Contrasting sonochemical with electrochemical and catalytic processes for degrading representative pharmaceuticals in synthetic urine – Treatment of binary mixtures, theoretical calculations, and life cycle assessment Yudy L. Martínez-Mena , Javier Silva-Agredo , Marcela Paredes-Laverde , Ricardo A. Torres-Palma , Efraím A. Serna-Galvis * Grupo de Investigación en Remediación Ambiental y Biocatálisis (GIRAB), Instituto de Química, Facultad de Ciencias Exactas y Naturales, Universidad de Antioquia UdeA, Calle 70 No. 52-21, Medellín, Colombia A R T I C L E I N F O Keywords: Advanced oxidation processes Electrochemistry Fenton reaction Pharmaceuticals degradation Ultrasound Urine matrix A B S T R A C T Active pharmaceutical ingredients (API) in urine are recognized as a primary source of environmental contamination. Thereby, the effective treatment of urine could limit the pollution by pharmaceuticals. Herein, three advanced oxidation processes (a sonochemical, a catalytic, and an electrochemical system) to deal with API in synthetic fresh urine were contrasted. Acetaminophen, losartan, and levofloxacin were considered as repre sentative pharmaceuticals. Effects of operational parameters and the main species responsible for API removal were established. API degradation at a frequency of 582 kHz in the sonochemical process was favored, while 500 µmol L− 1 of H2O2 and 90 µmol L− 1 of Fe2+ were proper conditions for the catalytic process (Fenton). In turn, a low current (9 mA) and the presence of NaCl led to good degradation by the electrochemical process employing a BDD anode. Also, it was found that HO• played the principal degrading role in the three systems. Theoretical calculations allowed for the identification of the moieties on the API susceptible to transformations by HO•. The treatments in the urine matrix showed that the sonochemical system was more selective than Fenton and electrochemical processes for degrading API. The treatment of binary mixtures of API in urine evidenced more superiority and selectivity of sonochemistry for degrading hydrophobic pharmaceuticals. The sonochemical process transformed API into simple structures having less phytotoxic effects than the parent pharmaceutical or initial by-products. Moreover, a brief life cycle analysis on sonochemical treatment revealed that the highest environmental impact is attributed to the electricity required for operating the ultrasound. 1. Introduction Pharmaceuticals are widely used worldwide for the treatment of diseases and control of other biological processes (e.g., contraception). After being consumed, the pharmaceuticals experience the pharmaco kinetic process recognized by the acronym ADME (i.e., absorption, distribution, metabolism, and excretion) [1]. Thus, a portion of the ingested pharmaceuticals are excreted from the human body mainly through urine, or a smaller percentage in feces [2,3]. In fact, urine is one of the chief media for the pharmaceuticals input to the sewage systems [4], in places where they are available; and, in places where they are not, it is directly discharged into natural water sources (rivers, lakes, or waterfalls). Among the active pharmaceutical ingredients (APIs) frequently detected/found in municipal sewage systems are the highly consumed analgesics (e.g., acetaminophen [5]), sartan-type antihypertensives (e. g., losartan [5]), and antibiotics such as those belonging to the fluo roquinolones class (e.g., levofloxacin [6]). Those APIs are not completely eliminated by conventional methods (such as filtration, sedimentation, coagulation, and aerobic or anaerobic biological pro cesses) in municipal wastewater treatment plants (WWTP) [5]. For instance, Botero-Coy et al. (2018) evaluated the presence of acetamin ophen, losartan, some fluoroquinolones, and other pharmaceuticals in influents and effluents of wastewater treatment plants in Colombia. After conventional treatments at the WWTPs, the presence of acet aminophen changed from 39.25 µg L− 1 to 29.66 µg L− 1, losartan from 2.18 µg L− 1 to 1.97 µg L− 1, and in the case of fluoroquinolones, changes were reported from 2.29 µg L− 1 to 0.81 µg L− 1. It can be noted that the * Corresponding author. E-mail address: efrain.serna@udea.edu.co (E.A. Serna-Galvis). Contents lists available at ScienceDirect Journal of Environmental Chemical Engineering journal homepage: www.elsevier.com/locate/jece https://doi.org/10.1016/j.jece.2025.118295 Received 9 May 2025; Received in revised form 14 July 2025; Accepted 24 July 2025 Journal of Environmental Chemical Engineering 13 (2025) 118295 Available online 25 July 2025 2213-3437/© 2025 The Author(s). Published by Elsevier Ltd. This is an open access article under the CC BY-NC license ( http://creativecommons.org/licenses/by- nc/4.0/ ). https://orcid.org/0000-0002-8230-0917 https://orcid.org/0000-0002-8230-0917 https://orcid.org/0000-0002-7275-8000 https://orcid.org/0000-0002-7275-8000 https://orcid.org/0000-0003-4583-9849 https://orcid.org/0000-0003-4583-9849 mailto:efrain.serna@udea.edu.co www.sciencedirect.com/science/journal/22133437 https://www.elsevier.com/locate/jece https://doi.org/10.1016/j.jece.2025.118295 https://doi.org/10.1016/j.jece.2025.118295 http://creativecommons.org/licenses/by-nc/4.0/ http://creativecommons.org/licenses/by-nc/4.0/ concentrations of pharmaceutical pollutants did not decrease completely in the treated wastewater, thus evidencing little efficiency of conventional processes [5]. Therefore, APIs enter the environment, where they can experience bioaccumulation and biomagnification through aquatic species, in addition to the induction of toxic effects for diverse environmental species [7]. In this regard, reproductive prob lems, cardiac abnormalities, and liver and kidney damage have been reported in fish due to analgesic medications (e.g., acetaminophen, ibuprofen, diclofenac) [8]. Regarding antihypertensives (e.g., losartan), acute and chronic damage (cardiac, liver, and metabolic) has been found, which can lead to the death of aquatic species [9]. Meanwhile, in the case of