0.24 | p=0.70 |
| Peroxisomal-Oxidation | |||||
| Acox1 | 5.63 | p<0.001 | p<0.001 | p=0.02 | p=0.17 |
| Abcd3 | 2.49 | p<0.001 | p<0.001 | p=0.70 | p=1.00 |
| Acaa1a | 4.22 | p<0.001 | p<0.001 | p=0.36 | p=0.85 |
| Hsd17b4 | 3.57 | p<0.001 | p<0.001 | p=0.01 | p=0.11 |
| Lipogenesis | |||||
| Agpat2 | 0.84 | p=0.026 | p=0.18 | p<0.001 | p=0.001 |
| Fads1 | 0.75 | p=0.014 | p=0.11 | p<0.001 | p=0.013 |
| Srebf1 | 0.62 | p=0.007 | p=0.069 | p=0.39 | p=0.89 |
| Mogat1 | 49.12 | p<0.001 | p<0.001 | p=0.95 | p=1.00 |
| Lipases/Lipid Droplet Proteins | |||||
| Cidec | 197.9 | p<0.001 | p<0.001 | p<0.001 | p=0.004 |
| Pnpla2 | 5.65 | p<0.001 | p<0.001 | p=0.35 | p=0.85 |
| Lipoprotein Uptake and Metabolism | |||||
| Angptl4 | 1.95 | p<0.001 | p=0.002 | p=0.43 | p=0.94 |
| Apoa1 | 0.48 | p<0.001 | p<0.001 | p<0.001 | p=0.005 |
| Apoc3 | 0.26 | p<0.001 | p<0.001 | p=0.007 | p=0.07 |
| Lipc | 0.27 | p<0.001 | p<0.001 | p=0.003 | p=0.03 |
| Pctp | 1.86 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Vldlr | 4.57 | p<0.001 | p<0.001 | p<0.001 | p<0.001 |
| Cholesterol/Bile Transport and Metabolism | |||||
| Abca1 | 1.37 | p<0.001 | p<0.001 | p=0.09 | p=0.39 |
| Abcb4 | 2.49 | p<0.001 | p<0.001 | p=0.24 | p=0.69 |
| Cyp7a1 | 0.14 | p<0.001 | p<0.001 | p=0.01 | p=0.09 |
| Cyp27a1 | 0.56 | p<0.001 | p<0.001 | p=0.25 | p=0.70 |
| Nr1h3/Lxrα | 0.83 | p=0.008 | p=0.08 | p=0.41 | p=0.91 |
| Scarb2 | 1.29 | p<0.001 | p=0.003 | p=0.14 | p=0.52 |
| Slc10a2 | 0.05 | p<0.001 | p<0.001 | p=0.34 | p=0.84 |
Discussion
Epidemiological studies focusing on several model PFAS have found exposure to be associated with increased risk of numerous chronic diseases related to lipid and sterol metabolism, including cardiometabolic disorders and steatosis (Chen et al. 2020; Gaston et al. 2020; Jin et al. 2020; Wang et al. 2017). Studies across species have also reinforced the association between PFAS exposure and diseases such as steatosis and dyslipidemia. Numerous studies in rodents have found PFAS exposure to be associated with liver injury, including hypertrophy, apoptosis, and increased steatosis (Butenhoff et al. 2012; Han et al. 2018; Martin et al. 2007; Nakagawa et al. 2012; Qazi et al. 2010). However, in contrast to the hypercholesteremia demonstrated in human studies, PFAS exposure in animal studies, historically performed at high doses, have often produced decreased circulating cholesterol (i.e. hypocholesteremia) (Butenhoff et al. 2012; Han et al. 2018; Martin et al. 2007; Qazi et al. 2010; Seacat et al. 2002). Various reasons for these discrepancies have been offered, including interspecies genetic differences (Fragki et al. 2021). For example, one well-established PFAS signaling mechanism is through PPARα (Intrasuksri et al. 1998; Rosen et al. 2008; Takacs and Abbott 2007). A study conducted by Nakamura et al. revealed that human PPARα was much less responsive than mouse PPARα to PFAS exposure (Nakamura et al. 2009). This is further supported by findings that PPARα expression in humans is approximately 1/10 that expressed in mice (Palmer et al. 1998). However, a recent study using mice with liver-specific humanized PPARα demonstrated PFOA exposure to be positively associated with increased serum cholesterol levels in males (Schlezinger et al. 2020). It should also be noted, though, that PPARα agonists are widely used in humans clinically to produce cholesterol lowering effects (Staels et al. 1998). As PFAS are known activators of PPARα, the association of serum PFAS levels with increased serum cholesterol creates appears at odds with a PPARα-dependent mode of action. Studies investigating the in vitro activation of human PPARα by PFAS using reporter gene assays showed that PFOA, PFOS, PFHxS and GenX were all capable of PPARα activation. However, while PFOA and GenX demonstrated the strongest effect on PPARα activation, they were only able to produce 60-90% activation compared to the positive control GW7647 even at concentrations 250 times greater (Behr et al. 2020b). Therefore, although PFAS can activate PPARα, PPARα activation may have only a minor role in the observed PFAS-mediated adverse health effects. Several studies have also demonstrated that PPARα is not required for the induction of PFAS-mediated health effects in mice, including lipid accumulation with steatosis (Filgo et al. 2015; Wang et al. 2014).
