https://orcid.org/0000-0001-8272-2258
Figures
Abstract
In this study, we utilized drinking water treatment sludge (WTS) to produce adsorbents through the drying and calcination process. These adsorbents were then evaluated for their ability to remove azithromycin (AZT) from aqueous solutions. The L-500 adsorbent, derived from the calcination (at 500°C) of WTS generated under conditions of low turbidity in the drinking water treatment plant, presented an increase in the specific surface area from 70.745 to 95.471 m2 g-1 and in the total pore volume from 0.154 to 0.211 cm3 g-1, which resulted in a significant AZT removal efficiency of 65% in distilled water after 60 min of treatment. In synthetic wastewater, the rate of AZT removal increased to 80%, in comparison, in a real effluent of a municipal wastewater treatment plant, an AZT removal of 56% was obtained. Kinetic studies revealed that the experimental data followed the pseudo-second-order model (R2: 0.993–0.999, APE: 0.07–1.30%, and Δq: 0.10–2.14%) suggesting that chemisorption is the limiting step in the adsorption using L-500. This finding aligns with FTIR analysis, which indicates that adsorption mechanisms involve π-π stacking, hydrogen bonding, and electrostatic interactions. The equilibrium data were analyzed using the nonlinear Langmuir, Freundlich, and Langmuir-Freundlich isotherms. The Langmuir-Freundlich model presented the best fitting (R2: 0.93, APE: 2.22%, and Δq: 0.06%) revealing numerous interactions and adsorption energies between AZT and L-500. This adsorbent showed a reduction of 19% in its AZT removal after four consecutive reuse cycles. In line with the circular economy principles, our study presents an interesting prospect for the reuse and valorization of WTS. This approach not only offers an effective adsorbent for AZT removal from water but also represents a significant step forward in advancing sustainable water treatment solutions within the framework of the circular economy.
Citation: Castro-Jiménez CC, Saldarriaga-Molina JC, García EF, Torres-Palma RA, Acelas N (2025) Azithromycin removal from water via adsorption on drinking water sludge-derived materials: Kinetics and isotherms studies. PLoS ONE 20(1): e0316487. https://doi.org/10.1371/journal.pone.0316487
Editor: Afzal Husain Khan, Jazan University, SAUDI ARABIA
Received: June 14, 2024; Accepted: December 11, 2024; Published: January 9, 2025
Copyright: © 2025 Castro-Jiménez et al. This is an open access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.
Data Availability: All relevant data are within the manuscript and its Supporting Information files.
Funding: The author(s) received no specific funding for this work.
Competing interests: The authors have declared that no competing interests exist.
Introduction
The United Nations strongly emphasizes the importance of promoting healthy lives for sustainable development. As a result, one of the key targets of Sustainable Development Goal -SDG- 3 (Good health and well-being) is to reduce the deaths and illnesses caused by water contamination [1]. In this context, antibiotics, classified as emerging contaminants (EC), have been frequently found in natural water sources, prompting widespread concern [2]. These pharmaceuticals are of particular concern due to the potential health and environmental risks associated with antimicrobial resistance [3–5].
Among the antibiotics, azithromycin (AZT) has been recognized as a priority substance on the second European Union watch list [6]. AZT belongs to the macrolide class, a group of antibiotics frequently detected in effluents, sewage sludge, sediments, soils, and food [7]. AZT is a broad-spectrum antibiotic used for various respiratory diseases, including COVID-19 [8–10], also AZT plays a crucial role in the medical field as a widely used antibiotic for treating various bacterial infections [11]. During the COVID-19 pandemic, pharmaceuticals like AZT were extensively used in treatment efforts [12]. This heightened usage resulted in increased excretion of these antibiotics into the environment via human waste [13].
Pharmacokinetic studies of AZT indicate that the body does not fully absorb it and that it can be excreted in its primary form into wastewater [12]. Additionally, no human metabolites or transformation products of this antibiotic have been identified [14]. Effluents from antibiotic manufacturing plants have been identified as another considerable source and pathway for the release of antibiotics like AZT, into aquatic environments [15].
When azithromycin enters aquatic environments, it can persist for extended periods, potentially affecting aquatic organisms and ecosystems. Its presence may also contribute to the emergence of antibiotic-resistant bacteria, posing threats to both human and animal health [16]. Moreover, the slow metabolism of AZT suggests that it may be poorly degraded in sewage treatment plants [17].
AZT has been detected in surface water sources in countries such as Spain (3 ng L-1) and China (4.3 ng L-1 to 935 μg L-1), as well as in surface water sediments (24 ng L-1) and groundwater (257 ng L-1) in Spain [18]. It has also been reported in raw and treated domestic wastewater in countries like the United States (raw: 18.3 μg L-1, treated: 3.25 μg L-1) and Colombia (raw: 5.24–7.00 μg L-1, treated: 3.02–4.66 μg L-1) [19]. In raw hospital wastewater, AZT concentrations range from 6.93 to 26.1 μg L-1 in Colombia [19] and 163 μg L-1 in Turkey [18]. The post-pandemic surge in AZT concentration in surface water, wastewater, and treated wastewater [18] and the ineffective conventional wastewater treatment methods for removing AZT [19] underline the urgency of developing alternative treatment approaches to effectively eliminate it from water [20].
At present, several treatments have been evaluated for removing AZT from aquatic systems, including adsorption [21–28], membrane filtration [29], advanced oxidation processes -AOPs- [30–33], and biological systems [34–38]. However, traditional methods like biological treatments face challenges due to their low efficiency and slow antibiotic decomposition rates [39]. Adsorption is considered as a highly effective, convenient, simple design, and cost-efficient method for eliminating pollutants from water without generating undesirable byproducts which can be produced in oxidation processes [2,8].
Adsorption is a surface process in which contaminants are transferred from a fluid bulk to the adsorbent surface through physical forces or chemical bonds [40]. Various adsorbents have been studied for the removal of AZT from water, including biochar, lignite, polyamide nanofibers [21], PAC (Powdered Activated Carbon) combined with Fe/Ag/Zn nanocomposites [22], activated porous carbon derived from the water fern Azolla filiculoides [23], mesoporous silica materials synthesized via the hydrothermal method [24], organoclays [25], natural clinoptilolite modified with surfactants [26], raw and modified nano-diatomites [27], and powdered zeolites [28].
Producing adsorbents from waste can enhance the process’s economic viability and exemplify waste valorization, aligning with the principles of the circular economy [41]. Among the waste materials that have garnered special interest for producing adsorbents are water treatment sludges (WTS). This waste is generated worldwide as a byproduct of drinking water treatment [42]. An average of 100,000 tons of WTS is estimated to be produced globally daily, which may triple in the coming decades [43]. Numerous studies have assessed adsorbents derived from WTS, particularly for removing metals, metalloids, phosphorus species, fluorides, and dyes across various water matrices [44].
However, water treatment sludges (WTS) have received relatively little exploration as a source material for antibiotic adsorption. To the best of our knowledge, the studies by He et al. [2], Punamiya et al. [45], and Saman [46] are the only ones to date that have reported on the adsorption of antibiotics from the tetracycline group by adsorbents made from WTS. Their findings highlight the need for further studies to explore the application of these adsorbents in real systems and to evaluated the effects of the adsorption process in complex matrices. To our knowledge, no studies have been reported on using adsorbents derived from WTS to remove the antibiotic AZT from water or complex matrices, such as simulated and real wastewater treatment plant effluents.
