Toxicity assessment of exposure to four short- and ultra-short PFAS using Daphnia magna and Hydra vulgaris as models Houda Hanana a, Marie Houédé b, Geneviève Farley c, Pascale Bouchard c, Roxane Sorel c, Quoc Tuc Dinh d, Sébastien Sauvé d, Magali Houde a* a Aquatic Contaminants Research Division, Environment and Climate Change Canada, 105 McGill Street, 7th Floor, Montreal, QC, H2Y 2E7, Canada. b University of Angers, Angers, France c Quebec Laboratory for Environmental Testing, Environment and Climate Change Canada, Montreal, QC, H2Y 2E7, Canada. d Université de Montréal, Department of Chemistry, C.P. 6128, Succursale Centre-Ville, Montreal, QC, H3C 3J7, Canada. Running title: Toxicity assessment of short- and ultra-short PFAS in aquatic invertebrates *Corresponding author: Dr Magali Houde Environment and Climate Change Canada 105 McGill Street, 7th Floor, Montreal, QC, H2Y 2E7, Canada. Email: magali.houde@ec.gc.ca mailto:magali.houde@ec.gc.ca 4 Abstract 1 Short- and ultra-short per- and polyfluoroalkyl substances (PFAS) such as perfluorobutanoic acid 2 (PFBA, 4C), perfluorobutane sulfonic acid (PFBS, 4C), trifluoroacetic acid (TFA, 2C) and 3 trifluoromethane sulfonic acid (TFMS, 1C) were detected in various environmental matrices but 4 studies addressing the effects attributed to exposure are still scarce. This study aimed to (1) 5 investigate acute toxicity of these emerging PFAS using Daphnia magna (immobility, mortality, 6 reproduction, and body size) and Hydra vulgaris (morphological changes and reproduction) and 7 (2) assess the sub-chronic/chronic toxicity of TFA in both species. Glutathione S-transferase 8 (GST) activity was also assessed in both species after long-term exposure to TFA. Chemical 9 analysis confirmed presence and stability of 4 PFAS in exposure media. For all tested PFAS, 10 endpoints examined after acute tests were not markedly affected by exposure and LC50 values were 11 > 1000 mg/L. However, this toxicity was significantly affected by media acidification induced by 12 the tested chemicals (LC50 = 316 and 31,6 mg/L for D. magna and H. vulgaris, respectively). Long-13 term exposure to TFA did not significantly induce any effect on both species and GST levels were 14 not altered. Overall, results suggest lower toxicity of ultra/short-chain PFAS to these species under 15 selected exposure conditions. However, additional studies investigating multigenerational effects 16 of these PFAS using realistic environmental concentrations are needed to overcome the significant 17 gaps in our understanding of short /ultra-short PFAS-induced toxicity. 18 19 20 21 22 5 Introduction 23 Per- and polyfluoroalkyl substances (PFAS) form a large and heterogenous group of over 9000 24 fluorinated chemicals (Shittu et al., 2023). Owing to their oleophobicity, hydrophobicity, heat 25 resistance, and persistence (Buck et al., 2011; Shittu et al., 2023), PFAS have been widely used 26 since the 1940s in a variety of industrial and commercial applications including aerospace 27 technologies, firefighting foam, and consumer products (Brennan et al., 2021) including clothes, 28 outdoor textiles, carpets (Kotthoff et al., 2015) cosmetics, and insecticides (Valsecchi et al., 2017). 29 A broad range of these substances were detected in various environmental and biotic samples 30 worldwide (Ma et al., 2022; Sharp et al., 2021). Several investigations and reviews reported 31 toxicity of PFAS in algae, aquatic invertebrates (crustaceans, bivalves, gastropods, 32 platyhelminthes, annelids) and vertebrates (amphibia, fish) (Ankley et al., 2021; Ma et al., 2022; 33 2023; Valsecchi et al., 2017). In addition, investigators noted that the toxicity of carboxylic acids 34 was correlated to the carbon chain length of these chemicals and that sulfonic acids exhibit more 35 toxicity than carboxylic acids for the same carbon chain length (Buhrke et al., 2013; Gaballah et 36 al., 2020; Ulhaq et al., 2013; Zheng et al., 2012). A few PFAS (PFOS; PFOA; perfluorohexane 37 sulfonate, PFHxS) and their respective salts and related compounds were listed under the 38 Stockholm Convention as Persistent Organic Pollutants (POPs) (UNEP, 2019). Long-chain 39 perfluorocarboxylic acids with carbon chain lengths from 9 to 21 carbons, their salts and related 40 compounds are currently under review by the Stockholm Convention (Stockholm Convention, 41 pops.int). 42 The regulation and phase-out of longer chain PFAS resulted in a significant shift in the 43 industry towards fluorinated alternatives such as short- (C = 4-7) and ultra-short (C = 2-3) chain 44 PFAS (Ateia et al., 2019; Xing et al., 2023), which are presumed to display less propensity to 45 6 bioaccumulate (Conder et al., 2008; Martin et al., 2003; Olsen et al., 2009) and induce toxic effects 46 (Wang et al., 2013). However, it has been found recently that physicochemical properties and 47 environmental fate of several substitutes were similar to long chain molecules (Gomis et al., 2015; 48 Militao et al., 2021). Short PFAS were measured and detected in surface water of rivers and 49 estuaries (Allinson et al., 2019; Zhao et al., 2016), groundwater (Gao et al., 2019), tap water (Gao 50 et al., 2019; Schwanz et al., 2016), Arctic environment (sediment and lake and ocean water) (Stock 51 et al., 2007; Yamashita et al., 2008) and wildlife (Gebbink et al., 2016). Due to their high solubility 52 in water and low/moderate sorption to sediments, short-chain homologs constitute a concern for 53 aquatic environments (Ateia et al., 2019). 54 Among dominant short PFAS, 4 carbon chain molecules such as perfluorobutanoic acid 55 (PFBA) and perfluorobutane sulfonic acid (PFBS), were widely detected in the environment (see 56 Table S1 for a compilation of studies). However, few studies investigated PFAS-initiated toxicity 57 on aquatic invertebrates, and some chronic testing was reported for these contaminants. Kadlec et 58 al. (2024) found no marked effects after short-term exposure of Hyalella azteca and Chironomus 59 dilutus to PFBA and PFBS and after chronic exposure of Ceriodaphnia dubia to these compounds. 60 However, sublethal exposure to PFBA led to metabolic perturbations in Daphnia magna via 61 disruption of the arginine and proline metabolism pathways (Labine et al., 2023). In fish such as 62 Oryzias melastigma and Danio rerio, toxicological studies indicated that PFBS exposure impaired 63 reproduction, embryonic development, pancreatic organogenesis, and energy homeostasis (Chen 64 et al., 2019; Sant et al., 2018). In addition, Hagenaars et al. (2011) noted that exposure to PFBS, 65 but not PFBA, induced malformations of the zebrafish embryos head, and only 3000 mg/L PFBS 66 resulting in alterations of embryo heart rates after 48hr and 72hr exposure. 