Purifying water for diverse uses is vital, but concerns lie with the sustainability and accessibility of purification materials. As such, this study converted readily available water treatment plant sludge (WTPS) into activated adsorbent for phosphate removal in wastewater. WTPS was activated via thermal activation at 300 °C temperature and chemical activation processes of 3 M acid concentration, 4 h activation time, and 75 °C activation temperature, and then characterized using Brunauer-Emmett-Teller (BET), scanning electron microscopy (SEM), Fourier transform infrared, elemental analyzer, and differential scanning calorimetry. SEM and BET analyses revealed a highly porous adsorbent (279.2 m2/g) for efficient adsorption. On top of the activation process, preliminary experiments and numerical optimization using response surface methodology (RSM) were designed and conducted. Through optimizing conditions, it was found that 70 min of contact time, pH 3, 3 g/L adsorbent dose, and 30 mg/L initial phosphate concentration as optimal, yielding 83% removal efficiency. Furthermore, adsorption kinetics and isotherm models were examined and the second-order kinetics and Langmuir isotherm models indicated best fit. Notably, the activated sludge could be regenerated for three cycles before efficiency dropped below 70%. Thus, activated WTPS presents a promising, sustainable, and readily available adsorbent for phosphate removal in wastewater treatment.

  • Water treatment plant sludge (WTPS) was activated through thermal activation and chemical activation.

  • The activated WTPS adsorbent has a surface area of 279.2 m2/g.

  • The activated WTPS showed an 83% removal efficiency.

  • The regeneration test indicated that the activated sludge can be used three times, repeatedly, before its efficacy falls below 70%.

Water consumption is rapidly increasing globally, as has been observed in the past two centuries. However, water quality deterioration continues to increase. Water resources are getting increasingly polluted due to anthropogenic and natural reasons (Lu et al. 2015). Phosphorus is a major source of water contamination. Eutrophication is also caused by the excessive release of phosphorus into water and results in water quality degradation. Wastewater runoff containing excess phosphorus species is a major environmental issue worldwide, as the nutrient leads surface water bodies to eutrophication. Various human activities, such as mining and industrial and agricultural use, release phosphorus into the aquatic system (Ali et al. 2019).

The increase of phosphate in water bodies promotes the growth of algae, which results in eutrophication and eventually consumes dissolved oxygen thereby adversely affecting the water quality (Xu et al. 2016; Berkessa et al. 2019). Municipal and industrial wastewaters are the main point sources for phosphate while runoff from agriculture is the dominant non-point source. Studies indicate that municipal wastewater may contain 4–15 mg/L phosphate, whereas effluent from chemical industries such as detergent manufacturing and metal coating processes may contain 14–25 mg/L phosphate (Afzaal et al. 2022; Kiprono et al. 2023). According to Kang & Cao (2014), the tolerable phosphate level in water should not exceed 0.05 mg/L to maintain an ecologically sustainable status. In such a case, lowering the phosphate loading in wastewater and runoff especially if local circumstances do not allow for advanced techniques such as membrane filtration became a challenge to local scientists and engineers (Rehman et al. 2018). The most widespread wastewater treatment technology for phosphate removal is based on precipitation processes, in which massive amounts of chemicals such as iron and aluminum salts are utilized (Adhikari et al. 2016). However, many techniques are suffering from either large amounts of sludge for disposal or high operational and maintenance costs with recurring expenses, which are not suitable for many developing countries such as Ethiopia. Therefore, searching for cost-effective and environmentally sound phosphate removal alternatives for low-income countries is essential.

A variety of physical, chemical, and biological methods have been developed in recent years for the phosphate removal from wastewater (Yan et al. 2014; Acelas et al. 2015; Mor et al. 2016; Usman et al. 2022). Advanced biological methods can remove up to 97% of phosphate and generate low amounts of sludge, but the method has limited applicability (Nobaharan et al. 2021). Similarly, physical processing (sedimentation, membrane filtration, etc.) techniques are too expensive, are accompanied by a high sludge production, and are often inefficient in the phosphate removal from wastewater effluent (Al Tahmazi 2017; Banks et al. 2020).