This study investigates the adsorption of the antibiotic AZT, using adsorbents produced by simple methods such as drying and calcinating of water treatment sludge (WTS), an easily obtained waste generated worldwide by drinking water treatment systems. Initially, three types of WTS were evaluated, each generated from raw water of varying qualities in the drinking water treatment plant. The key factors influencing AZT adsorption with the adsorbent that exhibited the most promising results were examined, including adsorbent dose, adsorption pH, the size of adsorbent particles and reuse. Furthermore, the experimental data were fitted to nonlinear adsorption kinetics models and adsorption isotherms for a more comprehensive understanding of the results. Additionally, AZT adsorption was studied in simulated and real municipal wastewater treatment plant effluents to assess the matrix impact. This study’s innovative use of WTS, combined with the detailed analysis of adsorption kinetics and isotherms, practical applicability in different water matrices, and alignment with circular economy principles, underscores the novelty and significance of the research.
Materials and methods
WTS sampling
Water treatment sludge samples (WTS) were collected from the settlers of a conventional drinking water treatment plant (DWTP) in Colombia. This DWTP employs aluminum sulfate as a coagulant for treating surface water. Three WTS samples were obtained, each corresponding to different turbidity levels in the source water and varying coagulant doses, as detailed in Table 1. WTS types were categorized based on a statistical analysis of raw water turbidity data from the DWTP spanning a 5 years period.
AZT powder was obtained from Zhejiang Guobang Pharmaceutical Co., Ltd., hydrochloric acid (HCl), sulfuric acid (H2SO4), sodium hydroxide (NaOH), sodium chloride (NaCl), ethanol (C2H6O) and methanol (CH₃OH) were obtained from Merck S.A. (Germany). In this study, distilled water was used for aqueous solutions.
WTS adsorbents preparation
Each WTS was initially dried in an oven at 100°C until a constant weight was achieved. Subsequently, they were ground and sieved (< 300 μm) to obtain dry WTS samples designated as L-100, M-100, and H-100. To assess the impact of calcination temperature, adsorbents were further prepared at 300°C (L-300), 500°C (L-500), and 700°C (L-700) using L-WTS. These calcinations were conducted in a furnace under an oxygen atmosphere for 2 h, with a heating rate of 10°C per minute. Finally, each calcined WTS sample was ground and sieved (< 300 μm) (Fig 1).
Characterization of WTS adsorbents
The specific surface area and pore structure of the absorbents were determined by the Brunauer, Emmett, and Teller (BET) method [47] with N2 adsorption at 77 K using an accelerated surface area and porosimeter system (Micromeritics, ASAP 2020 Plus). The surface chemical composition and morphology of the L-WTS were characterized by scanning electron microscopy (SEM, JEOL JSM 6490 LV) equipped with an Oxford energy-dispersive spectroscopy (EDS) system. Fourier transform infrared spectroscopy (FTIR, Spectrum-two, Perkin Elmer with UATR) was used to analyze the active functional groups of the samples for a range within 4000–450 cm−1. The zero-charge point pH (pHPZC) was calculated using the solids addition method [48]. The minerals in L-WTS were determined using an Empyrean 2012 X-ray diffractometer (Malvern-PANalytical) with copper [Cu, Kα = 0.15406 nm (1.5406 Å)] as the radiation source. Quantification was performed with the software HighScore Plus, using the Rietveld method and the ICSD FIZ Karlsruhe 2012–1 database.
Adsorption experiments
Initially, AZT removal was evaluated using adsorbents derived from three distinct types of WTS. Preliminary adsorption experiments were conducted with varying doses of the adsorbent. Each experiment involved 100 mL of a 50 mg L-1 (C0) azithromycin (AZT) solution, prepared in distilled water at a pH of approximately 7.0 (adjusted) and performed at room temperature (20–22°C) for 60 minutes (Fig 1A). The S1 Text details the AZT solution preparation procedure. The bottles were agitated at 200 rpm using an orbital shaker throughout the adsorption process. Samples were extracted at specific time intervals (t) and immediately filtered through a 0.45μm filter membrane (Fig 1B). The concentration of AZT (Ct) in the filtered samples was determined using a colorimetric method, as summarized in S2 Text [27,49–51]. The AZT removal was calculated according to Eq (1):
(1)
Where C0 (mg L-1) and Ct (mg L-1) represent the AZT concentrations at time 0 and t, respectively.
Considering the results of AZT removal and the prevalence of a specific type of WTS generated in the DWTP (as shown in Table 1), we chose L-WTS to investigate the impact of adsorbent preparation temperature on AZT removal (Fig 1C). Subsequently, using L-WTS calcined at 500°C (L-500) (Fig 1D), we explored the influence of pH (ranging from 3 to 12) and adsorbent particle size (<75, 75–150, 150–300, 300–600, and <300 μm) on AZT removal (Fig 1E). The adsorption capacity (qt) was determined using Eq (2).
Where m (g) represents the mass of the adsorbent, V (L) stands for the volume of the AZT solution, C0 (mg L-1) and Ct (mg L-1) have been previously defined.
Adsorption kinetic and isotherms models
At the optimal adsorption pH, the kinetics and isotherms of the adsorption process of AZT onto L-500 were evaluated (Fig 1F). In their nonlinear forms, the pseudo-first-order (PFO) and pseudo-second-order (PSO) models were employed to fit the kinetic experimental data. This choice was made because linearizing these models can introduce uncertainties and propagate errors, leading to inaccurate estimations of the model parameters [52,53]. Eqs (3) and (4) represent the kinetic models:
(3)
(4)
Where k1 (min-1) is the pseudo-first-order adsorption rate constant; qt (mg g-1) and qe (mg g-1) are the adsorption capacities at equilibrium and time t (min), respectively; k2 (g mg-1 min-1) is the pseudo-second-order adsorption rate constant. The constants of the kinetic models were calculated using the least squares model derived from the Rosenbrock-Newton optimization algorithm by Statistica software. Three isothermal models, Freundlich [54], Langmuir [55], and Langmuir–Freundlich [56–58] were applied to analyze the AZT adsorption on L-500 at the equilibrium. The mathematical models are presented in Table 2 (Eqs (5)–(8)). The coefficient correlation (R2), the average percentage error (APE) (S3 Text) [59], and the normalized standard deviation (Δq) (S3 Text) were used to validate the kinetics and isothermal models [60].
Where qe (mol kg-1) and Ce (mol L-1) are the AZT concentration and the adsorption capacity at equilibrium, respectively; KF (mol kg-1) and n are Freundlich constants representing the multilayer adsorption capacity and the adsorption intensity, respectively; qm (mol kg-1) is the maximum adsorption capacity, KL (L mol-1) is the Langmuir constant, RL is separation factor, and C0 (mol L-1) is the initial AZT concentration; Ka (L mol-1) is the adsorption affinity constant and nLF is the index of heterogeneity. RL serves to predict whether the adsorption process is favorable (RL < 1), unfavorable (RL > 1), irreversible (RL = 0), or linear (RL = 1) [61].
Reusability tests and complex matrix evaluation
The performance of the L-500 material was assessed over three reuse cycles under optimal conditions (Fig 1G). Batch systems were exposed to an AZT solution (100 mg L⁻¹) for 60 minutes with agitation at 200 rpm using an orbital shaker (Fig 1B). After the adsorption phase, the material was vacuum-filtered and treated with ultrasound using a Digital Pro (model: PS-30AL) for 30 minutes at 40 kHz. This treatment was performed in a mixture of 33.3 mL methanol (CH₃OH) and 6.6 mL of 3% sodium hydroxide (NaOH) solution. The material was then washed under vacuum with deionized water, dried at 100°C for 24 hours, and prepared for the next adsorption cycle [62].