67 7 Ivantsova et al. (2024) suggested that several biological responses are involved in the toxicity 68 attributed to exposure to PFBS and PFBA in different systems (cardiac, immune, hepatic, and 69 reproductive) of fish and indicated that understanding of the mechanisms underlying these effects 70 is limited by non-relevance of tested concentrations and lack of published reported studies. 71 To date, ecotoxicology research focused on a few PFAS and little was noted on ultra-short-chain 72 (C1-C3) PFAS (Björnsdotter et al., 2019). Trifluoroacetic acid (TFA, C2), the most studied 73 amongst the ultra-short chains (Björnsdotter et al., 2019), may be formed via degradation of 74 hydrochlorofluorocarbons (HCFCs), hydrofluorocarbons (HFCs), and hydrofluoroolifines (HFOs) 75 (Sidebottom and Franklin, 1996). Several chemicals frequently employed in pharmaceutical and 76 agricultural applications might also act as TFA precursors and contribute to its environmental input 77 (Freeling and Björnsdotter, 2023). In addition to TFA, trifluoromethanesulfonic acid (TFMS, C1) 78 was found by Zhou et al (2020) to be a POPs and mobile organic chemical which was detected at 79 concentrations up to μg/L in water samples such as urban effluent, surface water, ground water 80 and drinking water) collected from European countries (Montes et al., 2017; Zahn et al., 2016). 81 Despite global distribution and presence in many water samples including rivers, lakes, canals, 82 reservoirs, estuaries, tributaries, drainage basins, snowmelt, and precipitation, only two studies 83 determined the effects of TFA in aquatic species (Berends et al., 1999; Boutonnet et al., 1999). 84 Understanding the potential toxicity of these substances is therefore important to better support 85 their identification, evaluation and management. 86 In this context, it has been reported that PFAS interact with antioxidant proteins such as 87 glutathione-S-transferase activity, involved in antioxidant defense system (Rajak et al., 2024; 88 Hamid et al., 2024), and in detoxification processes via conjugating glutathione (GSH) to a variety 89 of hydrophobic and electrophilic compounds facilitating their solubilization and excretion (Hamid 90 8 et al., 2024). It was also reported that the mRNA level of GST was reduced in RTL-W1 exposed 91 to perfluoroethylcyclohexane sulphonate (PFECHS) at environmental concentrations (Mahoney et 92 al., 2023) and positively correlated to PFBA concentrations in coelomic fluid of sea cucumbers 93 (Cocci et al., 2025). Due to the marked increase in GST activity in the digestive gland of M. 94 galloprovincialis exposed in vitro to PFBS, Cunha et al. (2025) suggested that there is an enhanced 95 cellular response aiming to detoxify this pollutant or its potentially reactive metabolites. 96 The objectives of the present study were to investigate the acute and chronic/sub-chronic 97 toxicity of 4 short- and ultra-short-chain PFAS (TFA, TFMS, PFBA, and PFBS) utilizing two 98 aquatic invertebrates (Daphnia magna and Hydra vulgaris). These relevant model species are 99 widely distributed in the natural environment, easy to maintain and manipulate in the lab, possess 100 short life cycles, rapid reproduction, and are sensitive to a broad spectrum of contaminants 101 (Ahmed, 2023; Kovačević et al., 2024). In addition to mortality and physiological endpoints, GST 102 activity was also determined in both species as its regulation has been considered as a proxy for 103 cellular antioxidant and xenobiotic defense mechanisms. We hypothesize that although those short 104 PFAS are presumed to display less propensity to bioaccumulate than long chain PFAS, they 105 potentially have the capacity to elicit oxidative or electrophilic stress as suggested by Cunha et al. 106 (2025). To our knowledge, this study is the first to focus on the biological effects of short/ultra-107 short-chain PFAS, particularly for TFA and TFMS in aquatic invertebrates. 108 Material and methods 109 Chemical reagents and stock solutions 110 PFBA (heptafluorobutyric acid, purity 98%), PFBS (nonafluorobutane-1-sulfonic acid, purity 111 97%) and TFA (trifluoroacetic acid, purity 99%) were purchased from Sigma Aldrich (Oakville, 112 9 ON, Canada), and TFMS (trifluoromethane sulfonic acid, 95%) was purchased from Toronto 113 Research Chemicals (Toronto, ON, Canada) (Table S2). Stock solutions (1 g/L) were prepared 114 prior to exposure by dissolving different PFAS in the corresponding exposure media and stored 115 refrigerated (4°C) until testing. Since all tested PFAS markedly acidified the stock solutions (pH 116 less than 3), two acute tests with and without adjusting the pH of the stock solutions with sodium 117 hydroxide were carried out for each species. The final pH of the adjusted stock solutions was 118 similar to the pH of daphnid and hydra culture media. In addition, as evidence indicated that pH 119 of working solution is a key factor in accurately interpreting results (Wasel et al, 2021), physico-120 chemical properties (dissolved oxygen, conductivity, temperature and pH) of exposure media were 121 measured at the beginning and at the end of exposure (except conductivity) using standard methods 122 (Environment Canada, 1990; Environment and Climate Change Canada, 2020), in at least one 123 replicate per concentration (see Table S3 and S4). Nominal concentrations for D. magna (0.01, 1, 124 10, 100, and 1000 mg/L) and H. vulgaris (0, 0.0001, 0.001, 0.01, 1, 10, 100, and 1000 mg/L) 125 testing were selected based upon results from previous PFAS studies using congeners of similar 126 structures with aquatic species (Barmentlo et al., 2015; Jeong et al., 2016; Wasel et al., 2021). The 127 concentrations were selected to include environmentally relevant levels and higher doses to 128 determine LC50 and observe long-term effects. 129 Test organisms, quality assurance and quality control 130 Genetically homogenous Daphnia magna obtained from the Quebec Laboratory for 131 Environmental Testing of Environment Canada (Qc, Canada) were cultured in growth chamber 132 following Environment Canada’s method (Environment Canada, 1990). Cultures were kept at 20 133 ± 2 °C with a 16:8hr light:dark photoperiod in beakers in medium composed of reconstituted water, 134 0.02 g/L B12 vitamins (100 µL/L), at pH between 6.5 and 8.5. The daphnids were fed daily with 135 10 the green algae Raphidocelis subcapitata (3.5×105 cells/ml) and YCT preparation (0.0125 g/L of 136 yeast-cerophyll-trout chow) and the medium was renewed three times per week. Key water 137 parameters, such as pH and dissolved oxygen, were monitored weekly during routine water 138 changes. Neonates were used for the different experiments if the following criteria were met: the 139 % mortality of adult daphnia ≤ 25% during the 7day preceding exposures, first hatching ≤ 12 day 140 and mean of neonates /adults ≥ 15 (Environment Canada 1990). In order to confirm the sensitivity 141 of the tested organisms and suitability of experimental conditions in detecting potential toxicity, a 142 reference toxicant test was conducted using NaCl. The determined 48hr-CL50 (5.37 g/L; 95% 143 confidence interval (CI) = 4.98 - 5.8) was within the warning limit of the chart established as 144 reported in Environment Canada’s procedure method (Environment Canada 1990). 