Adsorption emerges as a robust process that could solve the aforementioned problems and render the treatment system more economically viable, especially if low-cost adsorbents are involved (Renu et al. 2017; Nadew et al. 2023). Water treatment processes produce millions of tons of solid waste annually worldwide, and the amount of sludge produced rises in tandem with rising water consumption. With an estimated 10,000 tons of dry water treatment residual produced daily worldwide, managing water treatment plant sludge (WTPS) continues to be a difficult environmental and financial issue for all water authorities (Nguyen et al. 2022). Landfills, once the default dumping ground for water treatment sludge, are filling up and revealing their environmental downsides. The cost of this disposal method is skyrocketing, prompting a shift toward circular economy solutions. Recovering, reusing, and recycling sludge are no longer fringe ideas but essential steps for water utilities seeking sustainable resource management. Circular economy eliminates waste, not by burying it but by finding new life for materials – a paradigm shift offering a welcome escape from the linear pitfalls of the past (Jung et al. 2016).

The effectiveness of water treatment sludge as a coagulant and an adsorbent of pollutants in water treatment has been demonstrated by many studies (Drechsel et al. 2015; Ooi et al. 2018; Bensitel et al. 2023; Kumari et al. 2023). However, the use of activated water treatment sludge for water purification or removal of impurities from wastewater is not yet more common and not pronounced. Therefore, this study aimed to synthesize activated adsorbent material from WTPS for the phosphate removal impurities in water. The study mainly focused on the investigation of phosphate removal efficiency from synthetic and wastewater treatment effluents by the adsorption process.

Materials

The WTPS was collected from the Legedadi water treatment plant in Addis Ababa, Ethiopia. The treatment plant is situated at a height of 2,450 m above sea level and in the geographic coordinates of 38° 60′ to 39° 07′ E longitude and 9° 01′ to 9° 13′ N latitude. The plant is 5,324 m2 in surface area, with a maximum water depth of 30m and a mean depth of 4 m. The sample was taken from the four discharging points of the treatment plant and mixed well to get a representative sample. Since the quantity and quality of the sludge vary depending on the seasons, samples were collected in the summer and winter seasons. All the reagents and chemicals used for the current investigation were analytical grades used without further purification.

Methods

Material preparation

The sludge samples collected from four disposal points in the two seasons were mixed to attain representative samples for further analyses, and all the samples were stored at a low temperature (4 °C). The collected samples were decanted to eliminate the liquid part and retain the sludge. Some unnecessary materials available in the sludge were removed manually, and the remaining sludge was dried in an oven at 105 °C for 24 h. The sample was cooled off to room temperature in a desiccator and then ground using a disk miller to a size of below 130 μm. The prepared sample was stored in a plastic bag to prevent any contamination (Nguyen et al. 2023).

Synthesis and activation process of WTPS adsorbent

To synthesize an activated adsorbent, the WTPS was subjected to thermal and acid activation processes in a row. The thermal activation processes were performed by putting 30 g of sample into a furnace and thermally treated with 300 °C temperature at a rate of 15 °C/min for 4 h under a continuous supply of nitrogen gas. The nitrogen environment prevents the sample from burning and allows only volatile components to escape (Volperts et al. 2021). When the volatile components are left off at elevated temperatures, it is assumed that a void volume would be created, which promotes considerable surface area of the sample. The sample was then allowed to cool to room temperature and ground to a size of below 100 μm using a hammer miller.