For studies that focus on cardiometabolic disease, it is important to acknowledge that lipoprotein metabolism differs between species and may also contribute to the discrepancies between mouse and human data. For example, there is complete absence of cholesteryl ester transfer protein in rodents and a faster clearance of Apolipoprotein B from the liver in rodents compared to humans, resulting in a higher ratio of HDL-C to LDL-C in rodents (Princen et al. 2016). Additional explanations for the initial apparent discrepancies between animal and human PFAS studies include single chemical exposures versus real world complex mixtures, chronic versus acute exposures, and interactions with other stressors such as diet. The vast majority of the early rodent PFAS studies were conducted using a standard rodent chow diet, which contains low amounts of fat and negligible cholesterol. Thus, some more recent studies have investigated the effects of PFAS in conjunction with diets containing high amounts of fat and/or cholesterol. However, male mice exposed to PFOS or PFOA in conjunction with a high-fat diet also exhibited reduced serum cholesterol levels (Bijland et al. 2011; Pouwer et al. 2019; Wang et al. 2014). However, a study by Rebholz et al. found that mice fed a western diet and exposed to dietary PFOA resulted in elevated serum cholesterol levels in both male and female mice (Rebholz et al. 2016). While the majority of these studies suggest that dietary fat does not modify PFAS-induced serum cholesterol alterations, further studies are warranted to better understand the interactions of diet and PFAS toxicity.
However, the majority of mechanistic studies rely on single chemical exposures instead of utilizing PFAS mixtures when examining PFAS toxicity. Therefore, we wanted to investigate mechanisms linking exposure to a PFAS mixture, while on an atherogenic diet, with alterations in lipid metabolism. Here, we utilize unbiased transcriptomics and metabolomics to begin to characterize the toxicity of exposure to a PFAS mixture that incorporates the use of an atherogenic rodent diet and results in liver injury and increased circulating cholesterol even with concurrently strong PPARα activation. We examined the effect of a 12-week exposure to a mixture of five PFAS of differing chain length that have been identified in the blood of humans and that represent both legacy and emerging PFAS of concern (PFOA, PFOS, PFNA, PFHxS, and GenX) on liver injury and alterations in lipid metabolism in C57BL/6J wild-type male and female mice. PFAS caused liver injury and inflammation through mechanisms that likely involve increased lipid utilization and oxidative stress, especially in females. In addition, PFAS exposure led to higher levels of circulating cholesterol, bile acids and other measure of fatty acid metabolism implicating critical PFAS-mediated impacts on the enterohepatic circulation and energy metabolism. We believe this is a useful model to study mechanisms linking PFAS exposure to increased risk of cardiometabolic outcomes. It will be important to follow-up these studies using hyperlipidemic mouse models, such as LDL receptor-deficient mice, to determine the effectiveness of this model in helping to understand the progression of cardiometabolic disorders that result in atherosclerosis.