Additionally, the AZT adsorption within a complex matrix was assessed by introducing 5 g of L-500 into a 100 mL synthetic wastewater solution spiked with AZT at a concentration of 100 mg L-1. This wastewater solution was prepared using the methods described by Paredes-Laverde et al. [63] and OECD [64], simulating the effluent of a municipal wastewater treatment plant (as detailed in S1 Table). Similarly, the adsorption of AZT was evaluated in a real effluent from a domestic wastewater treatment plant, using the same procedure as applied to the synthetic wastewater matrix (Fig 1G). All experiments were conducted at least by duplicate.
Results and discussion
Preliminary results
Fig 2A shows the AZT removal achieved with the three types of dry WTS by using several adsorbent doses (from 2.5 to 80 g L-1). It can be noted that similar AZT removal percentages were obtained for the dry WTS from sources with medium (M-100) and high turbidity (H-100), peaking at around 60% with an adsorbent dose of 80 g L-1. The highest AZT removal (~77% at the higher dose) was achieved with the WTS generated under conditions of low turbidity in the raw water (L-100). The AZT removals obtained with the three WTSs may be associated with the existence of the silica network functional groups observed from 465 to 1150 cm-1 in Fig 2B [65]. These functional groups have been reported as active sites of interaction with AZT through to the electrostatic interaction and H-bonding [28]. As the L-WTS had the highest AZT removal and is the most frequently produced sludge type at the DWTP, this WTS was selected for further investigation. The characterization of the L-WTS is presented in the following section.
(a) Effect of dried WTS doses on AZT removal, (b) FTIR spectra of L-100, M-100 and H-100. Experimental conditions: C 0: (a) 50 mg AZT L-1, contact time (t): 60 min, pH: 7.0, particle size (PS): <300 μm, temperature (T): 20–22°C.
L-WTS characterization
Fig 3A presents the thermogravimetric analysis of L-WTS. The analysis reveals that L-WTS undergoes a significant mass change up to approximately 200°C, attributed to the evaporation and hydration-induced loss of water, as Martins et al., [66] reported. Subsequently, between 200 and 600°C, a mass loss is attributed to the degradation of natural organic matter -NOM- [67]. Furthermore, the results from the X-ray diffraction pattern (Fig 3B) suggest that L-WTS is primarily amorphous (comprising 63.3%), with the few crystalline structures present being aluminum-silicate clays (kaolinite 1A, illite 2M1, and dickite 2M1), quartz, and low-level magnetite. These structures originate from particles in the raw water that are removed during water treatment.
(a) TGA analysis of L-WTS, (b) XRD pattern of L-WTS.
The surface chemical composition and morphology of L-WTS are depicted in Fig 4. Notably, particles with various sizes, irregular shapes, and rough texture can be observed. According to the results from EDS spectra, the surface of L-WTS is primarily composed of Si, Al, C, O, and Fe, commonly found in sludges from DWTPs [44].
Effect of thermal activation of L-WTS on AZT removal
The impact of the thermal activation of L-WTS on AZT adsorption by varying the adsorbent dosage was examined, as depicted in Fig 5A. Notably, adsorbents prepared at 500°C (L-500) and 700°C (L-700) exhibited the highest AZT removal percentages across various doses. Using these adsorbents, nearly complete removal of AZT (>99%) was achieved in 100 min with a dose of 100 g L-1. Calcination has been shown to enhance AZT adsorption, as reported by Saman et al. [46], who observed similar effects in the removal of cationic and anionic dyes as well as the antibiotics tetracycline and oxytetracycline. Considering its superior AZT removal efficiency and its lower energy for the adsorbent preparation compared to L-700, L-500 was selected for further adsorption experiments.
(a) Effect of thermal activation of L-WTS on AZT removal by varying the adsorbent dosage, (b) FTIR spectra of L-100, L-300, L-500, and L-700. Experimental conditions: C 0: (a) 100 mg AZT L-1, contact time (t): 60 min, pH: 7.0, particle size (PS): <300 μm, temperature (T): 20–22°C.
To elucidate the enhancement in AZT adsorption by L-500, the surface structure and functional groups of L-500 and L-100 were examined. Table 3 and S1 Fig exhibit the specific surface area and pore structures of the L-100 and L-500. Calcinating L-100 at 500°C resulted in a notable increase in its specific surface area (SBET) from 70.745 m2 g-1 to 95.471 m2 g-1. Similarly, it led to a 37% increase in the total pore volume of L-100, rising from 0.154 cm3 g-1 to 0.211 cm3 g-1. Both materials displayed a predominant pore size distribution centered around 3.5 nm, with an average pore diameter of around 9.0 nm, indicating a mesoporous structure (S1 Fig, inset). These findings align with previous reports on the thermal modification of WTS at similar temperatures (475–550°C) and highlight that the calcination of WTS can enhance active sites and pore diffusion capacity for pollutant adsorption [44,68].
The functional groups present in L-100, L-300, L500, and L-700 characterized by ATR-FTIR are depicted in Fig 5B. The band observed from 3020 to 3780 cm-1 in L-100 and L-300 corresponds to the OH stretching vibrations of sorbed water, mineral hydroxide phases, and organic hydroxide [68,69]. The decrease in intensity of this band in WTS observed after calcination at 500°C (L-500) and 700°C (L-700) can be attributed to increased decomposition of NOM and mineral dehydration [67]. NOM in the L-100 is associated with bands around 1640 cm−1 (C = C, C = O) and 1400 cm-1 (COO-, C-O). The presence of NOM in L-100 and L-300 may limit its adsorption capacity by competition with the target pollutant. The thermal treatment at 500°C and 700°C produces a loss of NOM, and the peaks corresponding to the bands 1640 cm−1 and 1400 cm−1 practically disappear in L-500 and L-700, which can enhance its adsorption capacity [67,69]. Other bands attributed to the silica network fall within the range of 465–1150 cm-1 [65]. These bands are related to the structures of the quartz and the clays identified in XRD analysis. The peaks at approximately 1030 cm−1, 532 cm−1, and 465 cm−1 are associated with Si–O vibrations, as indicated by Shamaki et al. [68]. These peaks persisted in L-500 and L-700, confirming the stability of the SiO2 phase following calcination. The band at 914 cm−1 is associated with the elongation of Al-OH bonds [66]. As demonstrated by the results of L-500 and L-700 samples, calcination reduces this band due to the dehydroxylation of clay minerals presented in L-100 and reported on XRD analysis [68]. The results of surface characterization for L-100 and L-500 suggest that the significant improvement in AZT removal observed with L-500 is primarily linked to enhancements in its surface structure (an increase in surface area and pore volume) and the NOM removal resulting from the thermal treatment at 500°C. As the AZM molecule is quite large, its adsorption onto L-100 and L-500 is supposed to primarily involve external surface interactions [28]. This is due to the limited ability of the molecule to access the adsorbent’s micropores. Also, these micropores only represent a small fraction of the total pore volume and SBET for both adsorbents (refer to Table 3).
Effect of particle size on AZT adsorption
AZT adsorption on L-500 was examined using various sieves, resulting in particle sizes ranging from 75 to 300 μm. The outcomes are illustrated in Fig 6. It is evident that the smallest particle size of L-500 (< 75 μm) results in a higher rate of AZT adsorption, achieving maximum AZT removal (57%) within just 5 minutes of treatment. Smaller particles offer a larger specific surface area, offering more sites for AZT adsorption, thus explaining this phenomenon [70]. Using larger particle sizes of L-500, AZT removal rates approaching 55% were also attained after 40 min of adsorption. Hence, a particle size of < 300 μm has been selected for further experiments.
