This is the accepted version of the article: Avila, Romina; Peris, Andrea; Eljarrat, Ethel; [et al.]. «Biodegradation of hydrophobic pesticides by microalgae : transformation products and impact on algae biochemical methane potential». Science of the Total Environment, Vol. 754 (February 2021), art. 142114. DOI 10.1016/j.scitotenv.2020.142114 This version is available at https://ddd.uab.cat/record/237998 under the terms of the license 1 Biodegradation of hydrophobic pesticides by microalgae: transformation products and 2 impact on algae biochemical methane potential 3 Romina Avilaa, Andrea Perisb, Ethel Eljarratb, Teresa Vicenta, Paqui Blánqueza* 4 5 6 a Chemical,Biological and Environmental Engineering Department, Escola d'Enginyeria, 7 Universitat Autònoma de Barcelona, E-08193, Bellaterra, Barcelona, Spain. 8 9 b Water, Environmental and Food Chemistry, Dep. of Environmental Chemistry, IDAEA-CSIC, 10 Jordi Girona 18-26, 08034 Barcelona, Spain. 11 12 * Corresponding author: Paqui Blánquez 13 E-mail address: paqui.blanquez@uab.cat 14 Full postal address: Chemical, Biological and Environmental Engineering Department, Escola 15 d'Enginyeria, Universitat Autònoma de Barcelona, E-08193, Bellaterra, Barcelona, Spain. 16 17 Abstract 18 Intensive and extensive use of pesticides has contributed to their wide distribution in soil, air, and 19 water. Due to their detrimental effects on non-target organisms, different technologies have been 20 considered for their removal. In this work, three hydrophobic pesticide active compounds, namely, 21 chlorpyrifos, cypermethrin, and oxadiazon, were selected to study the potential for their removal 22 from aqueous media by a microalgae consortium. An abiotic and a killed control (thermally 23 inactivated dead microalgae biomass) were employed to clarify their removal pathways, and 24 pesticide content was quantified in liquid and biomass phases for 7 days. At the final time, total 25 degradation (biodegradation plus photodegradation) contributed to the removal of 55% of 26 oxadiazon, 35% of chlorpyrifos, and 14% of cypermethrin. Furthermore, more than 60% of 27 chlorpyrifos and cypermethrin were removed by sorption onto microalgae biomass. Overall, the 28 three pesticides showed high removal from the liquid phase. O,O-diethyl thiophosphate was 29 identified in the liquid phase as a transformation product of chlorpyrifos formed by microalgae 1 30 degradation. Phycoremediation was coupled with anaerobic degradation of the microalgae 31 biomass containing the retained pesticides by sorption through biochemical methane potential 32 tests. Anaerobic digestion was not inhibited by the pesticides as verified by methane production 33 yields. The removal efficiency of the pesticides in the digestate was as follows: chlorpyrifos > 34 cypermethrin > oxadiazon. These results highlight the potential of low-cost algal-based systems 35 for the treatment of wastewater or effluents from agrochemical industries. The integration of 36 wastewater treatment with biogas production through anaerobic digestion is a biorefinery 37 approach that facilitates the economic feasibility of the process. 38 Keywords 39 Biodegradation - Phycoremediation – Microalgae – Agrochemicals – Metabolites – Biochemical 40 methane potential 41 42 1. Introduction 43 Concerns about the adverse effects of agrochemicals in the environment are public and widely 44 known. The broad use of pesticides in agriculture leads to diffuse contamination through spray 45 drift and runoff, which contributes to their distribution in soil, air, and water. Likewise, point sources 46 of pesticides include their use in public health, green areas, livestock and other industries, and 47 households. Once pesticides are released into the environment, their fate, mobility, and 48 transformation are influenced by complex physical, chemical, and/or biological processes, such 49 as degradation, volatilisation, accumulation in soil, uptake by plants or microorganisms, and 50 transport to ground and surface waters [1]. Moreover, their distributions in soil, water, and air are 51 affected by transfer between phases and adsorption/desorption processes [2]. Pesticides taken 52 up by living organisms are susceptible to bioaccumulation [3], co-metabolic or partial 53 transformation into other degradation compounds, or mineralisation [4,5]. 54 The European Drinking Water Directive 98/83/EC defined 0.1 µg L-1 as the threshold for a single 55 pesticide and 0.5 µg L-1 as the threshold for total pesticides in human water consumption. 56 Conventional processes in wastewater treatment plants (WWTPs) are not efficient in the 2 57 degradation of exogenous pollutants because they are not designed for these purposes [6,7]. 