antibiotics (e.g., levofloxacin), their environmental input can contribute to the development and proliferation of antibiotic-resistant bacteria, which is a major public health problem [7]. All these prob lems require the search for alternative solutions to decrease the entry of pharmaceuticals into the environment. Thus, the treatment of APIs in the initial pollution sources, such as urine, could be an option. The literature informs the application of advanced oxidation pro cesses (AOPs, which are based on the use of degrading radicals) for removing APIs in urine. These AOPs include homogeneous or hetero geneous processes based on photochemical, catalytic, and electro chemical systems, and, to a lesser extent, other systems such as ultrasound [4,10–18]. Such works are mainly focused on the individual treatment of diverse APIs. Most of them lack aspects such as the com parison with other AOPs or the evaluation of the environmental impacts of the AOPs through life cycle assessment. Moreover, a few reports show the treatment of pharmaceutical mixtures in the urine matrix [19]. Hence, this research covers those missed aspects. Herein, we considered three AOPs: sonochemistry, electrochemistry, and Fenton processes. Then, it should be mentioned that high-frequency ultrasound (200–1000 kHz) can be used for the generation of highly reactive species (such as HO•) from the acoustic cavitation phenomenon that cleavages water vapor molecules (Eq. 1). Also, in this AOP, in addition to the generation of HO•, the radical combination can occur, leading to the production of hydrogen peroxide (Eq. 2). So, H2O2 can be considered an indicator of the generation of radicals in the process. The sonogenerated HO• can degrade organic contaminants (Eq. 3), leading to innocuous products [4,20]. H2O + ))) → H• + HO• (1) 2HO• → H2O2 (2) HO• + Contaminants → Transformation products (3) Among catalytic AOPs, we can find the Fenton process, which is based on the reaction between hydrogen peroxide (H2O2) and ferrous ions (Fe2+) to produce radical species such as the hydroxyl radical (HO•) (Eq. 4). Other radical species such as the hydroperoxyl radical (HOO•), can be produced in the Fenton process by the interaction of Fe3+ with H2O2 (Eq. 5) [21,22]. Fe2+ + H2O2 → Fe3+ + HO• + HO- (4) Fe3+ + H2O2 → Fe2+ + HOO• + H+ (5) The hydroxyl radical can also be produced by electro-oxidation of water. Electrochemical AOPs are termed clean technologies in which green reagents, such as electrons, participate in the HO• formation from the water molecule. However, it should be considered that the genera tion of HO• radicals is dependent on the type of anode used in the process. Boron-doped diamond (BDD) anodes are typical for producing HO• electrochemically since they have a high overpotential (1.59–2.80 V) for oxygen evolution [23,24]. BDD anodes possess the capacity to generate physisorbed radical species (Eq. 6). Additionally, if chloride anions are present in the aqueous solution, BDD has also the tendency to generate other non-radical oxidizing species such as hypo chlorous acid (HClO) (Eqs. 7–8) [25], which can also act as degrading agents toward organic pollutants (Eq. 9) [25–27]. BDD anodes + H2O → BDD(HO•)ads + H+ + e- (6) 2Cl- → Cl2 + 2e- (7) Cl2 + H2O → HClO + HCl (8) Cl2/HClO + Contaminants → Transformation by-products (9) As mentioned above, this research studied the degradation of phar maceuticals in synthetic fresh urine by three different POAs (catalytic, electrochemical, and sonochemical systems). Three different pharma ceuticals were considered: an analgesic (acetaminophen), an antihy pertensive (losartan), and an antibiotic (levofloxacin); some of their physicochemical properties are shown in Table 1. Initially, the effect of parameters and degradation routes in the processes was established. Then, the degradations by the three AOPs were compared, showing ul trasound to have the highest selectivity for the degradation in urine. Thereby, this process was used to deal with the binary mixtures of the target pharmaceuticals. Besides, to better understand the action of sonogenerated HO• on the target APIs, theoretical calculations about the regions susceptible to the degrading species attacks were carried out. Also, a phytotoxicity test for the urine treated by ultrasound was per formed to evaluate the potential reuse for crop irrigation. Finally, the evaluation of the environmental impacts of the sonochemical system was determined through a life cycle assessment. Values taken from the references [14,28,29]. 2. Materials and methods 2.1. Reagents and samples Acetaminophen (ACE) from Laproff (Medellín, Colombia), losartan (LOS) from La Santé S.A. (Bogotá, Colombia), and levofloxacin (LEV) from Chemo-laboratories (São Paulo, Brazil) were used. Commercial tomato seeds (Solanum lycopersicum) of ANASAC® (Santiago de Chile, Chile) were used for the phytotoxicity test. For the preparation of syn thetic urine, substances such as urea, sodium chloride, potassium chlo ride, calcium chloride dihydrate, magnesium chloride hexahydrate, sodium phosphate, and D-glucose from Merck and sodium sulfate from Fisher Chemical were used. In addition, methyl orange from Merck was used. Iron sulfate heptahydrate (as the Fe (II) source for the Fenton re action) was purchased from Sigma. The pH of distilled water (DW) and synthetic urine was adjusted to 6.5 with sodium hydroxide and sulfuric acid as needed. For the experiments, distilled water (DW) was used in the prepara tion of the synthetic urine matrix according to the composition indicated in Table 2 (which was adapted from reference [10]). The concentration of all active pharmaceutical ingredients for the degradation tests cor responded to 66.2 µmol L− 1 (amount at mg L− 1 levels), which is a concentration that could be found in human urine typically [10,30,31]. 