145 Hydra vulgaris used in this study were obtained from the Quebec Laboratory for 146 Environmental Testing of Environment Canada (Qc, Canada). The polyps were cultured in covered 147 crystallization bowels under standard methods (Environment and Climate Change Canada, 2020). 148 Hydra were reared in a medium composed of 0.15g CaCl2, 2 H2O, and 0.1 g N-tris hydroxymethyl 149 1-2- aminoethanesulfonic acid buffer at pH 7. Polyps were fed daily with iodine-disinfected brine 150 shrimps, Artemia salina, hatched within 24hr and medium was renewed approximately every 3-151 5hr after hydra feeding. Key water parameters, such as pH and dissolved oxygen, were monitored 152 weekly during routine media changes. A reference test was also conducted using ZnSO4 7H2O and 153 EC/LC50 (0.22 mg/L; 95% CI = 0.16-0.3 and 0.81 mg/L; 95% CI = 0.64-1.02, respectively) which 154 was within control limits as reported in Environment Canada’s procedure method (Environnement 155 et Changement Climatique Canada, 2025). 156 Daphnia magna: acute and sub-chronic testing 157 11 For acute and sub-chronic exposures, neonates were maintained under 20 ± 1 °C with a 16:8hr 158 light: dark. Acute toxicity (48hr) was tested using two replicates of 5 neonates (< 24hr) for each 159 experiment with and without pH adjustment following Environment Canada’s standard procedure 160 (Environment Canada, 1990). Neonates obtained from adult daphnia (20-28 days old) after 161 approximately 9-12 days were transferred into 100 ml polypropylene beakers filled with 75 ml 162 culture medium alone (control) or containing increasing concentrations of tests solutions (0.01, 1, 163 10, 100, or 1000 mg/L). The exposure concentrations were obtained by serial dilution in 164 reconstituted water from the highest concentration made from the stock solution of 1g/L (with or 165 without adjusted pH). pH was adjusted to 8 ± 0.5 using 1N NaOH. Endpoints of immobilization 166 and death were monitored at the end of two acute exposures. An organism was considered 167 immobile when no swimming could be observed within 15sec after stirring and a cardiac activity 168 could be recorded. 169 For sub-chronic exposure, 4 replicates of 12 neonates (< 24hr) were exposed for 14 days. 170 Currently, no standardized sub-chronic exposure tests exist for D. magna. Therefore, this test is a 171 compromise between the acute (Environment Canada, 1990) and chronic bioassays from the 172 OECD guideline (OECD, 2012). The sublethal doses (0.01, 1, 10, or 100 mg/L) of TFA were 173 selected based on the LC50 (Table S5) determined after 48hr acute exposure. Untreated D. magna 174 exposed only to the reconstituted water were used as control. The tested solutions were prepared 175 similarly to those for acute toxicity. Organisms kept in culture media were used as a control group. 176 Daphnids were fed daily with algae and YCT and media renewed every 48hr. Temperature, pH, 177 and dissolved oxygen were monitored at 48-hr intervals before and after each media change. 178 Endpoints of immobility and mortality were monitored every day. The number of offspring was 179 also recorded daily and were removed from the vessel. At the end of the exposure, body length, 180 12 defined as the distance from the top of daphnia head to the base of the tail spine, was determined 181 for a subset of three surviving organisms in each treatment and measured under a digital image 182 analyzing system (Leica M165c Stereomicroscope, Wetzlar, Germany). The remaining organisms 183 were collected and stored at -80 °C for biomarker analyses. 184 Hydra vulgaris: acute and chronic testing 185 Acute and chronic exposures were carried out in multi-well microplates covered with parafilm to 186 prevent evaporation and kept in an incubator at 20 °C with 16:8hr light-dark cycle during the 187 exposure time. The selection of healthy (morphology score of 10 based upon the Wilby scale) 188 (Wilby, 1988) non-budding hydra polyps of similar size was done randomly using a microscope. 189 As for D. magna, acute toxicity (96hr) of PFBA, PFBS, TFA, and TFMS was assessed with and 190 without adjusting pH (Environnement et Changement Climatique Canada, 2025). Exposures were 191 conducted, without feeding, in a 12-well microplate (3 individuals per well in triplicate; N = 9) 192 and 3 ml/well of test solutions (0.0001, 0.001, 0.01, 1, 10, 100, or 1000 mg/L) or hydra medium 193 only for the control groups were added. Tested concentrations up to 100 mg/L were obtained by 194 serial dilution in hydra medium from the highest concentration (1000 mg/L) used as a stock 195 solution. pH was adjusted at 7 ± 0.5 using 1N NaOH. 196 For chronic exposure (7 days), only sub-lethal doses of TFA (0.0001, 0.001, 0.01, 1 or 10 mg/L) 197 selected based upon LC50 results (Table S5) along with untreated control hydra exposed to hydra 198 medium only were used to assess chronic effects of this ultra-short PFAS without adjusting pH. 199 For each concentration, 6 non-budding polyps were randomly placed in a well of 6 multi-well 200 microplates containing 8 ml test solutions (4 replicates for each concentration). The exposure was 201 performed in triplicate (N=72 polyps) and polyps were fed every 3 days, 2hr before media renewal. 202 At the end of the acute and chronic exposure, the morphological changes of the individual polyps 203 13 were observed by microscopic examination. Five classical morphological stages were examined 204 based upon the Wilby scale (Wilby , 1988): healthy polyps were normal and exhibit an extended 205 body and tentacles (score 10); first sign of intoxication was characterized by presence of clubbed 206 tentacles (Score 8), late phase of intoxication was marked by the presence of shortened tentacles 207 and body (Score 6); tulip stage was characterized by polyps exhibiting markedly shortened 208 tentacles and body (score 5) and animal death was marked by polyps disintegration (score 0). 209 Scores ≥ 6 are reversible and sub-lethal, while scores <6 are irreversible and lethal. 210 The population growth and reproduction rates of H. vulgaris were also determined during 211 chronic testing. Following the 7day exposure period, the mean population growth rate (K) and 212 budding rates (Ktot) were determined following the methodology outlined by Holdway et al. 213 (2001). The value K reflects the growth rate based solely upon the count of detached buds and is 214 calculated as follows: 𝐾 = 𝑙𝑛 (𝑛𝑡7)−ln(𝑛𝑡0) 7 where nt7 is the total number of hydras produced during 215 the assay (including detached buds and initial polyps). Similarly, the budding rate value Ktot, which 216 accounts for both attached buds and new polyps, was calculated as 𝐾𝑡𝑜𝑡 = ln(𝑛𝑡𝑜𝑡)−ln(𝑛𝑡0) 7 , where 217 ntot represents the sum of the number of attached (forming) buds and the number of detached polyps 218 as indicator of reproduction. 