58 Specially hydrophobic compounds are removed through sorption and accumulation in waste 59 activated sludge (WAS) [8], which could later be involved in other processes such as anaerobic 60 digestion or composting. However, if WAS is not previously treated and xenobiotics are not 61 degraded, they can accumulate in the media. Other conventional physical and chemical removal 62 treatments such as nanofiltration, advanced oxidation processes, and adsorption on activated 63 carbon are effective but expensive [9,10]. Over the past few years, assessments of the potential 64 of microalgae for directly transforming or enhancing the biodegradation of emerging 65 contaminants, heavy metals, and pesticides from wastewater is gaining attention among 66 researchers [11–13]. Microalgae-based wastewater treatment systems can facilitate algae 67 uptake, algae-mediated photolysis, bioaccumulation, and intracellular and extracellular 68 biodegradation of pollutants [14–16]. Microalgae can degrade complex parent compounds to 69 simpler molecules, highlighting their substantial biodegradation potential [17]. Moreover, 70 phycoremediation is a low cost and solar power-driven process that can be coupled with nutrient 71 removal and bioproduct recovery [12,18]. In algae-based systems, such as open ponds and 72 closed photobioreactors, microalgae produce the O2 required by heterotrophic aerobic bacteria 73 using CO2 released by these microorganisms [19]. Synergetic interactions between microalgae, 74 bacteria, and diverse microorganisms enhance the detoxifying potential of these systems [17]. 75 Therefore, microalgae-mediated bioremediation systems do not require the addition of carbon 76 sources or nutrients in stoichiometric balance, as required by bacteria and fungi [17]. 77 Among chemical families of pesticides, organophosphates and oxadiazol are widely employed 78 substances with effects on non-target organisms. Chlorpyrifos (CHL) is a chlorinated 79 organophosphate insecticide for pest control that is used broadly for a great variety of crops 80 [20,21]. Despite having a moderate persistence, these pesticides are highly toxic to mammals, 81 aquatic invertebrates, freshwater fish [5,22], and pollinisers, and they have neurotoxic, 82 immunological, and psychological effects in humans [4]. In January 2020, the European 83 Commission did not renew its approval of CHL (Regulation (EU) 2020/18). Cypermethrin (CYP) 84 is a pyrethroid insecticide used in pest control worldwide that is approved in the European Union. 85 Recently, the toxic effect of CYP on pollinisers [23] and the negative effects on the fertility, 86 immune system, cardiovascular system, and hepatic metabolism of mammals [24] were 3 87 confirmed. Oxadiazon (OXA) is a selective pre-emergence oxadiazoline herbicide against annual 88 dicotyledons, which is also used as a post-emergence measure against broad-leaved weeds 89 [25,26]. Due to its potential risks, OXA was included in the Watch List of Substances established 90 by EU Decision 2015/495, but this compound was removed from the second list (2018/840/EU). 91 Despite OXA being toxic to aquatic microorganisms [27], it continues to be detected in bodies of 92 water worldwide [28]. In humans, OXA could be associated with liver cancer and harmful effects 93 on reproductive and endocrine functions [29]. 94 Some authors have highlighted the potential of a microalgae–bacteria consortium for the removal 95 of xenobiotics, and some have studied degradation of pesticides using microalgae under sterile 96 conditions [30,31]; however, maintaining these conditions when scaling-up could be difficult and 97 the interaction effects between microorganisms and xenobiotic compounds have not been 98 evaluated. This work focused on assessing the potential of a microalgae consortium to degrade 99 three hydrophobic pesticides, CHL, CYP, and OXA, studying the evolution of their concentrations 100 in the liquid and solid phases while identifying their transformation products (TPs). With the aim 101 of biomass valorisation, the effect of the non-degraded pesticide taken up by the microalgae 102 biomass in methane production by anaerobic digestion was studied. 103 2. Materials and methods 104 2.1. Chemicals and reagents 105 The pesticides studied include three hydrophobic (log Kow > 4) compounds: CHL [(O,O-diethyl O- 106 (3,5,6-trichloro-2-pyridinyl) phosphorothioate], CYP [α-cyano-3-phenoxybenzyl ester of 2, 2- 107 dimethyl-3-(2, 2-dichlorovinyl) 2-2- dimethyl cyclopropane carboxylate], and OXA [5-terbutyl-3- 108 (2,4-dichloro-5-isopropoxyphenyl)-1,3,4-oxadiazol-2-one] (properties are provided in Table 1). 109 Stock solutions of each pesticide in methanol were prepared using analytical standards 110 purchased from Sigma-Aldrich (Steinheim, Germany) and stored in the dark at -20 °C until use. 111 Internal standards of d10-chlorpyrifos and phenoxy-d5-fenvalerate were also purchased from 112 Sigma-Aldrich; d7-oxadiazon was purchased from LGC standards (Teddington, Middlesex, UK). 113 Ethyl acetate, acetone, water, dichloromethane, hexane, and methanol solvents were purchased 114 form J.T. Baker (Waltham, Massachusetts, USA). Chloroform was purchased from Carlo Erba 115 (Val De Reuil, Eure, France). Formic acid (98–100%) was purchased from Merck (Darmstadt, 4 116 Germany). The quality of all solvents was according to organic trace analysis. Sulphuric acid was 117 obtained from Scharlau (Spain), and mineral salts and other chemicals were purchased from 118 Scharlab (Spain). GF/A and CG/C glass fibre filters were obtained from Whatman (GE Healthcare, 119 USA). 