Experimental conditions: C0: 100 mg AZT L-1, adsorbent dose: 50 g L-1, pH: 7.0, T: 20–22°C.
Effect of pH on AZT adsorption
Fig 7 shows the effect of pH on AZT adsorption onto L-500 adsorbent. As illustrated in Fig 7A, right after adding the adsorbent (50 g L-1) to the AZT solutions, there is a noticeable shift in pH, signifying the strong buffering capacity of L-500. Consequently, the pH levels at which the AZT adsorption process occurs transition from 12.0, 7.0, and 3.0 to 9.0, 5.7, and 4.8, respectively. The results in Fig 7B indicate that AZT removal increases as pH levels rise for all the time intervals considered. At adsorption pH 9.0 near the AZT pKa (~8.7 [71]), the AZT adsorption is maximum, achieving 94% of the AZT removal within a contact time of 40 min. At this pH, AZT adsorption is associated with the electrostatic interaction between the negatively charged surface of the adsorbent L-500, which has a pHPZC of 6.4 (Fig 8A) and the protonated form of AZT (Fig 8B). Additionally, the hydrogen bonding mechanism between the protonated and unionized forms of AZT (Fig 7B) and Si-O- groups present on the L-500 surface can also contribute to AZT adsorption at pH 9.0 [28]. Conversely, AZT removal dropped to 64% and 55% at adsorption pH values of 5.7 and 4.8, respectively. This effect can be attributed to electrostatic repulsion between the predominant cationic moiety of AZT (Fig 8B) and the L-500, which presents a positively charged surface at pH 5.7 (Fig 8A). These electrostatic repulsion forces and their impact on AZT adsorption become more pronounced at lower pH levels, thus elucidating the reduced AZT removal achieved at pH 4.8. The above findings confirm the significant role of pH in the adsorption of AZT onto L-500. Considering that L-500 exhibited relatively high AZT adsorption at an initial pH of 7.0, which falls within the typical range (7.0–9.0) of effluents from municipal wastewater treatment plants [72], pH 7.0 was chosen for further sections.
Effect of pH on AZT adsorption using L-500: (a) Final pH, (b) AZT removal. Experimental conditions: C0: 100 mg AZT L-1, adsorbent dose: 50 g L-1, PS: <300 μm, T: 20–22°C.
AZT adsorption kinetics
The results of the AZT adsorption process onto L-500 were fitted with two widely employed models: the pseudo-first-order (PFO) and pseudo-second-order (PSO) models. S2 and S3 Figs display the fitting of the PFO model (Eq (3)) and PSO models (Eq (4)), respectively. The results of kinetic parameters compiled in Table 4 indicate that both kinetic models fit the experimental data, with R2 values exceeding 0.970. However, the PSO kinetic model exhibited superior R2 values (0.993–0.999) and low values for APE and Δq, indicating that the qecalcul with this model closely aligns with the qeexper. Overall, the interpretation of experimental data is enhanced when employing the PSO kinetic model. These findings suggest that chemisorption is the rate-limiting process in the adsorption of AZT onto L-500 [74], and they indicate a direct relationship between AZT adsorption capacity and the number of active sites present on L-500 (8). Chemical adsorption can encompass hydrogen bonding interactions [28]. Hydrogen bonds could potentially form between the amine group of the AZT and the oxygen atoms present in the functional groups of L-500 [8]. The strong correlation between the adsorption data and PSO kinetics is in line with prior findings on AZT adsorption in water when employing various adsorbents, such as biochar synthesized from rice husk, impregnated with montmorillonite and activated by CO2 gas [8], nanocomposites prepared by impregnating activated carbon with metals (Fe, Ag, Zn) [22], natural zeolite (clinoptilolite) modified with surfactants [26], mesoporous silica SBA-15 [24], organoclays [25], Azolla filiculoides-based activated carbon [23] and zeolites [28].
Adsorption isotherms
Adsorption isotherm models are utilized to investigate the adsorption process for removing pollutants from water, aiming to optimize design parameters [75]. The Langmuir, Freundlich, and Langmuir-Freundlich isothermal models (Table 2) were employed to describe the equilibrium data for AZT adsorption onto L-500. Fig 9 depicts the nonlinear fitting of three models to the experimental data. To accurately calculate the equilibrium constants of the adsorption isotherms, Fig 9 represents Ce and qe in units of mol L-1 and mol kg-1, respectively, following the suggestion made by Tran et al. [76]. Furthermore, Table 5 summarizes the fitting results (R2) and the parameters for each isotherm.
Experimental conditions: C0: 90 mg AZT L-1 (1.2 x 10−4 mol L-1), L-500 dose: 5–100 g L-1, contact time: 40 min, pH: 7.0, PS: <300 μm, T: 22°C.
Fig 9 demonstrates a strong fit of the experimental data to the three models, as indicated by high correlation coefficients (R²: 0.89–0.93), along with low APE values (2.22–2.78) and Δq values (0.06–4.03), as shown in Table 5. The good agreement between the experimental data and the Langmuir model suggests that, at equilibrium, AZT adsorption corresponds to a homogeneous monolayer on the surface of L-500, with no interactions between the adsorbed species [28]. This model enables the calculation of the maximum adsorption capacity (qm) for AZT (0.0016 mol kg-1 or 1.20 mg g−1), as well as the determination of the separation factor RL (0.0025), which falls within the range of 0 to 1, indicating that the adsorption of AZT onto L-500 is a favorable process [61]. However, we can still observe a strong fit of the experimental data to the Freundlich model (R2 = 0.93, APE = 2.30, Δq = 3.06), which considers interactions on multilayer heterogeneous surfaces. These surfaces exhibit varying affinities and interaction energies between their active sites and the adsorbate. This model is instrumental in predicting a favorable AZT adsorption process onto L-500, as indicated by 1/n < 1 [77]. Therefore, there are multiple interactions and adsorption energies between AZT and L-500, involving a combination of adsorption mechanisms. Furthermore, the best fitting to the Langmuir-Freundlich model (R2 = 0.93, APE = 2.22, Δq = 0.06) and the heterogeneity index value between 0 and 1 with tends toward zero (nLF = 0.084) indicate that the isotherm approaches a Freundlich-like form and L-500 is a heterogeneous material [78].
As indicated in Table 6, the maximum adsorption capacity (qm) achieved for AZT adsorption by L-500 in distilled water (1.20 mg g−1) surpasses that of previous reports using materials like lignite (0.395 mg g−1) and polyamide nanofibers (0.033 mg g−1) [21]. However, other materials, such as biochars [8,21], nanocomposites [22], mesoporous silica [24], nano diatomites [27], and zeolites [28], have reported qm values for AZT adsorption that are superior to those obtained with L-500. This indicates that the adsorbent produced in this study through the thermal activation of WTS (L-500) exhibits a moderate capability for AZT adsorption in aqueous solutions. As a result, its ability to treat wastewater contaminated with AZT will be assessed in the following section.
Interaction characterization and reusability
FTIR analysis before and after AZT adsorption on L-500 (Fig 10) was performed to elucidate the molecular interactions between AZT and the material. The spectra reveal significant changes in the material after AZT adsorption.
The shift in the band observed between 3020 and 3780 cm⁻¹ highlights the involvement of the OH functional group from mineral hydroxides in the adsorption process, emphasizing its critical role via hydrogen bonding interactions. Additionally, changes in the signals corresponding to C–H stretching at 2852 and 2921 cm⁻¹ suggest that AZT interacts with these functional groups in the carbonaceous matrix, likely through π-π interactions [79].