219 GST activity 220 To better understand the potential sublethal effect of TFA, glutathione-S-transferase (GST) 221 activity, a marker of oxidative stress, was quantified in both species after 14 and 7d for daphnia 222 and hydra, respectively. Daphnids (4-5 organisms; N = 8) and hydra (pooled from 2 wells / 223 concentration; N = 6) were homogenized using buffer solution consisting of 100 mM NaCl, 25mM 224 HEPES-NaOH, 0.1mM dithiothreitol and 1 µg/L aprotinin at pH 7.4. The homogenates were 225 14 centrifuged at 15,000 g for 20 min at 4 °C. GST activity was determined according to the method 226 described by Gagné (2014). The reaction mixture consisted of 10 mM HEPES pH 6.5, 1mM GSH, 227 1mM 1-chloro-2-4-dinitrobenzene, 125 mM NaCl. The conjugation of glutathione (GSH) to 1-228 chloro-2, 4-dinitrobenzenem was evaluated by monitoring the increase in absorbance at 340 nm, 229 using the Synergy 4 microplate reader (BioTek, Winooski, VT, USA). Data were expressed as 230 nmole GSH/min/mg protein. Protein content was measured according to the method of Bradford 231 (1976) using a standard solution of bovine serum albumin for calibration. 232 Chemical analysis 233 Concentrations of the three PFAS were measured in media at 0 and 48hr for D. magna and at 0 234 and 96hr for H. vulgaris as exposure media was changed at these time points. Samples were 235 collected in 15 ml falcon tubes previously rinsed three times with methanol at 0h and 48h for D. 236 magna and 0 h and 96h for H. vulgaris (0 to 1000 mg/L) and stored at -20 °C until the analyses. 237 All TFA solutions (n = 24) and the non-adjusted pH media for PFBA, PFBS and TFMS were 238 analyzed (N = 12 for each compound). Samples were filtered (glass fiber, 0.3 μm, 25 mm) and 239 analyzed by on-line solid-phase extraction (SPE) coupled to hydrophilic interaction liquid 240 chromatography (HILIC) interfaced to high-resolution mass spectrometry (HRMS) (A Q-Exactive 241 Orbitrap - Thermo Scientific, Waltham, MA, USA). The on-line SPE and its connection to the 242 HILIC-MS system were performed by a dual switching-column array consisting of 6-port and 10-243 port valves. In the first steps, the sample was injected into 1 ml injection loop and loaded onto the 244 on-line SPE column (for sample pre-concentration. After sample loading, the on-line SPE mobile 245 phase was left to flow to remove the matrix/salt. The samples were then eluted at 0.5 ml/min with 246 HILIC-HRMS mobile phase, the analytes retention was conducted on a Thermo Scientific 247 Acclaim™ Trinity Q1 LC column thermostated at 35 °C, starting from strong organic conditions 248 15 (90/10 A/B). The HILIC mobile phases were 25 mM ammonium acetate (A) and HPLC-water (B). 249 Details the on-line SPE HILIC-HRMS acquisition method and chromatograms of the four studied 250 PFAS analyzed are provided in Tables S6 and Figure 2S. Limits of detection for PFBA, PFBS, 251 TFA and TFMS were 0.01, 0.01, 0.02, and 0.05 µg/L, respectively. 252 Statistical analysis 253 Statistical analyses were conducted using STATISTICA software package (version 7, France). All 254 data were tested for normality using Shapiro Wilks test and homogeneity using Bartlett's test. All 255 values were expressed as mean ± SD and were compared by using one-way ANOVA. When the 256 data did not follow a normal distribution, the non-parametric method Kruskal-Wallis test was used 257 to compare the difference between treatments and control of D. magna and H. vulgaris. The 258 multiple comparison test was carried out using Tukey's post hoc test. Significance was set at p ≤ 259 0.05. CETIS software version 1.9.7.10. (Tidepool Scientific Software, McKinleyville, CA, USA) 260 was used to analyze lethal and sublethal ecotoxicity data. The 50% lethal concentration (LC50) and 261 the sublethal effects (EC50) endpoints for acute exposures were determined using the binomial 262 method (Stephan, 1977). 263 Results 264 Acute and chronic toxicity in D. magna 265 Data demonstrated that concentrations of 4 tested PFAS determined in the D. magna and H. 266 vulgaris media at the beginning and the end of the exposure were generally close to nominal 267 concentrations (Table S7). Variation between the analyzed and the expected concentrations in the 268 media may be initiated by (1) differences in the standards used, (2) carryover cross-contamination 269 for low concentrations, (3) manipulations made during dilutions for high concentrations and (4) 270 16 method uncertainties. In addition, results obtained for TFA showed that measured concentrations 271 after each 48hr during sub-chronic exposure deviated slightly from the nominal values (for 272 example, accuracy was of 95 ± 10% (± SD, n = 6) and 81 ± 20% (± SD, n = 6) for TFA at 100 273 mg/L at the beginning and the end of the exposure, respectively) (Table S8), in agreement with an 274 observation reported by Xie et al. (2025) during their testing with PFBS. 275 Exposure of neonates to differing PFAS after adjusting the pH did not induce any marked 276 effects (mortality, immobility) in D. magna. Only the higher exposure concentration (1000 mg/L) 277 with non-adjusted pH led to 100% mortality for all tested substances (the pH was then below 4 278 and probably the main reason for the toxic effects). The EC50 and LC50 were estimated at 316 mg/L 279 for all PFAS, while these values were estimated to be higher than 1000 mg/L with adjusted pH 280 (Table S5). Similar to the acute testing, TFA 14day chronic exposure did not produce increase 281 mortality rates in D. magna (Figure 1A). No marked alteration in reproduction was found, and no 282 significant difference was observed in number of neonates between controls and TFA treatments 283 (Figure 1B). Daphnids from two replicates of the 10 mg/L treatment had their first neonates at day 284 9, all other treatments and controls reproduced at day 10. The average cumulative reproductive 285 output (number of neonates per female) after 14days for control (Figure 1C) was noted at 21 ± 2 286 and for tested concentrations this value varied from 20 ± 2 to 24 ± 1. Body measurements indicated 287 similar body sizes between exposed organisms and controls at the end of exposure (Figure 1C). 288 Analysis of GST activity in daphnids treated chronically with TFA did not exhibit any significant 289 variation in its levels compared to control (Figure 2A). 