120 Table 1 121 (a) [32], (b) [33], (c) [34], (d) [35], (e) [36] 122 123 124 2.2. Microalgae cultures 125 The microalgae samples employed in the experiments were collected from a 1 m 3 outdoor tubular 126 semi-open photobioreactor (PBR), described previously by Hom-Diaz et al. [37], operating in 127 semi-continuous mode with feeding once per week using Bold Basal Medium. The microalgae 128 employed in the degradation experiments were taken 24 h after feeding. The consortium was 129 composed mainly of Chlorella sp. and Scenedesmus sp., morphologically examined using an 130 optical microscope (Zeiss, AixoCam ERc 5s), although other microorganisms such as bacteria 131 and protozoa were present, as reported previously for outdoor microalgae-based systems [37– 132 39]. 133 Biomass evolution along experimental time was determined by optical density (OD) and was 134 correlated to total suspended solids (TSS) concentration according to Eq. 1: 135 TSS (g L-1) = 0.7565 x OD680 – 0.0422 (r2 = 0.962) (1) 136 2.3. Pesticides removal by microalgae consortium 137 Batch biodegradation experiments of the target pesticides were performed in 250 mL Erlenmeyer 138 flasks containing 100 mL of microalgae solution from the PBR described previously. Flasks were 139 under orbital agitation (100 rpm) in a 25 ± 1 °C temperature-controlled chamber. A standard 140 solution of the target pesticide in methanol was spiked to the reactors to a final concentration of 141 1 mg L-1. In addition to the flask containing the microalgae consortium and the studied pesticide 142 (microalgae reactor), a killed control (dead microalgae biomass) employing thermally inactivated 143 biomass (autoclaved at 121 °C for 20 min) was used to determine the removal by sorption onto 144 biomass. The influence of other losses was assessed by an abiotic control containing the 5 145 pesticide in absence of microalgae. A blank reactor with the microalgae consortium without the 146 pollutants was used to assess the chemical effect on biomass growth. OD was monitored in the 147 microalgae reactors as an indicator of microalgae biomass growth. TSS and volatile suspended 148 solids (VSS) were determined from the blank reactor at the beginning and the end of the 149 experiments. All conditions were assayed under the influence of light during a 24 h photoperiod 150 (31 µmol photon m-2 s-1) provided by cool white fluorescent lamps. Bearing in mind that the studied 151 xenobiotics are non-polar molecules with a great partition from the aqueous phase into organic 152 solvents, determination of pesticide concentration was carried out at 2 and 7 days of exposure in 153 the aqueous and the solid phase. Identification of the TPs was performed in the liquid phase at 154 0, 2, and 7 days. The study conditions were evaluated independently for each pesticide in 155 triplicate. At each sampling time, a triplicate of each condition was used to perform analytics, while 156 the other reactors remained until the next sampling time. Biodegradation experiments were not 157 performed in aseptic conditions. 158 The removal kinetics of the pesticides was adjusted to a first-order reaction model [31], according 159 to Eq. 2: 160 C = C0 ∗ exp⁡(−k ∗ t) (2) 161 where C0 and C are the initial and final (t = 7 d) concentrations of the pesticide in the solution, 162 respectively, k is the kinetic rate constant (d-1), and t is the removal time (d). The theoretical half- 163 life (t1/2) of pesticides was calculated as ln2/k (d) [31]. 164 2.4. Biochemical Methane Potential (BMP) tests 165 The potential methane production was evaluated by biochemical methane potential (BMP) tests 166 employing a concentrated microalgae suspension from the biodegradation experiments at the 167 final time. A volume of 500 mL of solution from the microalgae flasks was sedimented naturally 168 for 12 h. Later, the supernatant was removed, and 250 mL of the settled biomass was employed 169 as a substrate for the BMP tests. 170 Anaerobic batch assays were performed according to a previously described procedure [40], 171 taking into account suggestions from other authors [41,42]. BMP tests were performed 6 172 considering an inoculum-to-substrate ratio (ISR) of 2 based on VS [40]. Inoculum (total solids (TS) 173 = 16.6 ± 0.4 g L-1 and volatile solids (VS) = 10.2 ± 0.2 g L-1) was collected from the anaerobic 174 digesters of the Riu Sec WWTP (Sabadell, Barcelona) and pre-incubated at 37 °C for 13 days to 175 reduce background production of biogas. The assays were carried out at mesophilic conditions 176 (37 °C) in triplicate using 120 mL glass bottles. They were filled with inoculum, substrate, and tap 177 water until the working volume (80 mL) was attained. Subsequently, reactors were flushed with 178 pure N2 to ensure anaerobic conditions, closed with a gastight butyl rubber septum, and incubated 179 in a temperature-controlled chamber. Biogas production and accumulation in the headspace of 180 the bottles were measured employing an SMC pressure switch manometer (1 bar, 5% accuracy) 181 until biogas generation ceased. Blank reactors (containing only inoculum) were used to calculate 182 the background biogas production of the inoculum. Net biogas production was determined by 183 subtracting the biogas production of the blank reactor from the gross biogas production of 184 microalgae reactors. Additionally, crystalline cellulose was used as a substrate in control reactors 185 to assess the biological activity of the inoculum. Reactors were shaken manually before each 186 pressure measurement. Periodically, the methane content of the generated biogas was analysed 187 by gas chromatography. Results of the BMP tests were expressed as the volume of methane 188 generated per mass of VS of the added substrate (mL CH 4 g VS-1) under standard temperature 189 conditions (273.15 K and 1.0133 bar). 