2.2. Reaction systems For the sonochemical process, a Meinhardt-Ultrasonics® bath apparatus was used, carrying out 200 mL of the liquid sample in a glass cylindrical reactor. The reaction temperature was controlled and Table 1 Physicochemical properties of the target APIs. Properties ACETAMINOPHEN LOSARTAN LEVOFLOXACIN Water solubility in the molecular form (mg mL− 1) 14.0 < 1.0 1.44 pKa 9.46 5.5 6.25 LogKo/w (LogP) 0.46 4.01 − 0.4 Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 2 maintained at 20◦C utilizing a Huber Minichiller [4]. This sonochemical reactor allowed for varying the frequency and power. The actual ultra sound power was determined by the calorimetric method according to Kimura et al. [32]. In this system, the effect of frequency and power parameters on the APIs degradation was assessed initially. In the case of the Fenton process, a volume of 50 mL of sample was placed in a beaker. The concentrations of reagents (iron and hydrogen peroxide) were varied to determine their effects on the degradation of pollutants. For the electrochemical system, an undivided electrolytic cell, with 150 mL of sample and magnetic stirring, was used. The system was composed of a boron-doped diamond anode (BBD, 1.9 cm2) and a titanium cathode (Ti, 4.25 cm2). In this process, the effects of the current density and supporting electrolyte type were considered. It is relevant to mention that as the reaction systems involve different sample volumes, the comparison of the process was done in terms of a parameter that is independent of the volume of the reactor, which is the selectivity factor (ρ). This selectivity factor (ρ = %Removal in urine/% Removal in distilled water) is influenced by the intrinsic properties of the processes (e.g., their ability to produce degrading species such as radicals and the action routes involved in the pollutants degradation) [4]. 2.3. Analyses The degradation of the target APIs was followed by high- performance liquid chromatography (HPLC), using equipment with the following specifications: Thermo Scientific Dionex UltiMate 3000 HPLC with diode array detector and an Agilent 120 RP C18 column (5 µm, 4.6 ×150 mm). The mobile phases used were: formic acid buffer with acetonitrile in a ratio 85/15 for acetaminophen at a flow rate of 0.45 mL min− 1 and a wavelength (λ) of 243 nm; the same buffer, acetonitrile and methanol (MeOH) 46/44/10 for losartan at a flow rate of 0.7 mL min− 1 and λ of 230 nm; and, formic acid buffer with aceto nitrile in a ratio 85/15 for LEV at a flow rate of 1.0 mL min− 1 at a λ of 278 nm. A volume of 20 µL was injected. Experiments were performed at least in duplicate. A Shimadzu TOC-L Total Organic Carbon Analyzer Instrument (using the platinum combustion technique) was utilized for the mineralization analysis, following the procedure of the Standard Methods SM5310. For the phytotoxicity test, commercial tomato (Solanum lycopersi cum) seeds were subjected to four manual washing cycles for 15 min (the first with 2 % hydrogen peroxide solution and the other three with distilled water) to eliminate residues present on the seed surface. Then, the seeds were left to dry on a paper towel and, in different Petri dishes (60 mm diameter), 7.0 mL of the samples were added, taken at times 0T (non-treated), 1T (100 % removal of the contaminant) and 2 T (double the time T), respectively, with previous adjustment of pH to 7.0. In each petri dish with the sample, 5 seeds were added and distributed as uniformly as possible, and growth was expected for 7 days. Finally, the length of each germinated seed was measured (considering radicle and hypocotyl). Positive control samples in distilled water and a negative control with glyphosate were included. The experiments were per formed in duplicate. The life cycle analysis (LCA) was carried out using OpenLCA 2.1 software and the Ecoinvent v 3.7 database, with which the inventory was constructed. This approach was carried out considering the work of Paredes-Laverde et. al. [36]. Meanwhile, the identification of the moi eties on the target APIs more susceptible to radical attacks was done through theoretical calculations of the f0 (Fukui index) [37]. Such cal culations were carried out on the Software Rowan available online. This software utilizes the xTB family of semi-empirical methods (geometry optimization: GFN2-xTB, and level of theory: GFN1-xTB) developed by Stefan Grimme and co-workers at the University of Bonn [38]. All the specific details about the calculations can be found on the Rowan Sci entific platform at https://www.rowansci.com (accessed 2025–04–03). 3. Results and discussion 3.1. High-frequency ultrasound for degrading ACE The effects of frequency and ultrasound power for the treatment of ACE in distilled water (DW) were initially evaluated. Two frequency levels (582 kHz and 1144 kHz) and two power levels (7.6 W and 20.1 W) were studied (Fig. 1a). The effect of the ultrasound power (keeping the frequency at 582 kHz) was tested, showing that as the actual acoustic power augmented from 7.6 W to 20.1 W, the removal of ACE increased from 13.5 % to 55.3 % in 30 min of sonication. It should be considered that the generation of cavitation events (which lead to the production of degrading species) is directly proportional to the acoustic power. So, the higher the working power, the more ACE degradation is favored [39]. When the ultrasound frequency increased from 582 to 1144 kHz (operating at 20.1 W of acoustic power (Fig. 1a)), the degradation effi ciency diminished to 18 %. Such a result can be explained by consid ering that the predominance of the chemical effects that lead to the sonogeneration of degrading agents occurs in a frequency range be tween 200 and 600 kHz [39]. Some previous papers have shown that the degradation of pharmaceuticals like ACE by ultrasound is more efficient if the working frequency belongs to such a range [40,41]. Hence, based on our results, 582 kHz of frequency (at 20.1 W of actual acoustic power) was selected to perform the subsequent degradation experiments for the sonochemical system (SC). For the determination of the sonochemical degradation route, the accumulation of H2O2 in the presence and absence of ACE in distilled water was compared (Fig. 1b). Also, degradation experiments with ACE were performed using methanol [39]; which is a scavenger of hydroxyl radicals (kMeOH/HO•: 9.7 ×108 L mol− 1 s− 1, [42]). The comparative re sults of ACE degradation in the absence and presence of the scavenger are shown in Fig. 1c. It was evidenced that in the ACE presence, the hydrogen peroxide accumulation was lower than the absence (i.e., distilled water alone, Fig. 1b), this suggests the reaction of ACE with the sonogenerated hydroxyl radical. Furthermore, the degradation of the pharmaceutical was lowered by approximately 36 % by adding meth anol (MeOH, Fig. 1c), thus confirming the main degrading action of the hydroxyl radical in the sonochemical process [10]. 