On top of that, the thermally semi-activated sludge was further activated chemically using phosphoric acid to enhance its surface area and porosity and modify its affinity to attract adsorbate yet the more. To that end, three parameters, namely acid concentration, activation time, and activation temperature were used to activate the intermediate adsorbent. To examine the effect that these parameters have on the activation process of WTPS, the level of each parameter was set to a wide range considering that the variables have substantial influence on the adsorbing capacity of adsorbate (Table 1) (Zakaria et al. 2021). Hence, 30 g of partially activated sludge was subjected to different acid concentrations of 1, 2, 3, 4, and 5sM heated at a constant temperature of 75 °C for 4 h to scrutinize the effect of acid concentration variations. The activation process was allowed to take place in a 1 L cylindrical beaker, and the solution was agitated at 300 rpm using a mechanical stirrer to keep uniform temperature distribution and to get a gentle acid–sludge contact. Similarly, the sample activation time was varied from 2, 3, 4, 5, and 6 h while the acid concentration and activation temperature were fixed at 3 M and 75 °C, respectively. Moreover, the range of activation temperature was from 55, 65, 75, 85, and 95 °C to observe its effect on the surface of the adsorbent at a fixed acid concentration of 3 M and activation time of 4 h. In this fashion, 15 samples were activated, settled, neutralized with distilled water, and filtered to retain the solid part. The solid part was then dried at 105 °C for 24 h to remove residual water molecules within.

Table 1

Level of parameters for acid activation of WTPS

Varying parameterRange of parameterFixed parameter
Acid concentration (M) At 75 °C and 4 h 
Activation time (h) At 75 °C and 3 M 
Activation temperature (°C) 55 65 75 85 95 At 3 M and 4 h 
Varying parameterRange of parameterFixed parameter
Acid concentration (M) At 75 °C and 4 h 
Activation time (h) At 75 °C and 3 M 
Activation temperature (°C) 55 65 75 85 95 At 3 M and 4 h 

Physicochemical characterization of activated sludge adsorbent

Proximate determination: The proximate values such as moisture contents, volatile matters, fixed carbons, and ash contents of the raw sludge and prepared activated sludge adsorbent were studied using standard procedures of the American Society for Testing and Materials (ASTM) (Kassahun et al. 2022).

Elemental analysis: The implications of a possible application in the adsorption of phosphate impurities, as well as the suitability and impact on the environment, depend heavily on the elemental analysis. Using CHNS/O analyzers (EMA 502, VELP, China), the elemental analysis of raw and activated WTPS was examined. The elemental composition of raw and activated WTPS (C, H, N, and S) was determined using an ASTM-D5373 method. Five-milligram samples of completely dried WTPS (moisture content < 1%) were weighed into clean tin cups and analyzed using an elemental analyzer calibrated at 1,060 °C combustion furnace temperature. The EMA SoftTM software automatically calculates elemental weight percentages from real-time atomic ratios for each element. To determine the percentage of oxygen (O), the total of C, H, N, and S was deducted from 100% (Rai et al. 2016).

Point zero charge determination: The surface charge of the adsorbent depends on the pH of the solution and, furthermore, depends on its pH of point zero charges, pHpzc, at which the net charge on the adsorbent surface is zero (Puri & Kumar 2019). The adsorbent surface is negatively charged when pH > pHpzc and positively charged if pH < pHpzc (Belachew & Hinsene 2020). To determine the point zero charge of the samples, 3 g of sample was mixed with 50 mL of 0.1 N KNO3 solution of pH ranging from 2 to 14 at intervals of 2 units. The pH of the solution was made using NaOH and HCl solution in a 1 L flask and gently shaken for 48 h at ambient conditions. The pH of each solution was measured and the net charge (ΔpH) was calculated from the initial and final pH values. The pH values are plotted along the x-axis and ΔpH along the y-axis; the data obtained from the experiment are plotted, and the intersection point is taken as a reference for determining the pHPZC.

BET surface area study: The surface areas of both raw and activated samples were determined using the surface area analyzer (Horiba 96000 series). The surface area was determined as 0.6 g of samples were weighed and put into the sample preparation unit of the analyzer. The moisture of the samples was removed (degassed) for 1 h at 150 °C temperature, and thereafter the samples were cooled off and weighed. The total surface area of activated WTPS was determined based on the surface area of a single nitrogen molecule, following the procedure used by Nadew et al. (2023).

Scanning electron microscopy (SEM) analysis: SEM was used to analyze the surface morphology and shape of the raw WTPS and activated WTPS adsorbent. The samples were dried in an oven to remove the moisture content to improve the quality of the images. The morphological characteristics of the samples were examined using SEM (JSM-IT 300) by following a standard procedure (Nohl et al. 2022).