190 The modified Gompertz equation [43] was employed to model the biomethane production and 191 calculate kinetic parameters according to Eq. 3: R .e 192 Pnet (t) = Pmax . exp⁡{−exp [ Pmax (λ − t) + 1]} (3) max 193 where Pnet(t) is the net cumulative methane yield (NmL CH4 g VS-1) at time t, Pmax is the methane 194 yield potential (NmL CH4 g VS-1), Rmax is the maximum daily methane production rate (NmL CH 4 195 g VS-1 d-1), t is the digestion time, and λ represents the lag phase (d). The hydrolysis rate was 196 estimated using a first-order kinetic model [44], as shown in Eq. 4: 197 B = B0 . [1 − exp(−K H . t)] (4) 7 198 where B is the cumulative methane yield (mL CH4 g SV-1), B0 is the ultimate methane yield (mL 199 CH4 g SV-1), t is the digestion time (d), and KH is the hydrolysis rate (d-1). Kinetic analyses were 200 performed using the software Matlab R2015a (MathWorks Inc., Natick, MA, USA). 201 2.5. Analytical methods 202 2.5.1.Quantification of pesticides and identification of TPs 203 At each sampling time (days 0, 2, and 7), the total volume was removed from three flasks. To 204 determine the pesticide concentration in the liquid phase, a fibreglass filter (1.6 µm, GF/A, 205 Whatman) was used to separate biomass from the aqueous phase and then 50 mL of the filtrate 206 was spiked with internal standards (d10-chlorpyrifos, phenoxy-d5-fenvalerate, and d7-oxadiazon) 207 to a final concentration of 0.1 mg L-1. The liquid samples were collected in amber glass tubes, 208 while the biomass cake retained by the filter was collected with a spatula and stored inside 209 aluminium bags to evaluate later the concentration of pesticide in the biomass phase. Additionally, 210 to determine the pesticide sorption or retention during the filtration, the pesticide concentration in 211 the employed filters was determined. All samples were frozen at -20 °C until analysis. 212 A 30 mL volume of liquid sample was ultrasonically extracted for 5 min with chloroform. Then, the 213 sample was centrifuged (3.500 rpm, 5 min) and the organic phase was recovered. The aqueous 214 phase was extracted once more and the organic phases were evaporated until dry under a 215 nitrogen stream. The residue was reconstituted with 50 μL of ethyl acetate and then subjected to 216 gas chromatography coupled to tandem mass spectrometry (GC-MS/MS) analysis. 217 Freeze-dried biomass samples and filters were extracted by pressurised liquid extraction using a 218 350 ASE system (Dionex, USA). Prior to extraction, biomass samples were spiked with 15 ng of 219 internal standard and left overnight at 4 °C. Then, samples and 2 g of Florisil were loaded into an 220 ASE extraction cell (22 mL) previously filled with 6 g of Florisil. Hexane and dichloromethane (1:1, 221 v/v) were used as the extraction solvent. Temperature and pressure were set at 100 °C and 1650 222 psi, respectively. Extracts were evaporated to dryness under a nitrogen stream, and the residue 223 was reconstituted with 50 μL of ethyl acetate prior to GC-MS/MS analysis. 224 Pesticide concentrations were determined through GC-MS/MS on a 7890B GC coupled to a 225 7000C triple quadrupole (Agilent technologies, USA) equipped with a DB-5MS capillary column 8 226 (30 m × 0.25 mm, 0.25 µm). The operating conditions were 80 °C for 2 min, raised at 25 °C min- 227 1 to 180 °C for 6 min, at 5 °C min-1 to 240 °C for 5 min, at 10 °C min-1 to 280 °C for 5 min, and at 228 30 °C min-1 to 325 for 2 min. The temperatures corresponding to the transfer line and the ionisation 229 source were 300 and 280 °C, respectively. The collision energy was 70 eV. Two different 230 transitions were monitored for each pesticide. 231 No analytes of interest were observed in the method blank samples. Recoveries ranged from 50 232 to 58%. Detection limits were 1.7, 2.8, and 0.9 ng L-1 for CHL, CYP, and OXA, respectively. Limits 233 of quantification were 5.6, 9.2, and 2.1 ng L-1 for CHL, CYP, and OXA, respectively. 234 Identification of TPs was carried out on a Waters Acquity UHPLC system (Waters, Milford, MA, 235 USA) coupled to a hybrid quadrupole-Orbitrap mass spectrometer Q-Exactive (Thermo Fisher 236 Scientific; San Jose, CA, USA) equipped with a HESI II heated-electrospray ionisation source. 237 Chromatographic separation was performed on a Purospher STAR RP-18 end-capped (2 µm) 238 Hibar HR 150-2.1 UHPL column (Merck). The mass spectrometer performed a Fourier transform 239 mass spectrometry scan event of 50-700 m/z at a resolution of 70,000 and a subsequent MS/MS 240 scan event was acquired at a resolution of 35,000. To identify all potential TPs, the total ion current 241 chromatograms acquired at 2 and 7 days were compared with those obtained at initial time using 242 Compound Discoverer (Thermo Fisher Scientific). This software allows differential analysis of 243 selected sets of samples by simultaneously comparing thousands of MS spectra to find significant 244 differences between the control and samples. The accurate masses of the potential TPs were 245 then extracted to confirm their presence. Identification of the potential TPs was based on their 246 accurate mass, mass error, molecular formula, and degree of unsaturation of the parent ion and 247 product ions. 248 2.5.2.Other analyses 249 Biogas composition (carbon dioxide and methane content) was analysed using a gas 250 chromatograph (Hewlett Packard 5890, Agilent Technologies, Mississauga, Canada) equipped 251 with a thermal conductivity detector and a Supelco Porapack Q column (3 m x 3.2 mm) 252 (Pennsylvania, USA). Helium was the carrier gas (338 kPa). Oven, injector, and detector 9 253 temperatures were 70, 150, and 180 °C, respectively. Samples were injected with a 100 µL 254 syringe (VICI PS Syringe A-2, 0.74 mm x 0.13 mm x 50.8 mm). 