3.2. Degradation of ACE in DW by other oxidation processes 3.2.1. Catalytic process (Fenton) for degrading ACE 3.2.1.1. Parameters effects on the catalytic process. The degradation of ACE in distilled water was also tested by the catalytic process. In such a process, the effect of the concentration of the main reagents involved in the Fenton reaction was studied. From Fig. 2a, it is evident that when we Table 2 Composition of synthetic fresh urine. Substance Molecular weight (g mol− 1) Concentration (mol L− 1) Log P* Urea 60.1 0.2498 − 2.59 NaCl 58.4 0.0440 − 0.46 Na2SO4 142.0 0.0150 − 0.84 KCl 74.6 0.0400 − 0.46 MgCl2.6H2O 203.3 0.0040 0.61 NaH2PO4 119.9 0.0199 − 1.0 CaCl2.2H2O 147.0 0.0040 − 0.57 Glucose 180.2 0.0005 − 2.6 Methyl Orange** 327.3 0.000003 2.74 pH: 6.5 ± 0.1 Electrical conductivity: 14.1 mS cm− 1 **This compound was used as a model of a nitrogen-containing organic mole cule responsible for color in the synthetic urine. * Values taken from the references [10,33–35]; Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 3 https://www.rowansci.com used low concentrations of H2O2 or Fe2+, less than 20 % of the phar maceutical degradation was achieved. Nevertheless, when the reagent concentrations were increased up to 90 µmol L− 1 Fe2+ and 500 µmol L− 1 H2O2, in 5 min of treatment, the ACE was completely decreased. This may be associated with the fact that the concentration increase of the reagents is directly proportional to the production of species that exert degradative action on ACE [43]. Thus, based on the results in Fig. 2a, the conditions of 90 µmol L− 1 Fe2+ and 500 µmol L− 1 H2O2 were selected for the subsequent experiments. 3.2.1.2. Routes involved in the sonochemical process. To establish the degradation routes involved in the catalytic process, control experi ments for ACE degradation, using each component of the Fenton system (i.e., H2O2 and Fe2+ acting individually), were carried out (Fig. 2b). By performing these experiments, it was observed that in 5 min of treat ment, iron alone has no significant effect on degradation. Meanwhile, hydrogen peroxide only removed 36.7 % of ACE (H2O2 is an oxidizing agent (E◦: 1.76 V [44]) capable of inducing some oxidation of organic compounds [45]). Fig. 1. Results of ACE degradation in DW by the sonochemical process (SC). (a) Evaluation of experimental parameters (P: acoustic power, f: frequency), (b) hydrogen peroxide accumulation in the absence and presence of ACE, and (c) treatment of ACE in the presence of the HO• scavenger (MeOH: 6600 µmol L− 1). Experimental conditions: [ACE]: 66.2 µmol L− 1, volume: 200 mL, and initial pH: 6.5. Fig. 2. ACE degradation in DW by the Fenton process. (a) Evaluation of experimental parameters and (b) degradation routes involved in the catalytic process. Experimental conditions: [ACE]: 66.2 µmol L− 1, volume: 50 mL, initial pH: 6.5. Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 4 These results for H2O2 and Fe2+ acting individually contrast with 100 % degradation achieved within 1 min of treatment when applying the Fenton system. In addition to the control experiments, the partici pation of the hydroxyl radicals in the Fenton process was verified by the treatment of ACE in the presence of methanol (at 6.6 mM), obtaining an evident inhibition of ACE degradation (Fig. 2b). Hence, the degradative effect in this catalytic system was associated with the main action of the generated HO• and moderate participation of the direct oxidation by H2O2. 3.2.2. Electrochemical process for the ACE degradation 3.2.2.1. Parameters in the electrochemical process. The electrochemical treatment of ACE in distilled water using an electrolytic cell with a BDD anode and a titanium cathode (BDD/Ti system) was performed. The effects of the type of supporting electrolyte and current density were studied. Sodium chloride (NaCl) and sodium sulfate (Na2SO4) at 0.1 mol L− 1 were individually tested in distilled water (DW) (Fig. 3a). It was observed that the highest degradation value was found when using NaCl, obtaining a 100 % elimination in 30 min of treatment, while only 20 % of ACE degradation was observed when Na2SO4 was used. In the presence of sulfate anion, the BDD anode can promote the formation of persulfate anion (which is also an oxidizing agent) [46,47]. Thus, the determination of the persulfate accumulated during the elec trolysis of the sulfate anion in the absence of ACE was carried out (persulfate quantification was done following the reference [48]. It can be mentioned that the persulfates formation is more favored at high concentrations of sulfate (> 1.0 mol L− 1) and high current densities (e. g., 100 mA cm− 2) [46,47]. Under our experimental conditions, a low current density (4.7 mA cm− 2) and a moderate sulfate concentration (0.1 mol L− 1), we found that only ~ 41 µmol L− 1 of persulfate was accumulated after 30 min of electrolysis. Thus, the contribution of electrogenerated persulfate to the ACE degradation is very low. The results in Fig. 3a suggest that in the NaCl solution, more degrading species can be formed. Then, using NaCl as a supporting electrolyte, the role of the current density on the ACE degradation was assessed (Fig. 3a). It was noted that the increase in the current density did not improve the pharmaceutical elimination. This last result can be associated with the enhancement of side parasitic electrochemical re actions, such as oxygen