Fourier transform infrared (FTIR) spectra analysis: The FTIR analysis was used to obtain qualitative data and complementary evidence for the functional group of the sample. FTIR spectra of both raw and activated sludge samples were performed in the region of 400–4,000 cm−1 and at 32 resolutions and 16 scans (Turki et al. 2018). FTIR spectra were obtained using Spectrum (Thermo Scientific IS50 ABX, Germany) with samples prepared by the attenuated total reflection (ATR) disc method. All the spectra were recorded and processed using IR solution software.

Thermal properties: The thermal properties of raw and activated sludge samples were analyzed using differential scanning calorimetry (DSC; SKZ, 1052B). The samples were heated at a rate of 5 °C/min in a range of temperature from 25 to 125 °C with a void pan as reference. The enthalpy (ΔH, J/g) and the onset (Ton, °C), peak (Tp, °C), and offset (Toff, °C) temperatures of the observed transitions were computed from the thermal curves using the Universal Analysis Program 2,000 (TA Instruments) (Saadatkhah et al. 2020).

Batch adsorption experiments

To investigate the phosphate removal by activated sludge, different studies with a variety of operating conditions and parameters were carried out. These parameters are contact time, pH, adsorbent dose, and initial phosphate concentration. A preliminary study based on a one variable at a time approach was conducted to limit the range of parameter levels and examine the effect of factors on the adsorption process. Better yet, the phosphate adsorption process on the surface of activated sludge was optimized using the design expert response surface methodology–central composite design (RSM–CCD) approach after the interaction effect of these parameters was studied. The phosphate molecules and activated sludge surfaces were contacted under various circumstances and stirred at 300 rpm on a hotplate magnetic stirrer to perform these experiments.

The phosphate adsorbing capacity of each sample was evaluated by allowing 3 g of activated samples to adsorb 30 mg/L of phosphate at a pH of 3 for 70 min. The adsorption process was conducted in a 1 L beaker stirred at 300 rpm, and the solid and filtrate were separated through a filter paper. The residual of adsorbate in the filtrate was determined using a UV-spectrometer at 880 nm in which the absorbance was converted into concentration using a calibration curve (Monvisade & Siriphannon 2009). The phosphate removal percentage (R%) of the activated WTPS was calculated using Equation (1) (Harsha Hebbar et al. 2018).
(1)
where C0 (mg/L) denotes the initial phosphate concentration, and Cf (mg/L) denotes the final phosphate concentration.

Adsorption process optimization and validation

The performance and significance of the model equation were statistically evaluated using analysis of variance (ANOVA) and multiple correlation coefficients (R2). The parameters considered in this study, namely, contact time, adsorbent dose, pH, and initial phosphate concentration, were optimized using numerical optimization (RSM–CCD) to obtain a wastewater sludge (WTPS) adsorbent with high phosphate removal efficiency (Table 2). The validity of the optimum condition was confirmed with triplicate experimental data. A mathematical model was developed that relates the response (removal efficiency) with all parameters. Response surface regression for the design response was analyzed using a quadratic model generated by the design expert as given by Equation (2).
(2)
where Y is the response variable, β is the intercept constant, are the main linear effects constants, b1–b6 are the linear–linear coefficients, C1–C4 are the main quadratic effect coefficients, ξ is the error, and X1–X4 are the independent variables (Behboudi-Jobbehdar et al. 2013).
Table 2

Optimization process criteria to solve the quadratic equation

ParametersGoalsLower limitsUpper limitsImportance
Adsorption time (min) Minimum 50 90 
pH In a range 
Adsorbent dose (g) Minimum 
Initial concentration (mg/L) In a range 20 40 
Removal efficiency (%) Maximum 40 90 
ParametersGoalsLower limitsUpper limitsImportance
Adsorption time (min) Minimum 50 90 
pH In a range 
Adsorbent dose (g) Minimum 
Initial concentration (mg/L) In a range 20 40 
Removal efficiency (%) Maximum 40 90 

The criteria for optimization and solving the quadratic equation were set as shown in Table 2.