It is well established that PFAS exposure is associated with liver injury, both in humans and in animal models, often manifesting as increased steatosis and TAFLD (Butenhoff et al. 2012; Cave et al. 2016; Jain and Ducatman 2019a; Jin et al. 2020; Nakagawa et al. 2012; Schwingel et al. 2011). Our study reveals that, after 12 weeks of exposure, PFAS results in increased injury, which may be through a mechanism related to oxidative stress and that varies depending on sex. The PFAS-exposed mice exhibited increased plasma levels of ALT (Fig. 1H), increased liver weight (Fig. 1B), an enlargement of hepatocytes and lighter eosin staining, together illustrating a greater degree of hepatic injury and toxicity resulting from PFAS exposure. The genes Fabp4 and Fgf21, known to be associated with the progression of NAFLD, are also upregulated in our PFAS-exposed mice (Coilly et al. 2019; Rusli et al. 2016). Inflammation is also an important mediator of liver injury, and migration of inflammatory cells to the area of injury is an important indicator of level of injury (Del Campo et al. 2018). However, studies conducted over the past 20 years have shown that PFAS exposure generally produces a net anti-inflammatory effect, although the specific effect varies by exposure route, animal model, and dose (DeWitt et al. 2012). Additionally, the specific effect on inflammatory cytokine production can vary depending on the PFAS compound, dose, and the involvement of signaling through PPARα (DeWitt et al. 2012). Histological analysis of the mouse livers reveals that the PFAS-exposed mice exhibited a greater number of inflammatory foci, particularly portal and lobular, as well as increased migration of eosinophils to the liver, which are well known for activation of various proinflammatory and immunoregulatory pathways (Fig. 1C) (Marichal et al. 2017). In addition, the PFAS-exposed mice showed upregulated expression of several oxidative stress markers, including Sod2 and the Nrf2-target genes Gstm1 and Nqo1 (Ma 2013). Similar PFAS-induced oxidative stress has been demonstrated in recent studies, in which exposure to PFOS and PFHxS, in conjunction with a high-fat diet, led to increased hepatic expression of genes involved in oxidative stress (Pfohl et al. 2020). Oxidative stress is further exemplified through the increased hepatic lipid levels of 5-HETE in PFAS-exposed mice, which is an eicosanoid derived from arachidonic acid, shown to be upregulated during oxidative stress and that can be oxidized to 5-oxo-ETE, a potent eosinophil chemoattractant (Powell et al. 1995). The increased 5-HETE may at least in part, account for the increased inflammatory response especially in female mice. It is also possible that the increased 5-HETE in PFAS-exposed mice may be attributed to the elevated biotransformation of arachidonic acid, rather than its β-oxidation, especially given the reduced hepatic arachidoylcarnitine levels observed in female mice exposed to PFAS.
Another indicator of oxidative stress, specifically in the mitochondria, is the high levels of acylcarnitines and ketone bodies measured in the PFAS-exposed mice. PPAR agonists have been found to produce reactive oxygen species in a PPARα-dependent manner (Kim and Yang 2013). As known agonists of PPARα, PFAS have been shown to upregulate mechanisms of fatty acid beta-oxidation both in humans and mice, including higher levels of acylcarnitines related to β-oxidation (Chen et al. 2020; Lu et al. 2019; Wang et al. 2020). Fatty acid β-oxidation, the process by which fatty acids are broken down to generate energy, involves the addition of a CoA group to a fatty acid followed by binding of free carnitine to produce acylcarnitines. Acyl carnitines have an essential role in the carnitine shuttle, which aids in the transportation of fatty acids into the mitochondria for beta oxidation (Knottnerus et al. 2018). These previous studies showing PFAS-induced upregulation of fatty acid β-oxidation pathways supports our data, which demonstrates an increase in the relative hepatic levels acylcarnitines in the PFAS-exposed mice. Furthermore, circulating levels of ketone bodies, which are produced via hepatic fatty acid oxidation and circulate to other organs during gluconeogenesis, were increased in PFAS-exposed mice in both the liver and plasma, indicating that these mice experience a shift in energy metabolism that is not limited to the liver (Lee et al. 2016). However, it is possible another mechanism responsible for the high hepatic levels of acylcarnitines could be PFAS-induced oxidative stress in the mitochondria, leading to increased energy production via acylcarnitines and ketone bodies. Short-, medium-, and long-chain fatty acid metabolism in the liver is increased, while very long-chain fatty acid metabolism is decreased, suggesting a differential regulation of oxidation of long to very long-chain fatty acids when exposed to PFAS. Interestingly, increased levels of corticosterone were observed in both the male and female PFAS-exposed mice, which could also be a reason for the reduction in fat observed in the PFAS-exposed mice (Supplementary Table 1 and Supplementary Figure 1).