290 Acute and chronic toxicity in H. vulgaris 291 Morphological alteration after acute (96hr) exposures to PFBA, PFBS, TFA, and TFMS 292 indicated no increased mortality rates or sub-lethal effects for any treatment with adjusted pH by 293 17 comparison to controls (Table S5). However, in a non-adjusted pH media, no mortality was found 294 for the 4 PFAS tested acutely up to concentrations of 10 mg/L (all hydra were normal), while the 295 two highest concentrations (100 and 1000 mg/L) resulted in 100% mortality (hydra were at tulip 296 and degenerated stages and were unable to recover) by comparison to controls (Figure 3A). The 297 LC50 for PFBA, PFBS, TFA, and TFMS with non-adjusted pH after 96hr were determined to be 298 31.6 mg/L but exceeded 1000 mg/L when the pH was adjusted (Table S5); no significant 299 differences were observed between EC50 and LC50 for the 4 PFAS. These concentrations were 300 higher than environmental levels reported for these PFAS (exceeding the maximum concentrations 301 detected in the environment by more than 7x103fold higher) (Table S1). Regarding chronic 302 exposure (Figure 3B), data demonstrated that all tested TFA concentrations did not induce 303 morphological alteration on H. vulgaris compared to control. 304 Our results showed no significant differences in number of hydranths (hydranths develop on the 305 main body of the hydra and eventually detach to become separate individuals, a budding structure 306 is also considered a hydranth) after acute and chronic exposure between controls and treatments , 307 No significant, change was observed after exposure to the two lowest concentrations of TFA for 308 7day (Figure 4A). In addition, data demonstrated that the growth (K) and budding rates (Ktot-7d) 309 were not markedly affected by differing treatment (Figure 4B). Similarly to D. magna, GST 310 activity was not significantly altered after chronic exposure to TFA for 7day by comparison to 311 control (Figure 2B). 312 Discussion 313 The available literature on the toxicity of short- and ultra-short-chain PFAS is limited. To better 314 understand their potential toxicity on aquatic organisms, 4 short-chain PFAS (PFBA, PFBS, TFA, 315 and TFMS) were tested using D. magna and H. vulgaris. Our results showed that EC50/LC50 were 316 18 significantly different between organisms exposed in media with (>1000 mg/L) and without pH 317 adjustment (316 and 31.6 mg/L for D. magna and H. vulgaris, respectively). This finding 318 suggested that toxicity was significantly affected by media acidification initiated by different 319 PFAS as previously reported (Wasel et al. 2021). The different LC50 values observed between 320 species may be related to specific hydra media which seems to be more susceptible to acidification 321 by PFAS than that of daphnids. The addition of 100 mg/L of PFAS surpassed the buffering 322 capacity of the hydra medium and produced a drop in pH, from 7 to < 4, while the daphnid media 323 pH remained close to that of control; only addition of 1000 mg/L of PFAS induced media 324 acidification. The concentrations used in the present study may have been too low to provide 325 substantial acute and chronic toxicity data and a toxicity rank for buffered PFAS. Further testing 326 with higher concentrations is needed to determine which of these two species is more sensitive to 327 short-chain PFAS. 328 To the best of our knowledge, the effects of most PFAS on hydra remain largely unknown. 329 Only one other study investigated the effects of PFOA and PFOS on H. vulgaris and Wang et al. 330 (2021) reported that the minimum effective dose estimated at 100 mg/L resulted in 100% mortality 331 after 92hr treatment. Comparing the 48hr-EC50/LC50 values for PFBA found for D. magna in the 332 present study with literature data, demonstrated that acute value was in a similar range than that 333 reported by Ding et al. (2012) (180 mg/L without pH adjustment and 4260 mg/L with pH 334 adjustment) and Barmentlo et al. (2015) (48h-EC50 = 5250 mg/L) for this species. Barmentlo et al. 335 (2015) suggested that acute toxicity attributed to perfluoroalkyl acids in D. magna rose with 336 increasing carbon-chain length, as these investigators found that the 48 hr-EC50 values for PFHxA 337 (perfluorohexanoic acid, C6; 1048 mg/l) and PFOA (C8; 239 mg/L) were 5- and 21-fold lower 338 than that of PFBA, respectively. This observation indicating that chain length acts as a driver of 339 19 differential PFAS-initiated toxicity was also corroborated by mortality rates obtained with in vivo 340 and in vitro zebrafish models (larvae and embryonic fibroblast cells) (toxicity of PFOA > PFHxA 341 > PFBA) (Wasel et al., 2021). It was also stated that toxicity was elevated with sulfonate group, 342 since potassium perfluorobutane sulfonate K-PFBS (C4; 96 hpf-LC50 =1,394 mg/L) exhibited 343 higher toxicity than PFBA (C4; 96 hpf-LC50 = 9,703 and 83.6 mg/L for buffered and unbuffered 344 PFBA, respectively) (Wasel et al., 2021). In agreement with our findings, these outcomes confirm 345 that buffered PFBA and PFBS display toxicity at concentrations higher than in 1000 mg/L. 346 However, the role of chain length in the toxicity of PFAS was not supported by the study of Ulhaq 347 et al. (2013) indicating that the order of toxicity (144 hr-EC50/LC50) for zebrafish embryo was as 348 follow: PFBS (144 hr-EC50/LC50= 450 /1500 mg/L) > TFA (144 hr-EC50/LC50= 700 / >3000 mg/L) 349 > PFBA (144 hr-EC50/LC50= 2200 / >3000 mg/L). This was also in agreement with Wang et al. 350 (2014) in freshwater rotifer, Brachionus calyciflorus, reporting that acute effects (24hr-LC50) of 351 TFA (2C; 70 mg/L), PFPrA (perfluoropropionic acid, 3C; 80 mg/L), PFBA (4C; 110 mg/L), 352 PFPeA (perfluopentanoic acid, 5C; 130 mg/L) and PFHxA (perfluorohexanoic acid, 6C; 140 353 mg/L) were negatively correlated to the carbon chain length due to higher water solubility of 354 shorter PFAS. However, these findings need to be considered with caution since the observed 355 effects were related to the acidification of exposure media. A lower 24hr-EC50 (55 mg/L) was also 356 estimated by Rhône-Poulenc 1995 (cited in Boutonnet et al., 1999) for D. magna exposed to TFA 357 and this was attributed to the pH effect. It was also noted that 1200 mg/L sodium trifluoroacetate 358 (corresponding to 1000 mg/L trifluoroacetate) did not markedly induce any effects on D. magna 359 after exposure for 48hr (Berends et al., 1999; Boutonnet et al., 1999), and on Brachydanio rerio 360 (Boutonnet et al., 1999) and Danio rerio (Berends et al., 1999) after 96hr of treatment. 