Study of isotherms and kinetics of adsorption

Isotherm models: In the event of mass transfer within a system or from high concentration to low concentration of a given species, the equilibrium condition is an inevitable phenomenon. To relate the amount of adsorbate present on the surface of the adsorbent and the amount of adsorbate present in a solution when equilibrium is attained, different models were proposed, and the common ones are the Langmuir and Freundlich isotherm models. These models provide basic information and show the distribution of adsorbate molecules in the liquid and solid phases when the system reaches equilibrium (Elkady et al. 2011).

Adsorption isotherm experiments were conducted by allowing 3 g/L of adsorbent to adsorb different initial adsorbate concentrations of 20, 40, 60, 80, 100, and 120 mg/L at a pH of 3.0, and the adsorption process was agitated using a mechanical stirrer for 5 h in a 1 L beaker. When equilibrium was reached, the equilibrium concentrations of filtrate from each experiment were determined using UV-spectroscopy, and the corresponding amount of adsorbate at equilibrium (qe) was calculated from the Langmuir isotherm model, as is given in Equation (3) (Puri & Kumar 2019).
(3)
where Ce (mL/L) denotes the adsorbate concentration at equilibrium, qe (mg/g) denotes the amount of sorbate at equilibrium per unit mass of sorbent, qmax (mg/g) is the complete monolayer sorption capacity at equilibrium, and KL (L/mg) is the Langmuir equilibrium constant.
The data from the experiment conducted were regressed as Ce/qe versus Ce, from which the Langmuir constants qmax and KL were determined from the slope and intercept, respectively. Furthermore, the experimental data were also fitted to the Freundlich isotherm model, which is expressed in Equation (4).
(4)
where KF is the Freundlich constant related to adsorption capacity and n is the Freundlich constant related to adsorption intensity. The Freundlich constants were determined using the straight-line plot of ln(qe) versus ln(Ce) from which n was found from the slope and KF from the intercept.
Adsorption kinetics: To study the rate at which the phosphate was adsorbed on the surface of the activated sludge, pseudo-first-order (PFO) and pseudo-second-order (PSO) adsorption kinetic models were applied (Kumar & Chauhan 2019; Shobier et al. 2020). Using these models, the dynamic of the adsorption process was examined when the adsorbate made contact with the adsorbent surfaces. To scrutinize this process, 3 g/L of the sample was added to 1 L of phosphate solution with an initial concentration of 30 mg/L and a pH of 3.0 in a flask, and the solution was stirred at 300 rpm using a mechanical stirrer at ambient temperature. The rate of adsorption of phosphate molecules on the surface of activated sludge was studied at different contact times of 20–170 min with 10-min differences. Each experiment was conducted by changing only the concentration and keeping other parameters constant. The trace amount of phosphate that was not adsorbed by the adsorbent in the filtrate was quantified using UV-spectroscopy at a wavelength of 880 nm (Khosravi et al. 2018; Bai et al. 2020; Shobier et al. 2020). The amount of phosphate adsorbed with respect to time (qt) was generated using Equation (5) (Zhang et al. 2020). The data (amount of phosphate adsorbed versus time) were fitted to both PFO and PSO kinetics models, and the best fit was selected based on the higher value of R2 (Afolabi et al. 2020; Han et al. 2020).
(5)
where qt (mg/g) is the amount of adsorbate adsorbed at time t, V (L) is the volume of the solution, and m (g) is the mass of the adsorbent.
In addition, the amount of phosphate adsorbed with time and the amount of phosphate adsorbed when equilibrium attained (qe) after 5 h were calculated using Equation (6).
(6)
The simplified form of the PFO kinetics model can be given by Equation (7).
(7)
where K1 (min−1) is the PFO rate constant and t (min) is the adsorption time. K1 and qe can be determined from the slope and intercept of the plot of log(qeqt) versus t, respectively. In addition, the linearized form of the PSO kinetics model is expressed as Equation (8).
(8)
where K2 (g(mg)−1min−1) is the PSO adsorption rate constant. Plotting (t/qt) versus t, the constant K2 and qe were determined from the intercept and slope, respectively.