More pronounced changes are observed in the bands around 1030 cm⁻¹, 914 cm⁻¹, and 532 cm⁻¹, which are associated with Si–O and Al–OH groups. These findings confirm the vital role of Si–O and Al–O groups, whose negative charge density facilitates interactions with the positively charged nitrogen in AZT during adsorption.
The results indicate that multiple interactions—including π-π stacking, hydrogen bonding, and electrostatic interactions—contribute to the adsorption mechanism.
Reusability tests are crucial for evaluating the environmental, economic, and industrial feasibility of adsorbents [80,81]. The reusability of the L-500 adsorbent was evaluated over three consecutive cycles (Fig 11). This decrease in the adsorption capacity of L-500 is likely attributed to a loss of reactivity after several reuse cycles [82] and a reduction in available active binding sites on the adsorbent surface during regeneration cycles, particularly since chemisorption was the rate-limiting step in the adsorption process [83].
Experimental conditions: C0: 100 mg AZT L-1, L-500 dose: 50 g L-1, pH: 7.0, PS: <300 μm, T: 22°C.
AZT adsorption in complex matrices
The adsorption of AZT within complex matrices was evaluated by employing synthetic (S-WW) and real (R-WW) effluents from a municipal wastewater plant, spiked with AZT at a concentration of 100 mg L-1. Based on their chemical composition (S1 Table), the tested S-WW is characterized by a high concentration of inorganic salts and presence of organic matter. Similarly, the characterization results for the R-WW sample (S2 Table) reveal high concentrations of anions such as SO₄²⁻ and Cl⁻, along with organic matter (TOC), nitrogen compounds (Total Kjeldahl Nitrogen ‐ TKN, NO₂⁻, NO₃⁻), and phosphorus.
Consequently, cations (Ca2+, Mg2+, Na+, K+), anions (SO42-, Cl-, HCO3-, HPO42-), phosphorus and organic matter can compete with the AZT for the active sites on the adsorbent, affecting the adsorption process in both matrices as reported in previous studies Vrchovecká et al. [21] and De Sousa et al. [28]. As shown in Fig 12, the competitive effect in R-WW is evident, resulting in an 8% reduction in AZT removal after 60 min compared to distilled water (DW). However, AZT adsorption onto L-500 increases by 15% in S-WW compared to DW, achieving approximately 80% removal after 60 minutes of treatment. Additionally, there is a slight increase in the adsorption capacity at equilibrium (1.38 mg g-1) in S-WW, as indicated by the results of the PSO kinetic model (S4A Fig). This improvement in adsorption in S-WW matrix can be attributed to the higher pH levels attained in the S-WW during the adsorption process, in contrast to the pH of distilled water (S4B Fig). This elevated pH favors the adsorption of AZT onto L-500. Species such as HCO3- and HPO42- contribute to the buffering capacity in S-WW resulting in a smaller decrease in pH during the adsorption process with L-500.
Experimental conditions: C0: 100 mg AZT L-1, L-500 dose: 50 g L-1, pH: 7.0, PS: <300 μm, T: 22°C.
To emphasize the economic benefits of reusing WTS as an AZT adsorbent, Table 7 provides an approximate calculation of L-500 production costs, incorporating factors reported by Kumar et al. [59] and considering the specific conditions of the studied drinking water and wastewater treatment plants. While the values depend heavily on local conditions, L-500, derived from waste generated during the drinking water treatment process, offers a cost-effective solution. Adsorbents produced from waste typically have lower production costs compared to conventional options such as activated carbons, ion exchange resins, and zeolites [84].
Conclusions and perspectives
In this study, the adsorbent derived from drinking water sludge (WTS) through calcination at 500°C (L-500) demonstrated significant AZT removal efficiency in distilled water. A basic pH condition also enhanced AZT adsorption, primarily due to electrostatic attractions. The PSO adsorption kinetic model indicated that chemisorption is the rate-limiting step in AZT adsorption onto L-500, where the main interactions are π-π stacking, hydrogen bonding, and electrostatic attraction. It was also concluded that the Langmuir-Freundlich equilibrium isotherm model better represented AZT adsorption onto L-500, highlighting the varying interactions and adsorption energies involved in the process. The L-500 was reused for three cycles and showed a reduction of 19% in its AZT removal after the last cycle.
Future research on adsorbents derived from WTS to remove emerging pollutants should focus on continuous column tests, investigate synergistic treatment alternatives, and conduct pilot-scale studies to evaluate practical applications.
Ultimately, this work offers a promising avenue to reuse WTS, addressing two concurrent challenges simultaneously: mitigating water contamination by antibiotics and valorizing waste generated during drinking water treatment. By highlighting the potential of WTS-derived adsorbents, this study contributes to the ongoing efforts to develop sustainable solutions for water treatment and resource recovery.
Supporting information
S1 Table. Composition of synthetic municipal wastewater.
https://doi.org/10.1371/journal.pone.0316487.s004
(DOCX)
S2 Table. Characterization of real municipal wastewater (R-WW).
https://doi.org/10.1371/journal.pone.0316487.s005
(DOCX)
S1 Fig. N2 adsorption-desorption isotherm of L-100 and L-500.
Inset: Pore size distributions.
https://doi.org/10.1371/journal.pone.0316487.s006
(DOCX)
S2 Fig.
Pseudo-first-order kinetic for AZT adsorption onto L-500: (a) 50 mg L-1, (b) 60 mg L-1, (c) 70 mg L-1, (d) 80 mg L-1, (e) 90 mg L-1, (f) 100 mg L-1. Experimental conditions: C0: 50–100 mg AZT L-1, L-500 dose: 50 g L-1, pH: 7.0, PS: <300 μm, T: 22°C.
https://doi.org/10.1371/journal.pone.0316487.s007
(DOCX)
S3 Fig.
Pseudo-second-order kinetic for AZT adsorption onto L-500: (a) 50 mg L-1, (b) 60 mg L-1, (c) 70 mg L-1, (d) 80 mg L-1, (e) 90 mg L-1, (f) 100 mg L-1. Experimental conditions: C0: 50–100 mg AZT L-1, L-500 dose: 50 g L-1, pH: 7.0, PS: <300 μm, T: 22°C.
https://doi.org/10.1371/journal.pone.0316487.s008
(DOCX)
S4 Fig.
AZT adsorption on L-500 in municipal wastewater effluent. (a) Adsorption capacity, (b) Final pH. Experimental conditions: C0: 100 mg AZT L−1, L-500 dose: 50 g L−1, pH: 7.0, P.S.: <300 μm, T: 22°C.
https://doi.org/10.1371/journal.pone.0316487.s009
(DOCX)
References
- 1.
UN General Assembly. Transforming our world: the 2030 Agenda for Sustainable Development A/RES/70/1. 2015 Oct 21 [cited 2024 March 2]. Available from: https://www.refworld.org/docid/57b6e3e44.html [Accessed 26 October 2023].
- 2. He L, Chen Y, Li Y, Sun F, Zhao Y, Yang S. Adsorption of Congo red and tetracycline onto water treatment sludge biochar: characterisation, kinetic, equilibrium and thermodynamic study. Water Sci Technol. 2022 March 15; 85 (6): 1936–1951. pmid:35358080
- 3. Ashraf A, Liu G, Yousaf B, Arif M, Ahmed R., Rashid A, et al. Phyto-mediated photocatalysis: a critical review of in-depth base to reactive radical generation for erythromycin degradation. Environmental Science and Pollution Research. 2022; 29 (22): 32513–32544. pmid:35190984
- 4. Sosa-Hernández JE., Rodas-Zuluaga LI, López-Pacheco IY., Melchor-Martínez EM, Aghalari Z, Salas D, et al. Sources of antibiotics pollutants in the aquatic environment under SARS-CoV-2 pandemic situation. Case studies in chemical and environmental engineering. 2021; 4: 100127. pmid:38620862
- 5. Manaia CM, Macedo G, Fatta-Kassinos D, Nunes OC. Antibiotic resistance in urban aquatic environments: can it be controlled? Applied microbiology and biotechnology. 2016; 100: 1543–1557. pmid:26649735
- 6.