Epidemiological studies conducted over the past several decades have revealed that humans exposed to both high and lower “background” PFAS exposure levels in the general U.S. population are positively associated with increased cholesterol (Geiger et al. 2014; Mora et al. 2018; Nelson et al. 2010; Sakr et al. 2007; Steenland et al. 2009). Furthermore, recent human and experimental data report on the emerging role of altered lipid metabolism, especially impaired hepatic cholesterol homeostasis, in the pathogenesis of NASH and NAFLD (Bashiri et al. 2013; Musso et al. 2013; Perla et al. 2017). Hepatic accumulation of free cholesterol has been shown to promote pathogenesis severity of NAFLD and NASH (Caballero et al. 2009; Puri et al. 2007; Van Rooyen et al. 2011; Wouters et al. 2010). In addition, studies aiming to reduce the accumulation of hepatic free cholesterol suggest improved disease severity (Deushi et al. 2007; Park et al. 2010; Yoneda et al. 2010; Zheng et al. 2008). However, the mechanisms involved in PFAS-induced lipid alterations remain poorly understood. Our study revealed that mice exposed to a mixture of PFAS while on a low-fat atherogenic diet results in increased relative plasma cholesterol levels in both PFAS-exposed females and PFAS-exposed males compared to control at 12 weeks (Table 3). However, total cholesterol levels were also measured at 3 and 5 weeks of PFAS exposure (Supplementary Figure 1H&I). Between 3 and 5 weeks, total cholesterol levels in the control mice decreased from 130 mg/dl to 54 mg/dl, while the PFAS-exposed mice only decreased from 126 mg/dl to 97 mg/dl. It is commonly seen in studies utilizing hyperlipidemic diets, that there is an initial increase in cholesterol levels, followed by a decrease and normalization over time (Petriello et al. 2018). These data suggest that the relatively elevated cholesterol levels in the PFAS-exposed mice may result from PFAS exposure preventing a general decrease in cholesterol levels, such as those seen in the control mice. Thus, PFAS could instead be thought of as causing a disruption in cholesterol rebalancing, as opposed to stimulating a general surge in overall cholesterol levels. Additionally, in conjunction with the significantly reduced body fat observed in our PFAS-exposed mice, the changes observed in circulating cholesterol levels could indicate increased reverse transportation of cholesterol under PFAS exposure (Marques et al. 2018). Furthermore, given the lower hepatic bile acid levels in our PFAS-exposed mice, it is possible that the cholesterol to bile acid metabolism is impaired, and the increased Vldlr expression could indicate an increased secretion of cholesterol into the blood. Interestingly, the relative hepatic and plasma levels of 4-cholesten-3-one, a cholestanol precursor, were increased in both PFAS-exposed males and females (Salen et al. 1984). It is possible that, due to PFAS exposure in the presence of dietary cholesterol, that cholesterol is being metabolized to a greater extent through oxysterol production of cholestanol unrelated to bile acid production. Another cholesterol metabolism pathway involves sulfonation of cholesterol to cholesterol sulfate, which is the C3β sulfate ester derived from cholesterol and an endogenous steroid that has roles in stabilizing cell membranes, supporting platelet adhesion, and regulating signal transduction pathways involving blood clotting, fibrinolysis, and epidermal cell adhesion (Strott and Higashi 2003). Our data reveals that hepatic levels of cholesterol sulfate are decreased in PFAS-exposed mice, while plasma levels are unchanged (Table 1). Although the reason for the hepatic decreases are unknown, it could be related to the source of hepatic cholesterol in the mice. Another cholesterol metabolite, 7-alpha-hydroxy-3-oxo-4-cholestenoate (7-Hoca), is a monohydroxy intermediate in the synthesis of primary bile acids that is produced primarily in liver mitochondria or extrahepatically (Axelson et al. 1992; Meaney et al. 2003; Shoda et al. 1993). Our data reveals that hepatic levels of 7-Hoca were decreased in PFAS-exposed females, while plasma levels of 7-Hoca were increased in PFAS-exposed males. These trends mirror those generally seen in our mice for bile acids levels in the liver and plasma, and so may be due to similar mechanisms.