Regeneration and reuse of activated WTPS adsorbent

The regeneration and reusability of the activated sludge adsorbent were studied by allowing 3 g of the adsorbent to adsorb 30 mg/L phosphate solution at a pH of 3 for 70 min. The content was shaken at 300 rpm for 70 min to give it enough time so that the adsorbent could take the maximum adsorbate. However, the desorption process was affected by adding 500 ml of 0.1 M NaOH solution to the saturated adsorbent and shaking for 2 h. Finally, it was washed several (three) times using distilled water until the pH became neutral. The supernatant solution was analyzed and the removal capacity was determined. The adsorption–desorption process was repeated five times until the capacity of the adsorbent fell to a considerable level.

Activated WTPS adsorbent and parameters effect on the activation process

The effect of the activation process parameters (acid concentration, time, and temperature) on the performance of phosphate impurities removal from water was investigated (Figure 1(a)–1(c)). The effect of acid concentration on the removal efficiency of the activated WTPS in the event of adsorbing phosphate molecules at a given time and temperature is shown in Figure 1(a). At an activation time of 4 h and an activation temperature of 75 °C, the removal efficiency was observed to increase from 64 to 86% when the acid concentration increased from 1 to 3 M. The removal efficiency, however, decreases to 74% as the acid concentration further increases to 5%. The improvement of removal efficacy with acid concentration suggests that the activation process had taken effect on the sample surface to change into a substantial amount. The decrement in removal efficiency at relatively higher acid concentrations, however, could be due to the degradation of the sample matrix structure (Neolaka et al. 2023). The higher acid concentration used in the activation process would also damage the pores formed and cause the surface to be less and the removal efficiency to decline as compared to the moderate acid concentration (Sumari et al. 2018).
Figure 1

Effect of activation parameters such as (a) acid concentration, (b) activation time, and (c) activation temperature on the removal efficiency of activated sludge.

Figure 1

Effect of activation parameters such as (a) acid concentration, (b) activation time, and (c) activation temperature on the removal efficiency of activated sludge.

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The removal efficiency of the activated sludge with respect to the activation time at constant acid concentration and activation temperature is demonstrated in Figure 1(b). The activation time varied from 2 to 6 h at 1 h intervals while the acid concentration and activation temperature were kept constant at 3 M and 75 °C, respectively. It can be seen from the figure that the removal efficiency increases at a considerable rate from 58.6 to 84.1% as the activation time is changed from 2 to 4 h. Afterward, however, the efficiency changes by a small margin and tends to be almost constant to 85.4% efficiency when the activation time approaches 6 h. As time goes by, it seems that some volatile components that assume space in the sample volume escape, and pores, as a result, might be created. It also seems that the pores formed are the reason for the removal efficiency increase in this particular instance. Moreover, it is intuitive that the removal efficiency leans toward a more or less constant value even if the activation time increases yet more.

In a quite similar way, Figure 1(c) illustrates how the variation of activation temperature affects the removal efficiency of the activated WTPS in terms of phosphate molecules adsorbing. The result ascertains that the removal efficiency rapidly increases from 60 to 82.5% due to the change of activation temperature from 55 to 75 °C. Thereafter, the impact of the activation temperature on the activation process of the adsorbent and hence on the phosphate removal efficiency was less. The removal efficiency only increases from 82.5 to 84.5% despite the activation temperature increases from 75 to 95 °C. Given a 3 M acid concentration and 4 h activation temperature, it is evident that the activation temperature has a sound effect on the activation process of the sludge adsorbent. A more or less similar pattern can be seen from the results of Zakaria et al. (2021) where the authors indicated the role of activation temperature even at lower levels.