255 Acetic, propionic, and butyric acid concentrations were determined using a Dionex 3000 ultimate 256 high-performance liquid chromatography system (Barcelona, Spain) equipped with a UV/visible 257 detector (210 nm). The chromatographic separation was performed in an ICE-COREGEL 87H3 258 column (7.8 x 300 mm, Transgenomic, USA), heated at 40 °C, employing 0.006 mM of H2SO4 as 259 a mobile phase at a flow rate of 0.5 mL min-1. Samples were previously centrifuged (10 min, 8000 260 rpm, Beckman Coulter, Avanti J20 XP) and then filtered by 0.45 µm nylon syringe filters. 261 TS, VS, TSS, and VSS were determined following the procedures described in the Standard 262 Methods [45]. pH was measured using a pH meter (Crison, Spain), and OD was determined by 263 spectrophotometer (DR3900, Hach). 264 2.6. Data analysis 265 The experimental data were analysed statistically using a one-way analysis of variance (ANOVA) 266 of repeated measures, followed by a Bonferroni post-hoc test when significant differences were 267 identified (p < 0.05). P-values represent Bonferroni corrected significance levels. Statistical 268 calculations were carried out with R (version 3.6.3). 269 3. Results and discussion 270 3.1. Biodegradation of pesticides by microalgae consortium and TP identification 271 The characterisation of the initial microalgae biomass employed in batch experiments is shown 272 in Table 2. The biodegradation potential of the non-polar pesticides by the microalgae consortium 273 was evaluated by studying their behaviour under the following conditions: microalgae reactor, 274 killed control, and abiotic control. 275 Table 2 276 277 Figure 1 shows the evolution of the pesticide distribution in the liquid and solid phases in the 278 microalgae reactors. The remaining fraction of the target pesticide in the liquid and in the solid 10 279 phases through time was calculated as the relation between the residual quantity of the pesticide 280 in the liquid or solid phase and the total initial amount of pesticide in both phases. The solid phase 281 included the biomass contained in the flask and the filter used for biomass separation. Sorption 282 of the target pesticide onto the filter was determined for the microalgae reactors (1323.3 ± 131.3 283 ng of CHL, 11509.7 ± 3027.4 ng of CYP, and 920.3 ± 235.3 ng of OXA), and these values were 284 also applied to killed reactors. At the initial time, the three pesticides were mainly detected in the 285 liquid phase, where they were added; however, the residual pesticide distribution changed 286 gradually with time in the solid phase due to sorption onto the algae biomass. These compounds 287 have a tendency for sorption on solids and biosolids in concordance with their large log Kow value 288 and low solubility in water. They present a high solubility in lipids and an affinity for the microalgae 289 cell wall [46]. For instance, sorption onto the solid phase was also observed for pyrimethanil (a 290 fungicide with high log Kow value) when it was in contact with microalgae [47]. The pesticide 291 retained by the solid phase accounted for 62%, 60%, and 33% of the CHL, CYP, and OXA, 292 respectively (Fig. 1), highlighting the role of sorption in pesticide removal from the liquid phase. 293 The concentration of CHL in the aqueous media was reduced by 11 times after 2 days. The OXA 294 concentration in the solid phase remained fairly constant from day 2 to day 7 (36% to 33%). OXA 295 was gradually removed from the liquid phase due to sorption and degradation (by day 7, its 296 concentration in the liquid phase decreased by 8-fold). 297 298 Figure 1 299 (two-column fitting image) 300 301 Table 3 shows the distribution of the pesticides in the killed and abiotic controls after 7 days of 302 exposure. The remaining fraction of CHL in the solid phase of the killed control at the end of the 303 experiment (45.9 ± 11.5%) was lower than that in the microalgae reactor (61.9 ± 5.8%). With 304 regard to OXA, it showed a higher sorption capacity of the inactivated biomass (63.8 ± 32.1%) in 305 comparison with the active biomass (32.8 ± 12.7%). The cell wall of microalgae is constituted by 306 an aggregation of polymers with functional groups over its surface (such as carboxyl, phosphoryl, 11 307 and amine) that give it a negative charge [17]. Bearing in mind that the structure of the microalgae 308 surface is modified after heat treatment [48] for inactivation, sorption on the active biomass 309 (microalgae reactor) could be different from sorption on the inactivated biomass (killed control). 