evolution, which does not influence the ACE degradation (Eqs. 10–11) [49]. Also, it should be taken into account that at high current densities, the BBD anode can promote the formation of highly oxidized chlorine species (e.g., chlorate and perchlorate anions). Such highly oxidized chlorine species could act as degrading species, but they are not reactive under low temperature (~25◦C) and pressure (~1 atm) conditions [31], as these were used in our work. Therefore, based on the above-mentioned aspects, the subsequent electrochemical tests were carried out at 4.7 mA cm− 2 of current density with sodium chloride. Acid solution: 2 H2O → O2 + 4 H+ + 4e- (10) Alkaline solution: 4 OH- → O2 + 2 H2O + 4e- (11) 3.2.2.2. Routes in the electrochemical process. In the electrochemical process involving BDD and NaCl, three routes are plausible (i. attacks of hydroxyl radical from the water cleavage (Eq. 6), ii. direct oxidation of the pollutant on the anode surface, and iii. the attacks of the electro generated Cl2/HClO species (Eqs. 7–8). Then, to elucidate the contri bution of the hydroxyl radical to the ACE, the API treatment using methanol as an HO• scavenger (Fig. 3b) was evaluated. It was found that the elimination of ACE in DW decreased by more than 50 % when methanol was added. This indicated the participation of radical species in the process. Additionally, the results when Na2SO4 was used as a supporting electrolyte indicated that the direct oxidation of ACE on the BDD surface (Eq. 12, [50]) was low (~15 %, Fig. 3a). ACE + BDDsurface → By-products + BDD + e- (12) HClO + H2O ⇆ H3O+ + ClO- (13) The ACE degradation in the scavenger presence was not completely inhibited (Fig. 3b), indicating that the electrogenerated Cl2/HClO spe cies also participated in the process. As mentioned above, the interaction of Cl- with the BDD surface produces active chlorine species, such as molecular chlorine, hypochlorous acid, and hypochlorite anion (Eqs. 7–8,13) [25]. These species can also attack organic pollutants such as ACE [30]. Thereby, in the tested electrochemical system, the main degradation routes were the attacks of hydroxyl radicals and chlorine species. 3.3. Matrix effects on the degradation of ACE 3.3.1. Effect of synthetic fresh urine matrix on the degradation of ACE The degradation of ACE by ultrasound in the urine matrix was assessed (Fig. 4a). For ACE, a 55.26 % removal was obtained in distilled water (DW), while, in the urine matrix, the removal percentage obtained was 36.40 %. Degradation of ACE decreased when changing from DW to the more complex urine matrix. The effectiveness of the sonochemical system is related to the hydrophobic character of the treated substances, and high hydrophobicity allows them to be located near the interface of the cavitation microbubbles, which favors their interaction with the HO• Fig. 3. ACE treatment in DW by the electrochemical system. (a) Evaluation of experimental parameters and (b) Determination of the degradation routes. Experi mental conditions: [ACE]: 66.2 µmol L− 1, volume: 150 mL, and initial pH: 6.5. Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 5 generated at the moment of the bubble implosion [20,51]. Thus, although the hydroxyl radical is not selective, the sonochemical process applied has a certain selectivity against hydrophobic compounds; therefore, degradation is more affected by the more hydrophilic the substance treated is. In the case of ACE, since it is moderately hydro phobic according to its Log P (0.46), it would probably not be found in the interfacial zone; thus, a partial inhibition in degradation occurs in the urine matrix [28,41]. The degradation experiments of ACE in the urine matrix were carried out, using the Fenton process, and the results are presented in Fig. 4b. It is observed that when moving treatment to a more complex matrix such as urine, the degradation of the pharmaceutical is significantly inhibited, obtaining removal percentages lower than 10 %. This could be explained because the hydroxyl radical, a species that acts in the Fenton process, is not selective, so when changing to a complex sample such as urine, competition for the radicals between the urine compo nents and the pharmaceuticals occurs. The non-selective hydroxyl radical generated from the Fenton reaction can attack components of synthetic urine, such as urea, glucose, methyl orange, and even inor ganic ions (e.g., Cl-). Therefore, competition among the target phar maceutical and the substances that constitute the study matrix occurs, which can be related to the kinetic constants and concentration factors shown in Table 3 [4]. ACE degradation experiments in urine by the electrochemical Fig. 4. Urine matrix effect on the degradation of ACE. (a) Sonodegradation (SC) of ACE in urine, (b) degradation in urine by Fenton, (c) electrochemical process (EC) with BDD anode, (d) TOC evolution under the different processes, and (e) comparison of the selectivity parameter (ρ) in the removal of ACE in urine among the studied AOPs. Experimental conditions: as described in Figs. 1–3 previously. Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 6 process were carried out. At 30 min of treatment, less than 30 % of degradation was obtained (Fig. 4c). This contrasts with the treatment in DW, which had a contaminant elimination higher than 90 %. This is due to the urine components competing with the contaminant in the inter action with HO• and chlorine species [4,39]. For instance, urea degra dation can occur through the reaction with HO•, since this interaction has a high kinetic constant (kHO•-urea= 7.9 ×105 L mol− 1 s− 1) [10]. Moreover, the mediated route can also be affected by the urea presence, since it has high reactivity with chlorine species such as HClO (Eq. 8), which predominates at the working pH [52]. (NH2)2CO + 3HClO → CO2 + N2 + 3Cl- + 2H2O + 3H+ (14) On the other hand, to evaluate the mineralization, the total organic carbon (TOC) was analyzed for ACE in the urine samples treated by the three processes (Fig. 4d). It can be noted that none processes