The main route of cholesterol metabolism takes place through its conversion to bile acids, which is the primary method of cholesterol turnover (Luo et al. 2020). Bile acids are synthesized in the liver, secreted into the bile, reabsorbed and modified in the intestines, and circulated back to the liver via the portal vein, making up what is known as the enterohepatic circulation (Gonzalez 2012). Cholesterol is oxidized through multiple pathways into primary bile acids, two of the major primary bile acids being cholic acid and chenodeoxycholic acid in both mice and humans (Russell 2003). These primary bile acids can then be conjugated with taurine or glycine and secreted into the bile. In the intestines, primary bile acids can be deconjugated and converted by bacterial action to secondary bile acids. Previous studies have demonstrated the ability of PFAS exposure to impair bile acid metabolism and transport (Fenton et al. 2021). A recent study in humans has also found several PFAS to be highly associated with metabolites from predominantly lipid pathways (Salihovic et al. 2019). In vitro experiments exposing human cells to various PFAS report strong alterations in the expression of numerous genes involved in cholesterol and bile acid synthesis, metabolism, and transport, notably the strong decrease in Cyp7a1 expression (Behr et al. 2020a; Rowan-Carroll et al. 2021). Studies in mice have suggest that PFAS exposure impairs fatty acid transportation and bile acid synthesis from cholesterol, leading to steatosis via pathways that are PPARα-independent (Das et al. 2017; Filgo et al. 2015; Wang et al. 2014). Our study reveals that mice exposed to the PFAS mixture for 12 weeks have decreased levels of most hepatic primary bile acids in both males and females (Table 1). At the same time, PFAS-exposed mice exhibit increased circulating levels of bile acids, including many secondary bile acids, which may indicate an altered gut microbiome (Ridlon et al. 2014). An important factor affecting these bile acid levels involves downregulation of the bile acid synthesis pathways. High circulating levels of bile acids trigger feedback inhibition of cholesterol synthesis, bile acid synthesis, and bile acid transport (Eloranta and Kullak-Ublick 2008; Slijepcevic et al. 2017). Bile acids are a direct ligand for the nuclear receptor FXR (Ding et al. 2015; Hanafi et al. 2018). When bile acid levels are high, they can bind to and activate FXR, which induces expression of genes such as SHP and FGF-19 in order to protect against bile acid toxicity (Panzitt and Wagner 2021; Slijepcevic et al. 2017). FXR is enriched in the liver and ileum to allow for sensing and regulating bile acids in the enterohepatic circulation (Panzitt and Wagner 2021). Bile acids can bind to and activate FXR, which induces expression of target genes such as the inhibitory factor SHP1 (Goodwin et al. 2000). SHP1 inhibits transcription of CYP7A1 and CYP8B1 induced by liver receptor homolog 1 (LRH-1) (Goodwin et al. 2000). In the intestines, activation of FXR on ileal enterocytes leads to upregulation of fibroblast growth factor (FGF) 15/19, which can also act on hepatocytes to repress bile acid synthesis (Kliewer and Mangelsdorf 2015). Our RNASeq data supports activation of these negative feedback pathways, as Cyp7a1 and Cyp27a1 are both downregulated due to PFAS exposure, as well as other members of their downstream bile acid metabolism pathways (Figure 3). These pathways are likely activated by the high levels of circulating bile acids. It has also been shown that treatment with FXR agonists can downregulate cholesterol synthesis genes in mice (Gardès et al. 2013). In Supplementary Table 4, PFAS downregulation of cholesterol biosynthesis gene expression are reported. Only Hmgcs2, a direct PPARα target gene, was not downregulated. PFAS are also known agonists of PPARα (Wang et al. 2020; Wolf et al. 2012), which is widely known downregulate cholesterol synthesis genes when activated (Knight et al. 2005; König et al. 2007; Li and Chiang 2009). It is possible that the initial exposure to PFAS leads to downregulation of endogenous cholesterol synthesis. However, although endogenous cholesterol synthesis is repressed, our data shows that relative hepatic cholesterol levels do not significantly change in the PFAS-exposed mice. In our model, mice always had access to cholesterol via their diet throughout their exposure to PFAS. Thus, these results may better reflect the adaptation to changes in cholesterol source from endogenous to dietary. However, although hepatic cholesterol levels did not change significantly with PFAS exposure, hepatic bile acid levels saw an overall reduction, indicating a reduction in the cholesterol to bile acid conversion. Further studies are needed to determine the contributions from different sources of cholesterol to total cholesterol levels.