310 Furthermore, the absence of an active enzyme system after the biomass has been thermally 311 inactivated [14] could lead to dissimilarities when both conditions are compared. Additionally, 312 some authors have reported a major contribution of dead algal cells in the photolysis of 313 xenobiotics [14,49]. Luo et al. [14] proved that chlorophyll is the major active substance in dead 314 cells and generates singlet oxygen that acts as a catalyst stimulating and accelerating the 315 photodegradation of xenobiotics under light irradiation, which could also explain the removal 316 differences in both conditions. Concerning CYP, sorption onto solid phase was slightly higher for 317 killed reactors (Table 3) than for microalgae reactors (46% by day 2 and 60% by day 7). 318 Table 3 319 320 Table 4 shows the removal yields at the final time for the three pesticides. Removal from the liquid 321 phase was calculated assuming that the pesticide was in the liquid phase initially, and the removal 322 percentage was calculated based on the remaining pesticide in the liquid phase. The total 323 degradation is the difference between the removal from the liquid phase and the sorption onto the 324 solid phase. Photodegradation and other abiotic factors represent the pesticide removal 325 percentage of the abiotic control, and biodegradation by the microalgae consortium was defined 326 as the difference between total degradation and abiotic removal and other abiotic factors. 327 328 Table 4 329 330 As can be deduced from Figure 1 and is presented in Table 4, OXA showed the highest total 331 degradation (photodegradation or another mechanism such as biodegradation) by day 7 (55.5 ± 332 15.6%) in comparison with CHL (35.4 ± 7.1%) and CYP (13.8 ± 2.7%). OXA removal increased 333 with time from 41.0% at day 2 to 55.5% at day 7. In the case of CHL, the total degradation after 12 334 day 2 was higher (45.6 ± 10.6%) than the value at day 7, which could be attributable to pesticide 335 adsorption on the glass flasks and its later desorption, as was reported previously by others 336 authors considering hydrophobic pesticide removal [50]. The difference of pesticide degradation 337 of CYP between day 2 (20.8 ± 15.2%) and day 7 (13.8 ± 2.7%) could be associated with the same 338 cause. According to their removal, the theoretical half-lives of the pesticides in contact with 339 microalgae biomass were 38 d for CYP (r2 = 0.987), 13 d for CHL (r2 = 0.925), and 5 d for OXA 340 (r2 = 0.804). 341 As shown in Table 4, removal from the liquid phase included the contribution of pesticide sorption 342 to the solid phase plus the total pesticide degradation. In this sense, CHL exhibited the greatest 343 removal from the aqueous media (97.3%) after 7 days of treatment with the microalgae 344 consortium, followed by OXA (88.4%) and CYP (73.9%). Photodegradation was quantified by 345 considering pesticide removal in the abiotic control (Table 3) to analyse the influence of abiotic 346 processes on pesticide removal. CHL removal in the absence of the microalgae consortium 347 (abiotic control) was 16.3 ± 4.5%, suggesting a slight influence of photolysis and other abiotic 348 factors in pollutant removal. It has been reported that CHL has a limited potential for 349 photodegradation by natural sunlight irradiation in water [51,52]. With regard to CYP, 27.1 ± 350 12.0% was removed from the abiotic control. For OXA, no contribution of abiotic factors was 351 detected in the removal, suggesting that the pollutant elimination was influenced by neither 352 photodegradation nor volatilisation. Moreover, it has been reported that OXA presents a medium 353 to low volatility [27]. Bearing in mind that biodegradation was assumed to be the difference 354 between total pesticide removal and photodegradation (based on Hom-Díaz [53]), it can be 355 suggested that OXA removal can be mainly attributed to biodegradation (55.5 ± 15.6%) and 356 sorption (32.8 ± 12.7%). Regarding CYP, biodegradation was not identified, and the main removal 357 mechanisms seemed to be sorption (60.1 ± 11.1%) and photodegradation (27.1 ± 12.0%). 358 The results indicate that the total pesticide degradation in the microalgae reactors was higher for 359 OXA, followed by CHL and CYP, respectively. Biodegradation by the microalgae consortium 360 contributed to total degradation by 19.1% for CHL and 55.5% for OXA. Moreover, algal-mediated 361 photolysis seemed to have an effect in CHL and CYP losses (16.3% and 27.1%, respectively), 362 because oxygen and oxidant species generated by photosynthetic microorganisms are capable 13 363 of inducing indirect photolysis [54]. In the microalgae reactors, pesticide retention by sorption to 364 the solid phase was similar for CHL and CYP (61.9% and 60.1%, respectively) but lower for OXA 365 (32.8%). Overall, efficient removal from the liquid phase was accomplished for the three 366 pesticides (CHL > OXA > CYP). 