Loos R., Marinov D., Sanseverino I., Napierska D, Lettieri T. Review of the 1st Watch List under the Water Framework Directive and recommendations for the 2nd Watch List. Publications Office of the European Union: Luxembourg. 2018; https://doi.org/10.2760/614367
- 7. Samrot A. V., Wilson S., Sanjay Preeth R. S., Prakash P., Sathiyasree M., Saigeetha S., et al. Sources of antibiotic contamination in wastewater and approaches to their removal—An overview. Sustainability. 2023, 15 (16), 12639.
- 8. Arif M, Liu G, Zia ur Rehman M, Mian MM., Ashraf A., Yousaf B, et al. Impregnation of biochar with montmorillonite and its activation for the removal of azithromycin from aqueous media. Environmental Science and Pollution Research. 2023; 30 (32): 78279–78293. pmid:37269518
- 9. Echeverría-Esnal D, Martin-Ontiyuelo C, Navarrete-Rouco ME, De-Antonio M, Ferrández O, Horcajada JP, et al. Azithromycin in the treatment of COVID-19: a review. Expert review of anti-infective therapy. 2021; 19 (2): 147–163. pmid:32853038
- 10. Zhao L., Lv Z., Lin L., Li X., Xu J., Huang S., et al. Impact of COVID-19 pandemic on profiles of antibiotic-resistant genes and bacteria in hospital wastewater. Environmental Pollution. 2023; 334, 122133. pmid:37399936
- 11. Azari A., Malakoutian M., Yaghmaeain K., Jaafarzadeh N., Shariatifar N., Mohammadi G., et al. Magnetic NH2-MIL-101 (Al)/Chitosan nanocomposite as a novel adsorbent for the removal of azithromycin: modeling and process optimization. Scientific Reports. 2022; 12(1), 18990. pmid:36347864
- 12. Singh V., Gupta S. P., Samanta S. K. Water resource rejuvenation via AOP based degradation of pharmaceuticals extensively used during COVID-19. Journal of Water Process Engineering. 2024; 67, 106137.
- 13. Raut S., Behera A. K., Sahoo S. K. Electrospun polyacrylonitrile reinforced greenly synthesized iron oxide nanocomposite fibers sheet for remediation of azithromycin from water. Materials Today Communications. 2024; 40, 110113.
- 14. Sabater-Liesa L., Montemurro N., Ginebreda A., Barceló D., Eichhorn P., Pérez S. Retrospective mass spectrometric analysis of wastewater-fed mesocosms to assess the degradation of drugs and their human metabolites. Journal of Hazardous Materials. 2021; 408, 124984. pmid:33418519
- 15. Milaković M., Vestergaard G., González-Plaza J. J., Petrić I., Šimatović A., Senta I., et al. Pollution from azithromycin-manufacturing promotes macrolide-resistance gene propagation and induces spatial and seasonal bacterial community shifts in receiving river sediments. Environment international. 2019; 123, 501–511. pmid:30622075
- 16. Hussain A., Afzal O., Altamimi A. S., Ali R. Application of green nanoemulsion to treat contaminated water (bulk aqueous solution) with azithromycin. Environmental Science and Pollution Research. 2021; 28, 61696–61706. pmid:34184229
- 17. Koch D. E., Bhandari A., Close L., Hunter R. P. Azithromycin extraction from municipal wastewater and quantitation using liquid chromatography/mass spectrometry. Journal of Chromatography A. 2005; 1074(1–2), 17–22. pmid:15941034
- 18. Morales-Paredes CA, Rodríguez-Díaz JM., Boluda-Botella N. Pharmaceutical compounds used in the COVID-19 pandemic: A review of their presence in water and treatment techniques for their elimination. Science of the Total Environment. 2022; 814: 152691. pmid:34974020
- 19. Botero-Coy AM, Martínez-Pachón D, Boix C, Rincón RJ, Castillo N., Arias-Marín LP, et al. An investigation into the occurrence and removal of pharmaceuticals in Colombian wastewater. Science of the Total Environment. 2018; 642: 842–853. pmid:30045524
- 20. Cano PA, Jaramillo-Baquero M, Zúñiga-Benítez H, Londoño YA, Peñuela GA. Use of simulated sunlight radiation and hydrogen peroxide in azithromycin removal from aqueous solutions: optimization & mineralization analysis. Emerging Contaminants. 2020; 6: 53–61.
- 21. Vrchovecká S, Asatiani N, Antoš V, Wacławek S, Hrabák P. Study of Adsorption Efficiency of Lignite, Biochar, and Polymeric Nanofibers for Veterinary Drugs in WWTP Effluent Water. Water, Air, & Soil Pollution. 2023; 234 (4): 268.
- 22. Mehrdoost A, Yengejeh RJ, Mohammadi MK, Haghighatzadeh A, Babaei AA. Adsorption removal and photocatalytic degradation of azithromycin from aqueous solution using PAC/Fe/Ag/Zn nanocomposite. Environmental Science and Pollution Research. 2022; 29 (22): 33514–33527. pmid:35029828
- 23. Balarak D, Mahvi AH, Shahbaksh S, Wahab MA, Abdala A. Adsorptive removal of azithromycin antibiotic from aqueous solution by azolla filiculoides-based activated porous carbon. Nanomaterials. 2021; 11(12): 3281. pmid:34947630
- 24. Gholamian S, Hamzehloo M, Farrokhnia A, Mahdavifar Z. Response surface methodology optimizing the adsorptive removal of azithromycin using mesoporous silica SBA-15: mechanism, thermodynamic, equilibrium, and kinetics modeling studies. Journal of Environmental Science and Health, Part A. 2021; 56 (10): 1145–1164. pmid:34558387
- 25. Imanipoor J, Mohammadi M, Dinari M. Evaluating the performance of L-methionine modified montmorillonite K10 and 3-aminopropyltriethoxysilane functionalized magnesium phyllosilicate organoclays for adsorptive removal of azithromycin from water. Separation and Purification Technology. 2021; 275: 119256.
- 26. Saadi Z, Fazaeli R, Vafajoo L, Naser I, Mohammadi G. Promotion of clinoptilolite adsorption for azithromycin antibiotic by Tween 80 and Triton X-100 surface modifiers under batch and fixed-bed processes. Chemical Engineering Communications. 2021; 208 (3): 328–348.
- 27. Davoodi S, Dahrazma B, Goudarzi N, Gorji HG. Adsorptive removal of azithromycin from aqueous solutions using raw and saponin-modified nano diatomite. Water Science and Technology. 2019; 80 (5): 939–949. pmid:31746801
- 28. De Sousa DNR, Insa S, Mozeto AA, Petrovic M, Chaves TF, Fadini PS. Equilibrium and kinetic studies of the adsorption of antibiotics from aqueous solutions onto powdered zeolites. Chemosphere. 2018; 205: 137–146. pmid:29689527
- 29. Racar M, Dolar D, Karadakić K, Čavarović N, Glumac N, Ašperger D, et al. Challenges of municipal wastewater reclamation for irrigation by MBR and NF/RO: physico-chemical and microbiological parameters, and emerging contaminants. Science of the Total Environment. 2020; 722: 137959. pmid:32208282
- 30. Rueda-Márquez JJ, Palacios-Villarreal C, Manzano M., Blanco E., del Solar M R, Levchuk I. Photocatalytic degradation of pharmaceutically active compounds (PhACs) in urban wastewater treatment plants effluents under controlled and natural solar irradiation using immobilized TiO2. Solar Energy. 2020; 208: 480–492.