After these experiments were conducted, it appears on the surface that an acid concentration of 3 M, an activation time of 4 h, and a 75 °C activation temperature are typical parameters to activate wastewater sludge having about 85% phosphate removal efficiency. As a result, a mass synthesis of activated sludge at this condition was made, and the product was characterized to see whether it fulfills the standard requirement of adsorbents. The activated sludge adsorbent produced at the aforementioned conditions is shown in Figure 2 with raw WTPS and oven-dried WTPS.
Figure 2

(a) Raw WTPS, (b) oven-dried WTPS sample, and (c) activated sludge WTPS.

Figure 2

(a) Raw WTPS, (b) oven-dried WTPS sample, and (c) activated sludge WTPS.

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Characteristics of activated WTPS adsorbent

Proximate result analysis: The moisture content, volatile matter, ash, and fixed carbon content of the prepared raw and activated sludge were determined (Table 3). The physicochemical properties of the activated sludge adsorbent were enhanced following the activation process. It is evident that the activated WTPS has a high fixed carbon content, low moisture content, and low ash content. This could be because the molecules' bonds of the WTPS are broken and the carbon content rises due to the thermal and acid activation process (Kassahun et al. 2022).

Table 3

Proximate analyses result of raw and activated WTPS

ParameterRaw WTPSActivated WTPS
Moisture content 15 
Volatile content 30 29 
Ash content 43 50 
Fixed carbon 12 26 
ParameterRaw WTPSActivated WTPS
Moisture content 15 
Volatile content 30 29 
Ash content 43 50 
Fixed carbon 12 26 

Elemental Analysis: A summary of the CHNS/O analysis for the raw and activated WTPS is provided in Table 4. The raw WTPS was found to have a carbon content of 36.18%, which suggests potential uses for valuable adsorbents. The result obtained here is in the range of 25–60% carbon in the dried sludge as in the previous study by Krotz (2019). When compared to its raw WTPS, the activation of WTPS significantly increased the carbon content to 47.63% and decreased the oxygen content by 20%. This might be the result of some volatile substances being eliminated during the calcination process' chemical treatment and double-bonded carbon breakdown (Volperts et al. 2021). After the activation process, only very slight changes were seen in the other N and S.

Table 4

Elemental analysis of raw and activated WTPS

ElementsElemental compositions (%)
Raw WTPSActivated WTPS
Carbon (C) 36.18 57.63 
Hydrogen (H) 5.14 4.74 
Nitrogen (N) 2.26 2.19 
Sulfur (S) 1.05 1.03 
Oxygen (O)a 54.37 34.40 
ElementsElemental compositions (%)
Raw WTPSActivated WTPS
Carbon (C) 36.18 57.63 
Hydrogen (H) 5.14 4.74 
Nitrogen (N) 2.26 2.19 
Sulfur (S) 1.05 1.03 
Oxygen (O)a 54.37 34.40 

a100 − (C + H + N + S).

SEM analysis: The results of SEM experiments of raw and activated sludge are shown in Figure 3. The SEM image revealed that the activated sludge is more porous than a precursor. The image is evidence that the requirements of adsorbent are met in terms of porosity and size scale, which are important parameters for an adsorption process. Moreover, it seems that some volatile components have been, indeed, removed from the activated sludge adsorbent due to the preparation and activation processes. In recent research, Calderón-Franco et al. (2021) discussed that the morphology of the activated sludge was way more porous than the raw material due to thermal treatment.
Figure 3

The morphology of (a) raw and (b) activated WTPS using SEM.

Figure 3

The morphology of (a) raw and (b) activated WTPS using SEM.

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Surface area analyses using BET: Adsorption is a surface phenomenon where adsorbate finds a way to be attached to the surfaces of the adsorbent. Surface area, therefore, is the very basic nature of the adsorption process. To determine the surface area of the raw and activated sludge, BET analyses were performed. A raw and an activated sample of 0.6 g weight was prepared at 150 °C for 1 h, and the surface area was analyzed in a helium environment and a nitrogen adsorption–desorption system. The surface areas of the raw and activated sludge were found to be 10 and 279.2 m2/g of the sample, respectively. It is evident that the activation process synthesized a relatively high surface area of adsorbent. Some authors like Ros et al. (2006) discussed that the surface area of activated sludge was found to be 200–400 m2/g of the sample when activated in different stages and using different activating agents and impregnation ratios.