induced mineralization. This can be explained considering that the short treat ment time (i.e., 30 min, which is the focus of our work), there is not enough time to promote the mineralization of the pharmaceutical or other urine matrix components. Then, to better compare the three AOPs at short treatment time, the selectivity index (ρ, Removal in urine/ removal in DW (%)) was used. The results for the selectivity index for the three processes are shown in Fig. 4e. It was observed that the sonochemical (SC) process shows lower matrix interference thanks to urine components being so hydro philic ([10], Table 2). Thus, matrix components are less likely to react with the HO• generated, compared to the organic pharmaceutical under study. The Fenton (F) and electrochemical (EC) processes showed the lowest selectivity. This evidenced the great potential of the sonochem ical AOP for ACE degradation in urine. Therefore, the subsequent ex periments were focused on this last AOP. The present study focused on the approach to the initial stage of removal of APIs by the three processes (sonochemical, electrochemical and Fenton), so the comparison between them was based on the capacity of the reactors in terms of selectivity, rather than on kinetic aspects; for this reason, the difference between the volumes used in the different systems does not interfere in the approach given to the results. 3.3.2. Comparison of the sonochemical degradation of ACE with other relevant pharmaceuticals To determine the role of the pharmaceutical nature on the sonode gradation process, two other pharmaceuticals were considered; i.e., losartan (LOS) and levofloxacin (LEV). These compounds were selected considering that they belong to three relevant pharmaceutical classes and have diverse functional groups. The sonodegradation of the three active pharmaceutical ingredients (i.e., ACE, LOS, and LEV) in distilled water is compared in Fig. 5a. From this figure was noted that after 30 min of treatment, the degradation of the target APIs followed the order: LOS > ACE > LEV. The observed order for the sonodegradation can be linked to the hydrophobicity of pharmaceuticals, which is directly linked to the octanol/water partition coefficient (Log P) [10]. For LOS, ACE, and LEV, the Log P values are 4.01, 0.46, and − 0.4, respectively [10,28,29]. Hence, the antihypertensive is the most hydrophobic, while the anti biotic is the most hydrophilic compound. Considering this, the results in Fig. 5a make sense, since the most hydrophobic API (i.e., LOS) has more possibilities of being degraded by the sonochemical process (followed by ACE) because it is closer to the bubble interface, and consequently closer to the sonogenerated radicals [10]. On the other hand, as shown in Section 3.1 above, the main degrading species involved in the sonochemical process is the hydroxyl radical. Hence, to better understand the action of such a species on the target APIs, theoretical calculations about the regions susceptible to HO• attacks were carried out. Herein, the Fukui function for radical in teractions (f0) for ACE, LOS, and LEV was calculated [38]. The f0 index brings information about the moieties of the organic compounds more reactive to radical species [37]. Table 4 presents the f0 index values for the three target APIs. The results for ACE in Table 4 show that the central amide (f0: 0.058) and the ortho-positions (f0: 0.062) to the hydroxyl moiety on the ben zene ring of ACE had high f0 index values (i.e., these are the electron-rich Table 3 Rate constants between HO• and the components of urine and concentration factors. Substance M (mol L− 1) Concentration factor (substance/ACE) k2nd HO• - substance* ACE 66.2 × 10− 6 1 ~1.78 × 109 L mol− 1 s− 1 Urea 0.2498 3776.3 7.9 × 105 L mol− 1 s− 1 Cl- 0.1000 1510.6 4.3 × 109 L mol− 1 s− 1 SO4 2- 0.0150 226.8 6.5 × 102 L mol− 1 s− 1 H2PO4 - 0.0199 300.8 ~2.0 × 104 L mol− 1 s− 1 Glucose 0.0005 7.6 ~2.0 × 109 L mol− 1 s− 1 Methyl Orange** 3.0 × 10− 6 0.05 ~2.0 × 1010 L mol− 1 s− 1 * Rate constants were taken from the references [11,53–55]; **model com pound as a source of nitrogen and color for the urine matrix. Fig. 5. Treatment of the three APIs by the sonochemical process. (a) APIs degradation by sonochemical AOP in DW at 30 min of treatment, and (b) APIs degradation by sonochemical AOP in urine at 30 min of treatment. Experi mental conditions: as described in Fig. 1. Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 7 regions). Thus, such moieties can be more easily transformed by the action of the sonogenerated hydroxyl radical. Indeed, the literature re ports that acetaminophen can undergo the cleavage of the central amide and hydroxylation of the aromatic ring at the ortho position to the initial OH by the initial HO• action in distilled water [56–58]. In the case of LOS, the theoretical analyses revealed that atoms in the imidazole group, aromatic rings, and tetrazole moiety have the highest values for the f0 index. Thereby, such structures on LOS are very sus ceptible to the attacks of HO•. Thereby, the sonochemical treatment of LOS in distilled water can generate stable transformation products coming from imidazole ring cleavage and hydroxylations of the biphenyl moiety [10,59]. Meanwhile, high f0 values for LEV are placed on the nitrogen atoms (0.041–0.056) of the piperazyl ring and other atoms on the morpholine moiety (0.027–0.043), in addition to the double bonds (0.040–0.049) on the quinolone core; thus, suggesting the involvement of such atoms in the primary transformations of LEV in distilled water. The literature has informed that the piperazyl moiety of the fluo roquinolone antibiotics is very reactive toward HO•, which produces demethylation and intermediates of opened rings [60–62]. Also, the C––C bond close to the carboxylic acid can experience modifications by the HO• attacks [61]. Regarding the treatment of the three pharmaceuticals in the urine matrix (Fig. 5b), we should mention that the sonochemical process led to degradations between 24.3 % and 41.5 % at 30 min of treatment. As mentioned above (Table 1 and Table 2), the target pollutants are more hydrophobic than the matrix components; thus, the sonochemical sys tem was able to achieve high degradation percentages. Also, it is important to highlight that our results agree with those found by Gua teque-Londoño et al. (2020), who also evaluated the degradation of LOS, evidencing a favorable sonochemical elimination of this pharmaceutical thanks to its high hydrophobicity [10]. These results confirm the rele vance of hydrophobicity on the sonochemical treatment of pharma ceuticals and compounds, such as ACE, which exhibit a medium hydrophobic character, and have the chance to be degraded. 