- 31. Fiorentino A, Esteban B, Garrido-Cardenas JA, Kowalska K, Rizzo L, Aguera A, et al. Effect of solar photo-Fenton process in raceway pond reactors at neutral pH on antibiotic resistance determinants in secondary treated urban wastewater. Journal of Hazardous Materials. 2019; 378: 120737. pmid:31202058
- 32. Sayadi MH, Sobhani S, Shekari H. Photocatalytic degradation of azithromycin using GO@ Fe3O4/ZnO/SnO2 nanocomposites. Journal of Cleaner Production. 2019; 232: 127–136.
- 33. Serna-Galvis EA, Botero-Coy AM, Martínez-Pachón D, Moncayo-Lasso A, Ibáñez M, Hernández F, et al. Degradation of seventeen contaminants of emerging concern in municipal wastewater effluents by sonochemical advanced oxidation processes. Water Research. 2019; 154: 349–360. pmid:30818100
- 34. Bayati M, Ho TL, Vu DC, Wang F, Rogers E, Cuvellier C, et al. Assessing the efficiency of constructed wetlands in removing PPCPs from treated wastewater and mitigating the ecotoxicological impacts. International Journal of Hygiene and Environmental Health. 2021; 231: 113664. pmid:33212356
- 35. Tang K, Rosborg P, Rasmussen ES, Hambly A, Madsen M, Jensen NM, et al. Impact of intermittent feeding on polishing of micropollutants by moving bed biofilm reactors (MBBR). Journal of Hazardous Materials. 2021; 403: 123536. pmid:32823027
- 36. Kiki C, Rashid A, Wang Y, Li Y, Zeng Q, Yu CP, et al. Dissipation of antibiotics by microalgae: Kinetics, identification of transformation products and pathways. Journal of Hazardous Materials. 2020; 387: 121985. pmid:31911384
- 37. Liu PY, Chen JR, Shao L, Tan J, Chen DJ. Responses of flocculent and granular sludge in anaerobic sequencing batch reactors (ASBRs) to azithromycin wastewater and its impact on microbial communities. Journal of Chemical Technology & Biotechnology. 2018; 93 (8): 2341–2350.
- 38. Sbardella L, Comas J, Fenu A, Rodriguez-Roda I, Weemaes M. Advanced biological activated carbon filter for removing pharmaceutically active compounds from treated wastewater. Science of the Total Environment. 2018; 636: 519–529. pmid:29715656
- 39. Mojahedimotlagh F., Nasab E. A., Foroutan R., Vakilabadi D. R., Dobaradaran S., Azamateslamtalab E., et al. Azithromycin decomposition from simple and complex waters by H2O2 activation over a recyclable catalyst of clay modified with nanofiltration process brine. Environmental Technology & Innovation. 2024; 33, 103512.
- 40. Ramos B.D.P, Perez I.D, Aliprandini P, Boina R.F. Cu2+, Cr3+, and Ni 2+ in mono-and multi-component aqueous solution adsorbed in passion fruit peels in natura and physicochemically modified: A comparative approach. Environmental Science and Pollution Research. 2022; 29: 79841–79854. pmid:34981402
- 41. Sousa ÉM, Otero M, Rocha LS, Gil MV, Ferreira P, Esteves VI, et al. Multivariable optimization of activated carbon production from microwave pyrolysis of brewery wastes-Application in the removal of antibiotics from water. Journal of Hazardous Materials. 2022; 431: 128556. pmid:35255334
- 42. Yang J, Ren Y, Chen S, Zhang Z, Pang H, Wang X, et al. Thermally activated drinking water treatment sludge as a supplementary cementitious material: Properties, pozzolanic activity and hydration characteristics. Construction and Building Materials. 2023; 365: 130027.
- 43. Bensitel N, Haboubi K, Azar FZ, El Hammoudani Y, El Abdouni A, Haboubi C, et al. Potential reuse of sludge from a potable water treatment plant in Al Hoceima city in northern Morocco. Water Cycle. 2023; 4: 154–162.
- 44. Sharma A, Ahammed MM. Application of modified water treatment residuals in water and wastewater treatment: A review. Heliyon. 2023; 9 (5): e15796. pmid:37305496
- 45. Punamiya P, Sarkar D, Rakshit S, Datta R. Effectiveness of aluminum‐based drinking water treatment residuals as a novel sorbent to remove tetracyclines from aqueous medium. Journal of Environmental Quality. 2013; 42 (5): 1449–1459. pmid:24216422
- 46. Saman N., Subramanian K. K., Johari K., Mat Taib S., & Marčiulaitienė E. Physicochemistry properties of water treatment sludge (WTS) as adsorbents for dyes and antibiotics removal. Conference paper. 12th International Conference “Environmental Engineering”. 2023, Vilnius, Lithuania.
- 47. Brunauer S, Emmett PH, Teller E. Adsorption of gases in multimolecular layers. Journal of the American Chemical Society. 1938; 60 (2): 309–319.
- 48. Guechi EK, Hamdaoui O. Biosorption of methylene blue from aqueous solution by potato (Solanum tuberosum) peel: equilibrium modelling, kinetic, and thermodynamic studies. Desalination and Water Treatment. 2016; 57 (22): 10270–10285.
- 49. Martínez-Polanco MP, Valderrama-Rincón JA, Martínez-Rojas AJ, Luna-Wandurraga HJ, Díaz-Báez MC, Bustos-López MC, et al. Degradation of high concentrations of azithromycin when present in a high organic content wastewater by using a continuously fed laboratory-scale UASB bioreactor. Chemosphere. 2022; 287: 132191. pmid:34509021
- 50. Kumar V, Singh SK, Gulati M, Anishetty R, Shunmugaperumal T. Development and validation of a simple and sensitive spectrometric method for estimation of azithromycin dihydrate in tablet dosage forms: application to dissolution studies. Current Pharmaceutical Analysis. 2013; 9 (3): 310–317.
- 51. Sultana N, Arayne MS, Hussain F, Fatima A. Degradation studies of azithromycin and its spectrophotometric determination in pharmaceutical dosage forms. Pak J Pharm Sci. 2006 Apr; 19 (2): 98–103. pmid:16751118.
- 52. Revellame ED, Fortela DL, Sharp W, Hernandez R, Zappi ME. Adsorption kinetic modeling using pseudo-first order and pseudo-second order rate laws: A review. Cleaner Engineering and Technology. 2020; 1: 100032.
- 53. Wang J, Guo X. Adsorption kinetic models: Physical meanings, applications, and solving methods. Journal of Hazardous Materials. 2020; 390: 122156. pmid:32006847
- 54. Freundlich HMF. Over the adsorption in solution. J. Phys. Chem. 1906; 57 (385471): 1100–1107.
- 55. Langmuir I. The adsorption of gases on plane surfaces of glass, mica and platinum. J. Am. Chem. Soc. 1918; 40 (9): 1361–1403.
- 56. Sips R. On the structure of a catalyst surface. J. Chem. Phys. 1948; 16 (5): 490–495.
- 57. Sips R. On the structure of a catalyst surface. II. J. Chem. Phys. 1950; 18 (8): 1020–1026.