The observed decrease in the hepatic levels of bile acids is also Supplementary by the decrease in hepatic expression of transporters responsible for bile acid uptake, including Na+-taurocholate cotransporting polypeptide (NTCP), OATPs, and mEH (Kosters and Karpen 2008). NTCP, encoded by the gene Slc10a1, is a high-affinity Na+-dependent transporter located solely on the basolateral membrane of hepatocytes and known to transport bile salts, sulfated steroid hormones, and thyroid hormones as well (Döring et al. 2012). NTCP accounts for ~90% of bile acid uptake and its major bile substrates include all the main taurine- and glycine-conjugated bile acids (Hagenbuch and Meier 1994; Hata et al. 2003; Platte et al. 1996), while unconjugated bile acids are only moderate substrates (Hata et al. 2003; Mita et al. 2006). Our data demonstrates a significant downregulation of Slc10a1 is PFAS-exposed mice of both sexes. Another major uptake transporter for bile acids into the liver is the organic anion transporting polypeptide 1 (OATP1). OATP1, encoded by the gene Slco1a1, is a Na+-independent transporter located on the basolateral membrane of hepatocytes and most efficiently transports unconjugated or sulfated bile acids (Hata et al. 2003; Meier et al. 1997). Our data demonstrates a significant downregulation of Slco1a1 in PFAS-exposed mice. A recent study examining the effects of PFAS administration in the presence of a high-fat diet reported a similar reduction in expression of hepatic PFAS uptake transporters, including OATPs (Pfohl et al. 2021). However, unlike the NTCP and OATP1 transporters, the microsomal epoxide hydrolase (mEH) expression was shown to be upregulated in PFAS-exposed mice. mEH, encoded by the gene Ephx1 and located on the basolateral membrane of hepatocytes (Von Dippe et al. 1993; von Dippe et al. 2003), was originally associated with the conversion of epoxides to trans-dihydrodiols (Oesch et al. 1971). Although much about the preferred substrates for mEH remain unknown, it has been reported that mEH can transport cholate and taurocholate (von Dippe et al. 1996). The increase in Ephx1 could be due to its transcriptional regulation by GATA-4 (Zhu et al. 2004) or the HNF-4α/CAR/RXR/PSF complex (Peng et al. 2013), as opposed to the FXR pathway that regulates Ntcp and Oatps (Peng et al. 2015). In addition, mEH is a bifunctional enzyme that can be induced by Nrf2 via the antioxidant response pathway and plays an important role in oxidative stress (Ma et al. 2006). It is possible that the increased expression of Ephx1 seen in our data is due to mEH’s role in detoxification and oxidative stress. It has also been previously shown that some of these bile acid transporters, such as NTCP, OATPs, and ASBT to transport PFAS as well as bile acids (Zhao et al. 2015; Zhao et al. 2017), which could compete for binding with bile acids and contribute to the decreased hepatic bile acid levels observed.
The low hepatic levels of bile acids and high plasma levels of bile acids can also be regulated by transporters responsible for bile acid efflux from the liver and bile acid influx to the intestines. The hepatic bile acid efflux transporters include the bile salt export pump (Bsep), encoded by the gene Abcb11. Bsep is an ATP-dependent transporter located on the hepatocyte canalicular membrane and is the major efflux transporter in the liver, efficiently transporting conjugated bile acids and driving the flow of bile (Byrne et al. 2002; Stieger et al. 2007). Previous reporter gene studies have demonstrated the ability of FXR to directly bind the promoter region and positively control transcription of Abcb11 (Ananthanarayanan et al. 2001; Matsubara et al. 2013; Plass et al. 2002). Studies involving human patients have revealed that Bsep deficiencies are associated with intrahepatic cholestasis (Stieger et al. 2007; Strautnieks et al. 1998). Lacking or having a defective form of the Bsep protein results in an impaired ability to release bile from liver cells, leading to a build-up of bile in the liver and diminished liver function (Sohail et al. 2021). Surprisingly, our data reveal that Abcb11 is downregulated ~0.3-fold in PFAS-exposed mice compared to vehicle mice. Therefore, the low hepatic bile acid levels in the PFAS-exposed mice may occur via a mechanism unrelated to Bsep efflux. It is also possible Bsep is being downregulated by inflammatory injury, as has been previously described (Wagner et al. 2010), or that Bsep is being directly regulated by PFAS by an unknown mechanism. In the intestine, bile acid reabsorption is largely controlled by the apical sodium dependent transporter (ASBT), which is located on the apical membrane of ileal enterocytes and encoded by the gene Slc10a2 (Balakrishnan and Polli 2006). Studies have shown FXR to negatively regulate ASBT expression in the intestines of mice (Chen et al. 2003; Matsubara et al. 2013). This is in agreement with our results, which show Slc10a2 to be downregulated and high levels of circulating bile acids. Patients with ASBT deficiency have been shown to display substantial bile acid malabsorption (Oelkers et al. 1997). However, it is possible that hepatic gene expression levels do not reflect intestinal expression levels. Further analysis examining the intestine specifically are required in the future. Although our metabolomics data show a decrease in the majority of secondary bile acids in PFAS-exposed mice, some specific exceptions do appear. For example, relative hepatic levels of taurohyocholic acid significantly increased in PFAS-exposed males compared to vehicle, while levels in female mice did not change with PFAS exposure, which could possibly be due to sex-dependent differences in hepatic uptake of secondary bile acids or differences in hepatic taurine content.