3.3.3. Sonochemical degradation of binary pharmaceutical mixtures As the main focus of this research was on the selectivity at the pri mary stages of the processes (more than only the degradation percent age), and the sonochemical process showed a higher/good performance in term of the selectivity at short treatment times for the target phar maceuticals individually (see previous section), the degradation exper iments for rational binary mixtures among ACE, LOS, and LEV, in the urine matrix, were performed using this process. The results for the treatment of binary mixtures by ultrasound are shown in Fig. 6, where it is observed that the removal of LOS was still higher compared to ACE or LEV, respectively (Figs. 6a-6b). Additionally, it was observed that ACE and LEV removals were lower when found mixture compared to their individual treatment; whereas, LOS is favored when found in mixtures, as its removal was accelerated. Furthermore, the selectivity factor for the APIs degradation in the mixtures was evaluated (it was defined as β the ratio between the percentage removal of API in binary mixtures and the removal alone in solution, Fig. 6d). This shows that the applied sonochemical system also presents greater selectivity for LOS degradation in the binary mixtures in urine (Fig. 6d). The explanations of results in Fig. 6 are also based on the diverse hydrophobic/hydrophilic nature of the target pharmaceuticals. As LOS is more hydrophobic than ACE and LEV, in the mixture, the former is closer to the cavitation bubble interface, having a higher probability to interact with the sonogenerated HO•, thus affecting the ACE or LEV degradation. Furthermore, as ACE and LEV have more affinity for water than LOS, in the treatment of the binary mixtures, salting-out effects toward the antihypertensive are plausible. I.e., in the binary mixtures with LOS, the water molecules can surround ACE or LEV molecules more effectively, thus pushing LOS closer to the cavitation bubbles interface [10,39]. Otherwise, when ACE and LEV are in the mixture (Fig. 6c), both are placed far away from the interfacial zone of the cavitation bubble because of their less hydrophobic nature. Consequently, they compete for the hydroxyl radical and decrease the degradation in the individual case. 3.4. Treatment extent 3.4.1. Phytotoxicity Considering the results in Section 3.3, where LOS presented the highest removals by the sonochemical process, the subsequent tests were done with this pollutant. To go beyond the degradation (i.e., treatment extent), the assessment of the possibility of using the treated urine for crop irrigation was carried out by measuring phytotoxicity toward to mato seeds. For this purpose, the degradation of LOS was evaluated up to 100 % removal (this occurred after 180 min of treatment, which was termed T). Also, the non-treated solution (0 T) and that generated at 360 min of treatment (named 2 T) were considered. The results are shown in Fig. 7. We should mention that the urine matrix itself has some phytotoxicity due to its highly saline nature when compared with distilled water (DW) (Fig. 7a) [63]. Then, the test was performed with samples diluted at a 1/10 ratio. From Fig. 7b, it can be noted that at 1 T (180 min), the growth of the Table 4 Fukui index (f0) values for the three target APIs and their possible trans formations by the HO• attacks. API f0* Possible transformations ACE Cleavage of the central amide and hydroxylation of the aromatic ring at the ortho position to the initial OH [56–58]. LOS Imidazole ring cleavage and hydroxylations of the biphenyl moiety [10,59]. LEV Demethylation and opening of the piperazyl ring, in addition to hydroxylation and/or oxidation of the C––C bond close to the carboxylic acid [60–62]. * Higher f0 values on the API denote more reactive moieties toward HO•. Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 8 germinate length was lower than the non-treated sample (0 T); i.e., when 100 % of LOS was degraded, an increase in phytotoxicity was evidenced. This indicates that primary degradation by-products have a higher phytotoxicity than the parent pharmaceutical, so it is very important to be concerned about the transformation products generated in the process. Interestingly, when the treatment was 2 T (360 min), the resultant solution had a lower phytotoxicity than both the samples at 0 T and 1 T. Such results suggest that the primary by-products are trans formed by the sonochemical process into simple structures that have less phytotoxic effects than the parent pharmaceutical and its initial transformation products. Guateque et al. reported similar results for the treatment of LOS in DW, finding a decrease in phytotoxicity after twice the treatment time applied for the elimination of the active pharma ceutical ingredient [10]. Therefore, our results about the phytotoxicity diminution indicate a positive effect of the sonochemical treatment of LOS in the urine matrix. 