367 Given that the identification of TPs is directly associated with microalgae consortium activity, they 368 were analysed in the microalgae reactors. As CHL has low solubility in water, its bioavailability for 369 microbial degradation and its use as a carbon source could be limited [55,56]. However, O,O- 370 diethyl thiophosphate (DETP) has been identified as a TP of CHL in microalgae reactors. The 371 corresponding retention times, measured masses, molecular formulae (calculated on the basis of 372 their accurate mass measurements and the observed isotopic patterns), relative mass 373 measurement errors, and degree of unsaturation, expressed as ring and double bound 374 equivalents, for this TP are summarised in Table 5. Higher intensities of this metabolite were 375 observed at day 7, suggesting that it was produced gradually as degradation proceeded. 376 CHL degrading microorganisms are able to produce organophosphate hydrolysing enzymes that 377 hydrolyse the P–O bond, leading to the corresponding dialkyl phosphate (DETP) and 378 corresponding aryl alcohol (3,5,6-trichloro-2-pyridinol, known as TCP) [57]. In this study, DETP 379 was identified as a TP of CHL by microalgae degradation. Studies examining CHL microbial 380 degradation have focused mainly on bacteria and fungi [5,58–62], but few studies have assessed 381 the capability of microalgae and cyanobacteria to degrade this compound. In general, it has been 382 reported that degrading bacteria hydrolyse CHL, producing two main metabolites: DETP and TCP 383 [56,63]. In this study, TCP was not identified, suggesting further conversion into other compounds 384 or complete mineralisation, as Barathidasan et al. [64] found with a Phanerochaete 385 chrysosporium fungal strain that could use TCP as a carbon source and completely mineralise 386 CHL. 14 387 388 Table 5 389 390 In this work, no TPs were identified in association with OXA degradation, which could indicate 391 that the molecule was mineralised. This assumption is coherent with the high biodegradation 392 detected (55.5%). Some authors have demonstrated an ability of other microorganisms to 393 biodegrade OXA. For instance, Pseudomonas fluorescens CG5 isolated from a soil contaminated 394 with herbicides was able to use OXA as a carbon source and further catabolise the compound, 395 obtaining a few toxic metabolites such as indole, benzoic acid, and trimethyl benzene, including 396 dehalogenation [25]. 397 Most of the published studies regarding CYP biodegradation are focused on bacteria and fungi 398 [65–68] and not on microalgae. In general, it has been reported that hydrolysis of the ester bond 399 is the main degradation pathway of CYP, producing the corresponding alcohol (2-hydroxy-2(3- 400 phenoxyphenyl)acetonitrile, known as CPBA) and acid (3-(2,2-dichloroethenyl)-2,2- 401 dimethylcyclopropane-1-carboxylic acid, known as DCCA) [2]. CPBA is further transformed, 402 leading to small molecular weight aliphatic compounds with intermediate compounds such as 3- 403 phenoxybenzoic acid (PBA) [2]. These aliphatic compounds include oleic acid [1,2], 2- 404 pentadecanone [2], acetic acid [4], decanoic acid [4], 1-dodecanol [1], and isopropyl myristate [1]. 405 Likewise, phototransformation products of CYP have also been reported previously, with DCCA 406 and PBA being the main photoproducts [69]. In this work, no TPs related to microalgae 407 biodegradation of CYP were identified. This result is consistent with the low observed total 408 degradation (13.8%), which indicates that the main removal mechanism of CYP in the liquid phase 409 is sorption onto the microalgae. This low degradation yield probably led to poor formation of TPs 410 and hence no detection. 411 The physiological status of microalgae is in concordance with biomass growth, and cell dry weight 412 is considered an integral parameter of cellular metabolism [70]. Thus, the evolution of biomass 413 concentration in microalgae reactors could be an indicator of the toxicity and inhibition caused by 414 the pesticide. As shown in Figure 2, the biomass concentration in the microalgae reactors 15 415 increased by 8%, 19%, and 15% for CHL, CYP, and OXA after 7 days, suggesting that these 416 molecules were not toxic to the microalgae consortium. Moreover, no statistical differences were 417 identified between the mean values of biomass in blank and microalgae reactors for each 418 pesticide at each sampling time (p > 0.05, except for OXA at time 0). As shown in Fig. 2, the 419 pesticide concentration was not toxic for the microalgae consortium. On the other hand, 420 ecotoxicological data indicate that the acute 72-h EC50 values of CHL, CYP, and OXA are 0.48 421 mg L-1 (for unknown microalgae species), > 0.0667 mg L-1 (for Pseudokirchneriella subcapitata), 422 and 0.004 mg L-1 (for Scenedemus subspicatus), respectively [71]. Bearing in mind that these 423 toxicological studies were performed using pure microalgae species, it is not possible to relate 424 them with the microalgae consortium employed in the present work. The advantages of working 425 with a microalgae consortium are the synergetic interactions between microorganisms that 426 enhance the robustness of the system and overall removal efficiency [17,39,72,73]. 427 Figure 2 428 (one-column fitting image) 429 430 3.2. Anaerobic treatment and methane production 431 Anaerobic digestion has been proposed as a technique for microalgae biomass treatment and 432 valorisation after the aerobic phase [74,75]. The anaerobic biodegradation and the biogas 433 production potential of the microalgae biomass containing the retained pesticide by sorption was 434 assessed by BMP test. During anaerobic digestion processes (hydrolysis, acidification, and 435 methanisation), pesticides and other compounds can be physically, chemically, and biologically 436 transformed [76] by diverse microorganisms and enzymes involved in each phase. Given the high 437 sorption affinity of the studied pesticides, their concentration in BMP tests was determined in the 438 biomass at the initial (t0) and final time (t42). 