361 20 Overall, discrepancies observed in the data reported may be explained, at least in part, by 362 variations in culture conditions, experimental design, such as the lack of adjusting solutions to 363 neutral pH, and by differences in organism sensitivity. Although it is difficult to draw a conclusion 364 on the most toxic unbuffered PFAS (LC50 >1000 mg/L for all tested PFAS), the available results 365 suggest limited impacts for these individual short- and ultra-short chain PFAS to freshwater 366 organisms, as the estimated LC50 values were far higher than PFAS concentrations detected in the 367 environment. The PFAS tested in the present study may be considered as not toxic according to 368 the EU Directive 93/97/EEC classification of toxic contaminants (CEC, 1996) as the acute LC50 369 values were all above 100 mg/L. Additional acute toxicity research is needed to further address 370 potential ecotoxicity of short- and ultra-short chain PFAS across different aquatic species. 371 Since toxicity generally tends to rise with increasing exposure time, chronic exposure to TFA 372 was carried out using D. magna and H. vulgaris. Our sub-chronic results demonstrated that 14day 373 exposure of D. magna to TFA, up to the maximum dose of 100 mg/L, did not significantly induce 374 any effects on survival, body size, and number of neonates. In addition, data demonstrated that 375 exposure to 10 mg/L TFA did not markedly alter reproduction or growth rate of hydra after 7day 376 treatment. To the best of our knowledge, there is no apparent published information on chronic 377 exposure of aquatic organisms to TFA. A study that examined the combined effects of TFA salt 378 and trichloroacetic acid salt in Myriophyllum spicatum and Myriophyllum sibiricum noted that this 379 mixture did not induce marked effects in these aquatic plants after 35day exposure (Hanson et al., 380 2002). However, although no marked effects were observed under the present tested experimental 381 conditions, the findings need to be considered with caution. Previously, Xie et al (2025) found that 382 toxicity of PFBS in D. magna increased over 6 generations; the 21day-LC50 for F0 and F3 were 383 >1470 and 483 mg/L, respectively, and the F0-EC50 for reproduction (856 mg/L) and growth (886 384 21 mg/L) decreased respectively to 470 and 499 mg/L for F3 in relation with elevated PFBS 385 bioaccumulation in subsequent generations. Exposure over multiple generations may therefore 386 affect daphnia resistance and ultimately affect survival, reproduction, and population growth (Xie 387 et al., 2025). 388 Given the paucity of published literature, more information is needed regarding the chronic 389 toxicity and mechanisms of action of short-and ultra-short chain PFAS to better understand the 390 potential effects of these chemicals. Several investigators reported that the mechanisms of long-391 chain PFAS toxicity in aquatic organisms acted via a series of reactions in cells, including 392 oxidative stress which subsequently induced modulation in antioxidant defense system such as 393 GST activity (Jeong et al., 2016). An increase in GST activity was observed in offspring of D. 394 magna exposed to 10 mg/L of PFOS (Jeong et al., 2016), while a reduction in GST activity was 395 detected in D. magna chronically exposed to 8 mg/L of PFOS for 10 and 21days (Liang et al., 396 2017). Huang et al. (2022) reported a significant rise in GST activity after exposure of zebrafish 397 for 28days to 100 µg/L of a mixture of 5 PFAS (chlorinated polyfluorinated ether sulfonate (F-398 53B), PFOS, sodium p-perfluorous nonenoxybenzenesulfonate (OBS), PFHxS, and PFBS, 399 suggesting that GST played a role in the PFAS detoxification. The absence of GST induction in 400 the present study suggested that tested concentrations of TFA were not sufficient to trigger stress 401 in daphnids and hydra. However, it should be noted that one cannot attribute oxidative stress to 402 the actions of PFAS as only one parameter (GST activity) was examined. This constitutes a 403 limitation as future studies need to also determine carbonyl and MDA levels to assess oxidative 404 stress. 405 406 407 22 Conclusions 408 Overall, acute and sub-chronic/chronic results from this study indicated a lack of toxicity of the 409 tested short- and ultra-short chain PFAS to D. magna and H. vulgaris. Findings also indicated that 410 assessing the pH of media is essential to the accurate interpretation of PFAS toxicity results. Given 411 the low toxicity observed for PFBA, PFBS, TFMS, and TFA relative to the concentrations reported 412 in aquatic environments, it may be important to consider longer exposure (months, years, whole 413 life cycle) of aquatic organisms to better assess their toxicity at biochemical, molecular, cellular, 414 and physiological levels. It is also crucial to consider realistic environmental exposure and 415 multigeneration scenarios to overcome the significant gaps in our understanding of short /ultra-416 short PFAS toxicity. Further toxicity tests of individual and combined PFAS across different 417 aquatic species are needed to better determine their environmental impact. In this context, and with 418 growing concerns regarding short- and ultra-short PFAS, in vitro methods may provide a rapid 419 screening tool for their toxic potential and help in elucidating their mechanisms of action. 420 Acknowledgements 421 This project was funded by the Government of Canada’s Chemicals Management Plan. We thank 422 Sylvain Trottier for collaboration and coordination. 423 Disclosure statement 424 The authors report there are no competing interests to declare. 425 Data availability statement 426 Data are available on the Environment and Climate Change open government portal 427 (https://open.canada.ca/en). 428 429 430 23 CRediT authorship contribution statement 431 Houda Hanana: Conceptualization, methodology, formal analysis, writing - original draft, writing 432 - Review and editing, visualization. 433 Marie Houédé: Methodology, formal analysis, writing - original draft, writing - Review and 434 editing, visualization. 435 Geneviève Farley: Methodology, formal analysis, review and editing. 436 Pascale Bouchard: Methodology, formal analysis, review and editing. 437 Roxane Sorel: Methodology, formal analysis, review and editing. 438 Quoc Tuc Dinh: Methodology, formal analysis, review and editing. 439 Sébastien Sauvé: Supervision, review and editing. 440 Magali Houde: Conceptualization, supervision, project administration, funding acquisition, 441 review and editing 442 443 444 445 446 447 448 449 450 451 452 453 454 455 456 457 458 24 References 459 Ahmed, S. 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Chemosphere, 258, 127255. 