Throughout this study, sex has been a significant variable impacting the direction and degree of our results. For example, our study reveals that exposure to PFAS for 12 weeks results in increased injury, inflammation, and oxidative stress that varies depending on sex. Interestingly, the increase in liver weight and upregulation of TALFD-related gene Fgf21 was significantly increased in PFAS-exposed females compared to PFAS-exposed males, indicating a higher propensity for females to develop PFAS-induced hepatic toxicity (Fig. 1B&E). PFAS-exposed females also displayed greater lobular and portal inflammation (Fig. 1C). In addition, the upregulation of markers of oxidative stress Gstm1, Nqo1, and Sod2 by PFAS exposure, as well as hepatic levels of the metabolite 5-HETE, all demonstrated a significant interaction between PFAS exposure and sex, with greater expression in females (Fig.1 F&G). It is possible that the greater propensity for liver injury in PFAS-exposed females compared to males derives from the greater activation of these inflammatory and oxidative stress pathways in females. Furthermore, although the females did exhibit increasing water intake compared to males, only hepatic and plasma levels of PFOA were found to be increased in PFAS-exposed females compared to males (Supplementary Figure 1B and Table 3). A unique aspect of our work is the ability to see relative serum and hepatic PFAS levels in exposed male and female mice. Although there are examples of serum PFAS concentrations after PFAS exposure (Butenhoff et al. 2012; Loveless et al. 2006), sex-stratified direct liver-to-blood comparisons are not readily presented. It is possible that the sex-dependent progression of liver injury derives from toxicity induced specifically by PFOA, which was found at greater levels in both the plasma and livers of PFAS-exposed females (Table 3). Similar to our results, a recent study in which female and male mice exposed to PFOA and an American diet exhibited increased serum cholesterol in male mice only (Schlezinger et al. 2020). When examining human epidemiological data, in the general U.S. adult population, total cholesterol levels are higher in females than males, although this can largely be attributed to a higher HDL fraction in females than males and varied depending on age and race (Carroll et al. 2012; Jain and Ducatman 2019b; Swiger et al. 2014). This trend can be seen in data from NHANES surveys conducted from 1959-2008 ((CDC) 2009). In terms of age, younger women are largely protected from cardiovascular disease due to estrogens (Meyer and Barton 2016), while after menopause the risk factors for cardiovascular disease increase(Barrett-Connor 2013). In one study of primary care patients either at-risk for cardiovascular disease or having cardiovascular disease, women had higher total cholesterol and higher LDL-C than, but also received less lipid-lowering treatment (Rachamin et al. 2021). In a study of PFAS-exposed populations, it was found that men had higher total cholesterol than women, although the trend varied with age (Sakr et al. 2007).
Our data also reveals significant interaction between PFAS exposure and sex for transcriptional regulation of the bile acid influx transporters Slc10a1, Slco1a1, and Ephx1. Slc10a1 had greater fold change repression in PFAS-exposed males compared to females. Slco1a1, which is highly expressed in male vehicle mice and expressed at low levels in female vehicle mice, was almost completely repressed in both sexes. Lastly, Ephx1 was upregulated to a significantly greater degree in PFAS-exposed females compared males. Previous studies have also observed sex-specific differences in the expression and abundance of bile acid transporters, such as the male-predominant Oatp1a1 and Oatp1a3 in mouse liver (Cheng et al. 2005; Gotoh et al. 2002; Isern et al. 2001; Melià et al. 1998). These differences have been shown to be the result, at least in part, of the abundance of sex-specific hormones, such as androgen (Cheng et al. 2006). These effects contribute to the gender differences observed in bile acid synthesis and bile acid pool composition (Li-Hawkins et al. 2002; Phelps et al. 2019). For example, males have been shown to have a much larger bile acid pools than females (Bennion et al. 1978; Turley et al. 1998; Xiang et al. 2012). Furthermore, gender differences in bile acid transporters could be having an effect on hepatic PFAS levels. Our data reveals male vehicle mice have significantly higher levels of PFOA, PFOS, and PFHxS in the liver compared to female vehicle mice. This could be due to the greater male expression of transporters such as OATP1a1, which also contribute to PFAS uptake into the liver. Our data also showed greater relative concentrations of hepatic PFOA in PFAS-exposed females compared to PFAS-exposed males, which could be due to gender-specific transporter abundance along with the specific affinity of PFOA for certain transporters. It is not likely that these differences were due to the observed increased intake of water in PFAS-exposed females, as the other PFAS measured did not follow the same sexually dimorphic pattern. Certain limitations are present throughout this study. The dosing regimen used mirrors blood levels of high occupational exposure to single PFAS, such as those found in workers at PFOA manufacturing plants, where studies