439 The net methane production (Fig. 3) was better for BMP reactors containing CYP and CHL, with 440 yields of 4880 and 4558 NmL CH4 g VS-1, respectively (differences were not statistically 441 significant). The methane yield in flasks containing OXA was remarkably lower (2919 NmL CH4 g 442 VS-1, p < 0.05). According to the Gompertz model, the lag phase was around 2.9 to 3.5 days 16 443 (Table 6), after which biogas production increased sharply at a rate of 27.4, 24.3, and 21.8 mL 444 CH4 d-1. VFAs quantified after the methanisation verified the stability of the process (Table 6), 445 indicating neither inhibition from the residual pesticide nor the potential TPs. 446 Analysis of the final digestate indicated that CHL and CYP were efficiently removed by 87.4 ± 447 0.5% and 58.7 ± 3.9%, respectively (difference of the mean pesticide content in the solid phase 448 between the initial time and final time was statistically significant in both cases, p < 0.01) (Fig. 4). 449 This is in agreement with the work of Lian et al. [77], in which contaminated biomass with the 450 insecticides parathion and malathion (organophosphorus compounds) was efficiently 451 anaerobically transformed by reduction reactions and enzymatic hydrolysis driven by hydrolases, 452 suggesting that hydrolytic activities provide a potential tool for biodegradation of 453 organophosphorus compounds such as CHL. Likewise, García-Mancha et al. [78] reported a 454 good removal efficiency (77%) of CHL from wastewater under thermophilic conditions (55 °C). 455 The results indicate that although OXA was highly degraded by the microalgae consortium under 456 aerobic conditions (55% biodegradation), its anaerobic removal was lower (18.7 ± 4.4%, p < 0.05) 457 (Fig. 4), leading to a minor methane yield. Otherwise, CHL and CYP were successfully degraded 458 by the anaerobic microorganisms as confirmed by their removal in the digestate. The results 459 indicate that anaerobic degradation of CYP was more effective than aerobic degradation with the 460 microalgae consortium. 461 Figure 3 462 (one-column fitting image) 463 464 Figure 4 465 (one-column fitting image) 466 467 Table 6 468 17 469 After the aerobic treatment, water could be treated in a municipal WWTP or discharged if 470 adequate parameters were fulfilled. Moreover, the solid phase containing pesticides could be 471 valorised as proposed in this work through methanisation for biomethane production. This 472 suggestion is in agreement with other studies [79–81]. Even though anaerobic digestion 473 contributed to the elimination of pesticides, the digestate obtained after the anaerobic digestion 474 could be dewatered. The solid fraction could be employed for fertilisation [82], and the water could 475 be treated in a WWTP or properly discharged if parameters were adequate given environmental 476 restrictions. 477 4. Conclusions 478 Phycoremediation coupled with anaerobic degradation of contaminated biomass was assessed 479 for hydrophobic pesticides removal. The distribution of pesticides on both liquid and solid matrices 480 was evaluated to determine the capacity of a microalgae consortium to degrade complex 481 molecules. Total degradation (biodegradation and photodegradation) was higher for OXA (55%), 482 followed by CHL (35%) and CYP (14%); whereas CHL and CYP exhibited greater sorption to the 483 solid phase (62% and 60%, respectively). Overall, a large removal from the liquid phase (total 484 degradation + sorption) was achieved for the target compounds (CHL: 97%, CYP: 74%, and OXA: 485 88%). The results evidence the potential of algae-based bioremediation technologies in the 486 bioconversion of agrochemicals. Further anaerobic degradation of the biomass containing the 487 target pesticides was performed without inhibition and led to high methane generation and 488 removal of the pesticides (removal efficiency was CHL > CYP > OXA). Based on this study, 489 microalgae biomass containing pesticides could be valorised efficiently by anaerobic digestion, 490 fostering biogas production. 491 492 Acknowledgements 493 This work was supported by the Spanish Ministry of Science and Innovation (Project CEX2018- 494 000794-S), the Spanish Research Agency partly supported by the European Regional 495 Development Fund (ERDF) through the project BECAS (CTM2016-75587-C2-1-R and 2-R) and 496 by the Generalitat de Catalunya (Consolidated Research Groups 2017-SGR-1404 and 2017- 18 497 SGR-014). The Department of Chemical, Biological and Environmental Engineering of the 498 Universitat Autònoma de Barcelona is member of the Xarxa de Referencia en Biotecnologia de 499 la Generalitat de Catalunya. 500 501 Competing interest statement 502 We declare that no conflict of interest exists in the submission of this manuscript. 503 504 CRediT author statement 505 Romina Avila: Research, Writing: original draft. 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