724 https://doi.org/https://doi.org/10.1016/j.chemosphere.2020.127255 725 726 727 728 https://doi.org/https:/doi.org/10.1016/j.scitotenv.2014.02.028 https://doi.org/https:/doi.org/10.1016/j.envint.2013.08.021 https://doi.org/10.1080/15287394.2020.1842272 https://doi.org/https:/doi.org/10.1016/j.ecoenv.2025.118078 https://doi.org/https:/doi.org/10.1016/j.scitotenv.2023.167313 https://doi.org/https:/doi.org/10.1016/j.chemosphere.2007.07.079 https://doi.org/https:/doi.org/10.1016/j.watres.2016.05.082 https://doi.org/https:/doi.org/10.1016/j.scitotenv.2016.05.221 https://doi.org/10.1007/s11356-012-0977-y Figure captions Figure 1. Effect of sub-chronic exposure to different TFA concentrations on survival (A), hatching (B), number of neonates per female (C) and body length (D) of D. magna after 14d of exposure. Data are expressed as mean ± SD. Figure 2. Changes in GST activity level in A) D. magna after 14d and B) H. vulgaris after 7d of exposure to TFA. Data are expressed as mean ± SD. Figure 3. Stacked bar chart showing the number of H. vulgaris with sublethal (normal, clubbed and shortened tentacles) and lethal (tulip and disintegrated) morphological changes after A) 96h of exposure to short- and ultra-short PFAS and B) after 7d of exposure to TFA. Figure 4. Mean number of (A) hydranth after acute and chronic exposure (96h and 7d) and (B) population growth rate (K) and budding rate (Ktot) after 7d of exposure to TFA. Data are expressed as mean ± SD. Figure 1 Figure 2 Figure 3 B A Figure 4 1 Supplementary Material Toxicity assessment of exposure to four short- and ultra-short PFAS using Daphnia magna and Hydra vulgaris as models Houda Hanana a, Marie Houédé b, Geneviève Farley c, Pascale Bouchard c, Roxane Sorel c, Quoc Tuc Dinh d, Sébastien Sauvé d, Magali Houde a* a Aquatic Contaminants Research Division, Environment and Climate Change Canada, Montreal, QC, Canada b University of Angers, Angers, France c Quebec Laboratory for Environmental Testing, Environment and Climate Change, Montréal, QC, Canada d Chemistry Department, Université de Montréal, Montréal, QC, Canada *Corresponding author: Magali Houde Environment and Climate Change Canada Tel: 1-514-742-5124 Email: magali.houde@ec.gc.ca mailto:magali.houde@ec.gc.ca 2 Table S1. Concentrations (ng/L) of TFA (2C), TFMS (1C), PFBA (4C), and PFBS (4C) in environmental waters reported in the literature after 2000. See Jordan and Frank (1999) and Zhang et al. (2005) for reviews of TFA concentrations reported in waters before 2000. *Review of PFBA and PFBS reported in China surface waters between 2006-2022. nd = not detected. Location Year of collection Media TFMS TFA PFBA PFBS Reference Canada 1997 Lakes (<0.5-360) Scott et al. 2000 Switzerland 1996-1997 Rivers 87 (12-328) Berg et al. 2000 Midland lakes 119 (37-204) Mountain lakes 141 (46-360) Canada/United States 1998 Lakes, river 18-150 Scott et al. 2002 Africa Lake 1-5 China 2001 Lakes (6.8-98) Zhang et al. 2005 Rivers (10-221) 1999-2001 Seawater (6-190) Canada 2001 Lakes (61-170) Scott et al. 2006 Germany 2007 River 1.7-3.4 Ahrens et al. 2009 India 2008 Rivers nd Yeung et al. 2009 Netherlands 2008 Rivers, seawater (0.22-335) (<0.08-181) Möller et al. 2010 France 2010 River 1.3 (0.6-2.6) Labadie and Chevreuil 2011 3 Faroe Islands (Kingdom of Denmark) 2012 Lakes (<0.04) (<0.04-1.2) Eriksson et al. 2013 Norway 2006 Lake, river, seawater (0.06-1.2) a Kwok et al. 2013 Japan 2019 Rivers, seawater (nd-18) (nd-49) Takemine et al. 2014 Germany 2013-2014 River, estuary, coast (<0-3.7) (1.4-40) Heydebreck et al. 2015 Netherlands River (<0-1.9) (0.8-3.6) China River (4.8-4770) nd Spain 2010 Rivers 20 (0.1-111) 0.3 (0.4-4.1) Campo et al. 2015 Italy Rivers 5 (0-52) 11 (0-66) Castiglioni et al. 2015 Vietnam 2013 Rivers, canals nd Duong et al. 2015 France 2012 Lakes, rivers (<0.17-11) (<0.02-29) Munoz et al. 2015 China 2012 Landscape waters 643 (345-828) Zhai et al. 2015 India 2014 River (<0.004-1) <0.04–10.2 Sharma et al. 2016 Germany 2016-2017 River, streams (<50-140 000) Scheurer et al. 2017 China and South Korea 2016 Rivers, seawater (<1.3-35) (<0.13-1.5) Zhou et al. 2018 Australia 2012 River, estuary (1.7-11) (0.4-7) Allinson et al. 2019 Finland 2016-2017 Rivers 1 (max 5.3) 0.23 (max 1.5) Junttila et al. 2019 Canada 2008-2017 Lakes max 4.9 < limits of detection Gewurtz et al. 2019 4 a Range of means Norway 2014-2016 Lakes, rivers, seawater (<0.08-2) <0.003 Skaar et al. 2019 Philippines not found Lake (<0.06-22.1) (0.5-5.15) (<0.03-1.2) Guardian et al. 2020 Thailand not found River, dams (<0.06-30.7) (0.4-6.5) (<0.03-2.4) South Korea 2017-2018 Lake 17 (1.4-43.6) Lee et al. 2020 United States 2016 River (0.9-1.4) (0.5-2.5) Penland et al. 2020 Sweden 2019 Streams, lake (0.11−15) (30-820) (<0.1-12) (<0.1-22) Björnsdotter et al. 2022 United States 2020 Rivers (nd-79) Viticoski et al. 2022 Canada 2019 River (3.6-87) (0.42-0.52) Munoz et al. 2022 Canada 2013 to 2020 Lakes, rivers max 73 max 138 Lalonde and Garron 2022 Canada 2012-2015 Proglacial rivers, creeks (0.6-6.2) (<0.002-0.3) MacInnis et al. 2022 China 2018 River estuaries 17 (0.3-283) 12 (nd-575) Du et al. 2022 Tibet 2020-2021 Rivers, Tibetan plateau (0.09-2.8) (0.04-4.0) Ren et al. 2023 Portugal 2017-2018 Rivers (<0.040-23) (<0.020-7.1) Barbosa et al. 2023 China 2002-2022 Lakes, rivers, reservoirs (0.03-3720) (0.03-3720) Zhang et al. 2023* 5 Table S2. Chemical information for the four studied PFAS. Compound Acronym Structure Mass Internal standard used Trifluoroacetic acid TFA CF3COO- 112.98449 TFA-13C2 Perfluorobutanoic acid PFBA CF3CF2CF2COO- 212.9781 PFBA-13C4 Trifluoromethane sulfonic acid TFMS CF3SO3- 148.95148 PFBA-13C4 Perfluorobutane sulfonic acid PFBS CF3CF2CF2CF2SO3- 298.94189 PFBS-13C3 6 Table S3. Physico-chemical parameters of Daphnia magna and Hydra vulgaris exposure media without pH adjustment. Daphnia magna Nominal concentration (mg/L) Time (h) PFBS PFBA 0 0.01 1 10 100 1000 0 0.01 1 10 100 1000 Dissolved oxygen (%) 0 96.8 96.9 98.2 97.3 98.1 99.0 97.6 98.4 99.0 98.3 98.7 98.7 48 104.8 105.4 105.3 105.7 105.3 107.7 104.5 105.7 104.5 102.0 102.6 105.7 Temperature (ºC) 0 20.3 20.4 20.5 20.1 20.4 20.3 20.6 20.5 20.6 20.5 20.4 - 48 19.5 19.6 19.7 19.6 19.7 19.8 19.7 19.5 19.6 19.7 19.7 20.1 pH 0 8.03 8.05 8.01 7.79 6.92 2.98 8.01 8.04 8.02 7.71 6.82 2.82 48 7.81 7.86 7.85 7.73 7.13 2.90 7.88 7.92 7.92 7.75 7.05 2.85 Conductivity (uS/cm) - 317 317 317 317 313 1007 317 318 318 318 313 1469 Nominal concentration (mg/L) Time (h) TFMS TFA 0 0.01 1 10 100 1000 0 0.01 1 10 100 1000 Dissolved oxygen (%) 0 97.6 97.7 98.2 98.2 98.3 100.5 97.2 98.0 96.8 97.0 98.0 99.8 48 101.6 101.0 101.1 101.3 101.2 102.7 99.9 100.6 100.1 99.4 100.7 101.3 Temperature (ºC) 0 18.9 18.9 19.1 19.1 19.1 19.4 19.0 19.1 19.1 19.2 19.2 19.3 48 19.8 19.7 19.7 19.6 19.7 19.7 20.1 19.9 19.9 19.9 19.9 20.1 pH 0 8.02 8.05 7.99 7.57 6.41 2.60 8.04 8.03 7.94 7.45 5.94 2.50 48 7.90 7.97 7.94 7.74 6.89 2.72 7.85 7.84 7.85 7.60 6.34 2.58 Conductivity (uS/cm) - 317 317 317 317 317 2220 317 317 317 317 315 3240 7 Hydra vulgaris Nominal concentration (mg/L) Time PFBS PFBA 0 0.000 1 0.001 0.01 1 10 100 1000 0 0.0001 0.001 0.01 1 10 100 1000 Dissolved oxygen (%) 0 95.0 97.9 97.8 98.9 98.3 97.6 97.6 99.0 95.4 97.2 98.7 98.4 100.5 97.7 99.8 99.6 96 104.7 103.5 104.4 104.8 102.3 103.5 102.7 102.3 104.8 104.4 103.9 103.2 103.5 103.7 104.5 103.7 