3.4.2. Life cycle assessment To determine the environmental impact of the sonochemical process for degrading LOS in urine, an LCA was carried out. Regarding the effect Fig. 6. Removal of APIs in binary mixtures by sonochemical AOP: (a) LOS in mixture with ACE, (b) LOS in mixture with LEV, (c) ACE in mixture with LEV, and. (d) Selectivity to the API degradation in the binary mixtures. Experimental conditions: Concentration of LOS, ACE, LEV: 66.2 µmol L− 1; frequency: 582 kHz; Power: 20.1 W; volume: 200 mL. Fig. 7. Phytotoxicity test associated with LOS treatment by the sonochemical process. (a) Control experiments in the phytotoxicity test and (b) phytotoxicity results for LOS treated at times 0 T, 1 T, and 2 T. Experimental conditions: LOS: 66.2 µmol L− 1; urine samples diluted 1/10; tomato seeds; 7 days of germination. Y.L. Martínez-Mena et al. Journal of Environmental Chemical Engineering 13 (2025) 118295 9 of this study on the environment, the system and boundaries of the process were established based on the ISO 14040 standard. Subse quently, the inputs/outputs required in the ultrasound treatment of 0.200 L of synthetic urine containing LOS were included in the in ventory (see Fig. 8 and Table 5). The calculations were done using the OpenLCA 2.1 software and the Ecoinvent v 3.7 database, and the process impacts were determined for 11 environmental impact categories using the Characterization factors for life cycle impact assessment baseline approach [36,64]. The environmental impacts of the process on global warming (GWP 100a), abiotic depletion, abiotic depletion (fossil fuels), acidification, eutrophication, freshwater aquatic ecotoxicity, human toxicity, marine aquatic ecotoxicity, ozone layer depletion (ODP), photochemical oxidation, and terrestrial ecotoxicity are reported in Table 6. The results for all evaluated categories indicated that the highest impact is almost 100 % attributed to the electricity required for the operation of the ul trasound equipment (Table 6). Consequently, these results motivate the scientific community to search for renewable energy alternatives for the green operation of ultrasound (e.g., solar light/solar panels) for the treatment of pharmaceuticals in complex matrices such as urine. 4. Conclusions From the development of this research can be concluded that: -The operational conditions that favored the API degradation by the sonochemical, catalytic (Fenton system), and electrochemical processes were those that increased the hydroxyl radical generation. -From the theoretical calculations, the central amide and aromatic ring on ACE, the imidazole and byphenyl moieties on LOS, and the piperazyl ring plus the morpholine structure on LEV were recognized as the moieties more susceptible to transformations by the radical attacks due to they are the electron-rich regions on the API. -The sonochemical system was more selective than the Fenton and electrochemical processes for degrading the target API, because the system based on ultrasound allows matrix components differentiation depending on the hydrophobic nature of the substances. -The treatment of binary mixtures of API in the urine matrix showed the high suitability of the sonochemical process to deal with hydro phobic pharmaceuticals in complex samples. -After a long treatment time, the sonochemical process was able to transform LOS into structures less phytotoxic than the parent pharma ceutical and its initial transformation products, thus denoting the rele vance of going beyond the decrease in the simple API concentration during the process action. -The life cycle assessment on the sonochemical treatment uncovered the highest environmental impacts of the electricity required for the operation of the ultrasound equipment, suggesting the need for searching for renewable energy alternatives for the green operation of this process to treat pharmaceuticals in complex matrices such as urine. -Finally, it is very important to remark that this work presented an initial comparison of three advanced oxidation processes through the approach of selectivity (as a key factor at short treatment times, beyond than the reaction kinetics), thus opening the possibility for future works, where a more exhaustive evaluation of each process at long treatment times, considering degradation by-products, toxicity, and mineralization level must be performed to provide a wider comparison of the envi ronmental relevance and safety of the three tested processes. CRediT authorship contribution statement Efraím A. Serna-Galvis: Writing – review & editing, Supervision, Project administration, Funding acquisition, Formal analysis, Concep tualization. Ricardo A. Torres-Palma: Writing – review & editing, Resources, Conceptualization. Marcela Paredes-Laverde: Visualiza tion, Methodology, Formal analysis. Javier Silva-Agredo: Writing – review & editing, Supervision, Methodology, Conceptualization. Yudy L. Martínez-Mena: Writing – original draft, Methodology, Investigation. Declaration of Competing Interest The authors declare that they have no known competing financial Fig. 8. Schematic representation of ultrasonic treatment for LOS degradation considered for life cycle analysis (gate to gate). Table 5 Inventory of inflows and outflows considered in the ultrasonic treatment of 0.200 L of LOS-contaminated urine. Flow Unit Sonochemical treatment Inputs ​ ​ Electricity M.J. 6.2 Synthetic urine (composed of inorganic salts) L 0.2 Pharmaceutical (LOS) kg 6.1 × 10− 6 Outputs ​ ​ Treated synthetic urine L 0.2 Table 6 Life cycle analysis for the sonochemical process. Impact group* Units Result of the impact assessment Percentage contribution Global warming (GWP 100a) Kg CO2eq 6.4 Electricity: 100 % Abiotic depletion Kg Sb equivalent 2.5 × 10− 9 Abiotic depletion (fossil fuels) M.J. 77.8 Acidification kg SO2 equivalent 0.03 Eutrophication kg PO4 equivalent 0.002 Aquatic ecotoxicity in freshwater kg 1,4-DB equivalent 0.01 Human toxicity kg 1,4-DB equivalent 0.6 Marine aquatic ecotoxicity kg 1,4-DB equivalent 833 Ozone layer depletion (ODP) kg CFC− 11 equivalent 4.2 × 10− 7 Photochemical oxidation kg C2H4 equivalent 0.002 Terrestrial ecotoxicity kg 1,4-DB equivalent 0.003 * Environmental assessment of the sonochemical treatment of LOS in synthetic urine in eleven impact categories using the reference method. Conditions: 0.200 L of synthetic urine contaminated with 6.1 mg of LOS, and 3 h of treatment. Y.L. Martínez-Mena et al. 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