have found average serum PFOA concentrations of 1.3-12.9ug/mL (Costa et al. 2009; Sakr et al. 2007). The average serum PFOA concentrations in our mice was 7.3 ug/mL in males and 23.4 ug/mL in females (Table 3B). The concentrations of other PFAS used in our model are unknown in these studies. Using NHANES as a guide for the PFAS serum levels in the general population, it is known that our measured concentrations of PFOA were 2000-6000-fold higher than the general population (Nelson et al. 2010). While average serum GenX concentrations in our mice were 0.10 ug/mL in males and 0.19 ug/mL in females (Table 3B), previous studies measuring serum levels of GenX in a population exposure to PFAS-contaminated water found GenX not to be detected above their limit of detection (2ng/mL) (Kotlarz et al. 2020). Our study is not attempting to mimic water exposures, but rather blood concentrations similar to highly exposed individuals. Investigation of the effects of lower PFAS dosage regimens will be valuable in future studies. The critical next steps would be to utilize a mixture of PFAS that more closely matches the observed water toxicity levels and that reflect the measured concentrations found in human blood. The body weights, food intake, and water intake were recorded weekly until around week 5-6, when facilities were shut down due to COVID-19 for all except emergency personnel. Mice continued to receive treatment via their water source during this time, but detailed monitoring could not be performed and additional blood and tissue samples could not be collected until study completion. Because of this, it is unknown what the pattern of weight gain or weight loss looks like over time, especially the definitive time the PFAS-exposed mice began losing weight. Future studies that utilize the same diet but with exposures of lower concentration are critical to limit the impact of weight loss as an additional variable. Similarly, although the PFAS-exposed female mice displaying a trend in increased water intake, the trend beyond week 5 is unknown. In addition, after 12 weeks of exposure to PFAS, the female mice displayed signs of wasting that could by itself obscure the specific pathways being mediated by PFAS exposure. Furthermore, the estrogenic influences on PFAS toxicity were not examined in this study, but are an interesting variable for further study. Moreover, although our study focused largely on PFAS exposure’s effect on sterols, there are numerous different species in the lipid milieu that could provide the setting for follow-up studies in the future (Quehenberger and Dennis 2011). Our untargeted metabolomics approach did not include measurement for metabolites involved in the cholesterol to bile acid metabolism pathway, which would help to highlight specific PFAS-induced modulations of these pathways leading to the overall effects. Furthermore, levels of cholesterol subsets, such as HDL/LDL, were not measured. Additionally, although we exposed mice to 5 distinct PFAS chemicals, this should still be considered a simple mixture model. Future studies investigating full factorial combinatorial mixtures may help tease out synergistic effects of these PFAS pollutants.
Studies of the effects of PFAS on human health have and continue to demonstrate an association with increased risk of chronic diseases such as steatosis and cardiometabolic disorders. Here, we begin to characterize the toxicity of exposure to a PFAS mixture in combination with an atherogenic diet that results in a pattern of liver injury and increased circulating cholesterol similar to that seen in humans. In our model, exposure to the PFAS mixture resulted in accelerated liver injury, especially in females, that likely occur through mechanisms involving increased oxidative stress and altered lipid metabolism. Furthermore, mice exposed to the PFAS mixture displayed increased circulating cholesterol levels and alterations in levels of bile acid and fatty acid metabolism that indicate impaired functions of the enterohepatic circulation and energy metabolism. Our results suggest that the positive associations observed in epidemiological studies between PFAS and circulating cholesterol may be more related to impacts on cholesterol metabolism and transport than on endogenous cholesterol formation.
Supplementary Material
Highlights.
PFAS exposures are associated with dysregulated lipid and sterol homeostasis
A mixture of 5 PFAS increased cholesterol and bile acids in mice
Female mice displayed increased liver injury
Circulating PFOA were increased in exposed females compared to males.
Acknowledgments:
We thank Katherine Gurdziel at the Wayne State University Genomics Core for support related to RNA sequencing and generation of heatmaps.
Funding:
This research was supported in part by the National Institute of Environmental Health Sciences [P30ES020957, R00ES028734] and the National Institute of Diabetes and Digestive and Kidney [R01DK106540] at the National Institutes of Health and the Office of the Vice President for Research at Wayne State University. The content is solely the responsibility of the authors and does not necessarily represent the official views of the NIH.
Footnotes
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The authors declare that there are no competing financial interests.
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