- 58. Jeppu GP, Clement TP. A modified Langmuir-Freundlich isotherm model for simulating pH-dependent adsorption effects. Journal of Contaminant Hydrology. 2012; 129–130: 46–53. pmid:22261349
- 59. Kumar N. S., Asif M., Poulose A. M., Al-Ghurabi E. H., Alhamedi S. S., & Koduru J. R. Date palm fiber agro-waste biomass for efficient removal of 2, 4, 6-Trichlorophenol from aqueous solution: Characterization, Kinetics, Isotherms studies and Cost-effective analysis. Desalination and Water Treatment. 2024, 100405.
- 60. Paredes-Laverde M, Salamanca M, Silva-Agredo J, Manrique-Losada L, Torres-Palma RA. Selective removal of acetaminophen in urine with activated carbons from rice (Oryza sativa) and coffee (Coffea arabica) husk: Effect of activating agent, activation temperature and analysis of physical-chemical interactions. Journal of Environmental Chemical Engineering. 2019; 7 (5): 103318.
- 61. Nagaraj A, Sadasivuni KK, Rajan M. Investigation of lanthanum impregnated cellulose, derived from biomass, as an adsorbent for the removal of fluoride from drinking water. Carbohydrate Polymers. 2017; 176: 402–410. pmid:28927624
- 62. Ospina-Montoya V., Cardozo V., Porras J., Acelas N., Forgionny A. Valorization of coffee husks for the sustainable removal of pharmaceuticals from aqueous solutions. H2Open Journal, 2024, 7(3), 303–317.
- 63. Paredes-Laverde M, Silva-Agredo J, Torres-Palma RA. Removal of norfloxacin in deionized, municipal water and urine using rice (Oryza sativa) and coffee (Coffea arabica) husk wastes as natural adsorbents. Journal of Environmental Management. 2018; 213: 98–108. pmid:29482094
- 64.
Organization for Economic Co-operation and Development -OECD-. Test No. 303: Simulation Test ‐ Aerobic Sewage Treatment ‐ A: Activated Sludge Units; B: Biofilms. In: OECD Guidelines for the Testing of Chemicals, Section 3. Paris: OECD Publishing; 2001. doi: https://doi.org/10.1787/9789264070424-en.
- 65. Teixeira SR, Santos GTA, Souza AE, Alessio P, Souza SA., Souza NR. The effect of incorporation of a Brazilian water treatment plant sludge on the properties of ceramic materials. Applied Clay Science. 2011; 53 (4): 561–565.
- 66. Martins DS, Estevam BR, Perez ID, Américo-Pinheiro JHP, Isique WD, Boina RF. Sludge from a water treatment plant as an adsorbent of endocrine disruptors. Journal of Environmental Chemical Engineering. 2022; 10 (4): 108090.
- 67. Everaert M, Bergmans J, Broos K, Hermans B, Michielsen B. Granulation and calcination of alum sludge for the development of a phosphorus adsorbent: from lab scale to pilot scale. Journal of Environmental Management. 2021; 279: 111525. pmid:33168303
- 68. Shamaki M, Adu-Amankwah S, Black L. Reuse of UK alum water treatment sludge in cement-based materials. Construction and Building Materials. 2021; 275: 122047.
- 69. Jeon EK, Ryu S, Park SW, Wang L, Tsang DC., Baek K. Enhanced adsorption of arsenic onto alum sludge modified by calcination. Journal of Cleaner Production. 2018; 176: 54–62.
- 70. Paredes-Laverde M., Salamanca M., Diaz-Corrales JD, Flórez E., Silva-Agredo J, Torres-Palma RA. Understanding the removal of an anionic dye in textile wastewaters by adsorption on ZnCl2 activated carbons from rice and coffee husk wastes: A combined experimental and theoretical study. Journal of Environmental Chemical Engineering. 2021; 9 (4): 105685.
- 71. Sidhu H, D’Angelo E, O’Connor G. Retention-release of ciprofloxacin and azithromycin in biosolids and biosolids-amended soils. Science of the Total Environment. 2019; 650 (Part 1): 173–183. pmid:30196217
- 72. Nie J, Yan S, Lian L, Sharma VK, Song W. Development of fluorescence surrogates to predict the ferrate (VI) oxidation of pharmaceuticals in wastewater effluents. Water Research. 2020; 185: 116256. pmid:32768661
- 73. Rodríguez-López L, Santás-Miguel V, Núñez-Delgado A, Álvarez-Rodríguez E, Pérez-Rodríguez P, Arias-Estévez M. Influence of pH, humic acids, and salts on the dissipation of amoxicillin and azithromycin under simulated sunlight. Spanish Journal of Soil Science. 2022; 12: 10438.
- 74. Ramirez A, Ocampo R, Giraldo S, Padilla E, Flórez E, Acelas N. Removal of Cr (VI) from an aqueous solution using an activated carbon obtained from teakwood sawdust: Kinetics, equilibrium, and density functional theory calculations. Journal of Environmental Chemical Engineering. 2020; 8 (2): 103702.
- 75. Subramanyam B, Das A. Linearised and non-linearised isotherm models optimization analysis by error functions and statistical means. J Environ Health Sci Engineer. 2014; 12: 92. pmid:25018878.
- 76. Tran HN, Lima EC, Juang RS, Bollinger JC, Chao HP. Thermodynamic parameters of liquid–phase adsorption process calculated from different equilibrium constants related to adsorption isotherms: A comparison study. Journal of Environmental Chemical Engineering. 2021; 9 (6): 106674.
- 77. Tran HN, You SJ, Hosseini-Bandegharaei A, Chao HP. Mistakes and inconsistencies regarding adsorption of contaminants from aqueous solutions: a critical review. Water Research. 2017; 120: 88–116. pmid:28478298
- 78. Turiel E, Perez-Conde C, Martin-Esteban A. Assessment of the cross-reactivity and binding sites characterisation of a propazine-imprinted polymer using the Langmuir-Freundlich isotherm. Analyst. 2003; 128 (2): 137–141. pmid:12625553
- 79. Bougrine O., El Fellah I., Kada I., Rabie F. A., Lanjri A. F., & Ammari M. Advancing Circular Economy: A study of Drinking Water Sludge for Potential Uses. Results in Engineering. 2024; 102426.
- 80. Upoma B. P., Yasmin S., Ali Shaikh M. A., Jahan T., Haque M. A., Moniruzzaman M., et al. A fast adsorption of azithromycin on waste-product-derived graphene oxide induced by H-bonding and electrostatic interactions. ACS Omega. 2022; 7(34), 29655–29665. pmid:36061663
- 81. Igwegbe C. A., Oba S. N., Aniagor C. O., Adeniyi A. G., & Ighalo J. O. Adsorption of ciprofloxacin from water: a comprehensive review. Journal of Industrial and Engineering Chemistry. 2021; 93, 57–77.
- 82. Angaru G. K. R., Lingamdinne L. P., Choi Y. L., Koduru J. R., & Chang Y. Y. Catalytic binary oxides decorated zeolite as a remedy for As (III) polluted groundwater: Synergistic effects and mechanistic analysis. Journal of Environmental Chemical Engineering. 2023; 11(2): 109544.
- 83. Al-Hakkani M. F., Gouda G. A., Hassan S. H., Mohamed M. M., & Nagiub A. M. (2022). Environmentally azithromycin pharmaceutical wastewater management and synergetic biocompatible approaches of loaded azithromycin@ hematite nanoparticles. Scientific Reports. 2022; 12(1), 10970. pmid:35768496
- 84. Mangla D., Sharma A., & Ikram S. Critical review on adsorptive removal of antibiotics: Present situation, challenges and future perspective. Journal of Hazardous Materials. 2022; 425, 127946. pmid:34891019