Temperature (ºC) 0 19.7 19.6 19.7 19.5 19.5 19.5 19.5 19.4 20.1 20.1 20.2 20.1 20.1 20.1 20.0 20.0 96 20.4 20.2 20.1 20.1 20.1 20.1 20.0 20.0 20.1 19.9 19.7 19.7 19.6 19.6 19.6 19.6 pH 0 7.04 7.02 7.03 7.03 7.02 6.90 3.85 2.88 7.04 7.02 7.00 7.03 7.00 6.80 3.61 2.72 96 6.81 6.84 6.85 6.88 6.88 6.75 3.84 2.84 6.82 6.87 6.85 6.87 6.86 6.64 3.55 2.78 Conductivity (uS/cm) 0 266 269 269 269 269 269 323 1405 266 268 268 268 268 268 371 1847 96 265 269 269 269 269 269 322 1390 265 269 269 268 269 269 369 1839 Nominal concentration (mg/L) Time (h) TFMS TFA 0 0.000 1 0.001 0.01 1 10 100 1000 0 0.0001 0.001 0.01 1 10 100 1000 Dissolved oxygen (%) 0 96.6 98.0 96.6 96.6 98.0 98.0 96.6 96.7 93.0 96.7 98.1 92.5 96.7 95.5 96.3 98.0 96 107.2 106.8 106.1 108.3 107.7 107.1 106.6 107.2 109.1 107.5 106.7 107.5 107.5 106.9 107.1 103.9 Temperature (ºC) 0 18.6 18.4 18.2 18.0 18.1 18.0 18.0 18.1 17.8 17.6 17.6 17.6 17.7 17.5 17.4 17.5 96 21.2 20.9 20.8 20.7 20.7 20.6 20.6 20.5 20.8 20.5 20.4 20.4 20.3 20.3 20.2 20.3 pH 0 7.14 7.14 7.13 7.14 7.08 6.65 3.36 2.64 7.18 7.15 7.15 7.15 7.11 6.82 3.48 2.72 96 6.98 6.93 6.98 7.01 6.96 6.62 3.48 2.81 6.97 6.95 6.97 6.98 6.92 6.51 3.34 2.58 Conductivity (uS/cm) 0 268 262 269 269 269 271 529 3170 268 270 269 270 270 272 462 2770 96 267 271 270 270 270 272 354 2690 267 271 269 269 278 271 521 3080 8 Table S4. Physico-chemical parameters of Daphnia magna and Hydra vulgaris exposure media with pH adjustment. Daphnia magna Nominal concentration (mg/L) Time (h) PFBS PFBA 0 0.01 1 10 100 1000 0 0.01 1 10 100 1000 Dissolved oxygen (%) 0 97.7 96.9 97.7 97.8 98.1 98.4 99.0 99.3 99.3 99.0 99.2 100.2 48 105.6 106.6 106.1 105.7 104.7 106.2 104.0 105.5 106.3 105.3 104.9 105.1 Temperature (ºC) 0 20.6 20.6 20.5 20.1 20.5 20.7 20.6 20.5 20.5 20.5 20.6 20.7 48 19.5 19.5 19.5 19.6 19.4 19.5 19.8 19.8 19.8 19.8 19.8 19.8 pH 0 8.07 8.07 8.08 7.79 8.07 7.99 8.05 8.06 8.07 8.07 8.04 7.77 48 7.83 7.88 7.89 7.73 7.89 7.87 7.82 7.92 7.88 7.85 7.76 7.64 Conductivity (uS/cm) - 316 316 317 317 339 530 317 317 317 320 350 639 Nominal concentration (mg/L) Time (h) TFMS TFA 0 0.01 1 10 100 1000 0 0.01 1 10 100 1000 Dissolved oxygen (%) 0 98.2 99.1 98.5 99.1 98.5 99.8 96.8 97.1 98.2 97.1 97.1 98.3 48 103.4 104.1 103.9 104.0 104.0 104.3 101.4 99.9 100.0 100.2 100.4 101.3 Temperature (ºC) 0 19.0 18.9 18.9 19.0 19.1 19.4 19.0 18.8 18.9 18.9 19.0 19.1 48 20.0 19.9 20.0 20.0 19.9 19.9 19.7 19.6 19.7 19.8 19.6 19.6 pH 0 8.02 8.03 8.04 8.05 8.11 8.68 8.01 8.05 8.05 8.03 7.88 7.23 48 7.88 7.92 7.93 7.95 7.96 8.41 7.80 7.86 7.87 7.86 7.80 7.34 Conductivity (µS/cm) - 318 317 318 323 371 840 318 317 317 324 383 959 9 Hydra vulgaris Nominal concentration (mg/L) Time (h) PFBS PFBA 0 0.0001 0.001 0.01 1 10 100 1000 0 0.0001 0.001 0.01 1 10 100 1000 Dissolved oxygen (%) 0 96.2 101.7 102.6 102.0 102.4 101.5 102.2 102.1 96.8 101.8 102.0 101.3 102.2 101.0 102.4 102.6 96 103.4 103.9 104.2 104.2 104.6 105.3 103.7 104.2 105.5 103.6 103.6 104.4 104.3 103.4 103.9 103.4 Temperature (ºC) 0 19.6 19.6 19.5 19.4 19.4 19.4 19.4 19.6 20.0 19.9 19.9 19.8 19.7 19.7 19.6 19.8 96 19.9 19.9 19.9 19.9 19.9 19.8 19.7 19.7 20.0 19.7 19.6 19.6 19.5 19.5 19.5 19.5 pH 0 7.07 7.05 7.06 7.04 7.03 7.02 7.02 7.02 7.00 7.03 7.03 7.03 7.01 7.03 7.06 7.32 96 6.87 6.91 6.92 6.90 6.91 6.88 6.84 6.85 6.89 6.98 6.97 6.99 6.95 6.95 6.96 7.17 Conductivity (uS/cm) 0 270 269 268 269 269 271 287 456 272 267 267 267 268 271 302 596 96 270 269 268 269 270 271 288 455 274 268 268 268 268 272 302 594 Nominal concentration (mg/L) Time (h) TFMS TFA 0 0.0001 0.001 0.01 1 10 100 1000 0 0.0001 0.001 0.01 1 10 100 1000 Dissolved oxygen (%) 0 93.0 91.5 95.0 94.7 95.1 94.4 94.8 95.6 94.5 96.6 97.6 96.5 96.2 96.0 95.9 96.3 96 108.6 107.6 108.0 108.4 107.0 107.3 107.2 106.3 102.6 108.6 108.7 108.4 107.7 107.8 108.1 106.8 Temperature (ºC) 0 18.4 18.6 18.5 18.5 18.5 18.4 18.4 18.5 18.3 18.5 18.5 18.4 18.4 18.3 18.4 18.6 96 20.4 20.4 20.4 20.4 20.4 20.4 20.4 20.4 20.5 20.1 20.1 20.1 20.2 20.1 20.1 20.1 pH 0 7.18 7.19 7.19 7.18 7.17 7.15 7.15 7.15 6.97 7.02 7.04 7.07 7.10 7.11 7.11 7.07 96 6.84 6.96 6.98 6.99 7.00 7.01 7.01 6.99 6.91 6.98 6.97 6.98 6.95 6.92 6.94 6.89 Conductivity (µS/cm) 0 270 268 268 268 268 274 327 825 265 268 268 268 269 277 344 1008 96 267 269 268 268 269 274 326 818 271 269 268 268 270 277 344 1000 10 Table S5. EC50 and LC50 (mg/L) of short- and ultra-short PFAS tested in the current study. CL= Confidence Level. Daphnia magna Hydra vulgaris Chemicals pH 48h-EC50 (95% CL) 96h-LC50 (95% CI) 96h-EC50 (95% CI) 96h-LC50 (95% CI) TFA Non-adjusted for pH 316 (100-1000) 316 (100-1000) 31.6 (10-100) 31.6 (10-100) pH adjusted > 1000 > 1000 > 1000 > 1000 PFBA Non-adjusted for pH 316 (100-1000) 316 (100-1000) 31.6 (10-100) 31.6 (10-100) pH adjusted > 1000 > 1000 > 1000 > 1000 TFMS Non-adjusted for pH 316 (100-1000) 316 (100-1000) 31.6 (10-100) 31.6 (10-100) pH adjusted > 1000 > 1000 > 1000 > 1000 PFBS Non-adjusted for pH 316 (100-1000) 316 (100-1000) 31.6 (10-100) 31.6 (10-100) pH adjusted > 1000 > 1000 > 1000 > 1000 11 Figure S1. Example of concentration-response curves analyzed by the CETIS software after acute exposure of D. magna neonates (A) and H. vulgaris (B) to the different tested PFAS without pH adjustment. 12 Table S6. Details on the on-line SPE HILIC-HRMS method for media analysis. Injection Volume 1000 μL SPE Column Biotage S Isolute ENV Online SPE cartridge (30 mm x 2.1 mm, 20- 80 μm particle size) On-line SPE mobile phases A: 0.0125% FA in HPLC-water, B: Acetonitrile On-line SPE gradient Time (min) % A % B Flow rate μL/min 0.0 100 0 2500 1.6 100 0 2500 1.7 0 100 1500 10.5 0 100 1500 10.6 100 0 1000 19.0 100 0 1000 20.0 100 0 2500 HILIC -HRMS system Dionex Ultimate 3000 UHPLC chain. heated electrospray ionization source (negative ion mode). Thermo Q-Exactive Orbitrap high- resolution mass spectrometer. Separation column Thermo Scientific Trinity Q1 column (100 × 2.1 mm. 3 µm particle size) Ionization Heated electrospray ionization source (HESI). negative ion mode Column oven temperature 35 °C HILIC mobile phases A: 25mM AmAc in HPLC-water, B: Acetonitrile. Mobile phase flow rate 450 μL/min HILIC gradient Time (min) % A % B 0.0 10 90 2.6 10 90 9.6 70 30 13.6 70 30 16.6 10 90 20.0 10 90 Source Sheath gas flow rate 50 a.u. Aux gas flow rate 20 a.u. Sweep gas flow rate 2 a.u. Capillary temperature (°C) 350. Vaporizer temperature (°C) 350. S-lens RF level 65. Q-Exactive Orbitrap settings Full Scan MS mode. Scan range (m/z) 150-600. Resolution 70.000 at m/z 200. AGC target 3e6. Max. Inject Time (ms) 100. 13 Table S7. Measured concentrations (mg/L, % nominal concentration) of the studied PFAS in exposure media of D. magna and H. vulgaris following acute exposure. Daphnia magna Nominal concentration (mg/L) PFBA-0h (mg/L) PFBA-48h (mg/L) PFBS-0h (mg/L) PFBS-48h (mg/L) TFA-0h (mg/L) TFA -48h (mg/L) TFOH-0h (mg/L) TFOH - 48h (mg/L) 0