FTIR spectra analysis: The FTIR spectra of raw WTPS and activated WTPS adsorbent are shown in Figure 4. As can be seen from the peaks of the figure, there are function groups for both raw and activated sludge at 3,274.46, 2,921.75, 2,853.27, 1,743.01, 991.89, and 521.65 wavenumbers (cm−1). The broad peak at 3,274.46 wavenumbers (cm−1) may show the existence of a trace amount of water, and it is evident that the activated sludge adsorbent contains a lower amount of water compared to the control due to activation processes. The peaks at 2,921.75 and 2,853.27 may be attributed to the presence of the C–H stretching group, probably an alkane compound. In addition, the strong peak at 991.89 might show the presence of the C = C bending group (Kowalski et al. 2018).
Figure 4

FTIR spectra of raw WTPS and activated WTPS adsorbent.

Figure 4

FTIR spectra of raw WTPS and activated WTPS adsorbent.

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Thermal properties analysis: The thermal properties of both samples (raw WTPS and activated WTPS adsorbent) were evaluated based on the parameters calculated from DSC curves as shown in Figure 5, where the samples underwent both endothermic and exothermic processes. The activated WTPS and raw WTPS peak temperatures were around 97.4 and 103.3 °C, respectively. The onset and offset temperatures of activated sludge were 31.2 and 131.3 °C and for raw (prepared) sludge were 42.3 and 128.3 °C, respectively. The activated sludge has lower onset and peak temperature values as compared with the raw sludge.
Figure 5

DSC analyses of raw WTPS and activated WTPS adsorbent.

Figure 5

DSC analyses of raw WTPS and activated WTPS adsorbent.

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Point zero charge of activated sludge adsorbent: The graph of pH change and initial pH values is shown in Figure 6. It was found that the point zero pH of the activated sludge adsorbent was 5.2. As a result, it can be stated that the surface of the activated sludge adsorbent would assume a net positive charge below a pH of 5.2 and a net negative charge above a pH of 5.2. Moreover, it is apparent that for adsorbates that are negative charge in nature, the solution in which the adsorption takes place should be below a pH of 5.2.
Figure 6

Point zero pH of activated sludge adsorbent.

Figure 6

Point zero pH of activated sludge adsorbent.

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Adsorption process parameters effect and optimization

Effect of individual parameters

In this preliminary study, the influence of specific parameters on the adsorption process was investigated by varying one parameter while keeping the other two constants. Consequently, the contact time in the range of 30–150 min, adsorbent dosage from 1 to 7 g/L, pH from 1 to 7, and phosphate initial concentration from 10 to 70 mg/L have been properly considered (Figure 7(a)–7(c)). Overall, the removal efficiency was found to vary from 50 to 86% as the parameters change within the abovementioned range. The effect of contact time on the phosphate adsorption process by the activated sludge adsorbent at a given pH, adsorbent dose, and initial phosphate is shown in Figure 7(a). The removal efficiency showed a significant increase from 50 to 82% when the contact time increased from 30 to 70 min. Up next, however, the removal efficiency increases 84.4% at a very slow rate though the contact time increases to 150 min. The increment of removal efficiency at any rate when the contact time stepwise rises is straightforward. Longer time allows phosphate molecules to pass all adsorption barriers; film diffusion, intraparticle diffusion, and physical or chemical adsorption of molecules on the active sites (Inglezakis et al. 2020). It can be seen from the research of Saad et al. (2012) that the adsorption capacity rapidly increases when the contact time varies from 20 to 80 min and remains almost constant thereafter.
Figure 7

Effect of (a) contact time, (b) pH, (c) adsorbent dose, and (d) initial concentration on the removal efficiency of activated sludge adsorbent.

Figure 7

Effect of (a) contact time, (b) pH, (c) adsorbent dose, and (d) initial concentration on the removal efficiency of activated sludge adsorbent.

Close modal

Figure 7(b) depicts the removal efficiency of the adsorbent when the pH of the solution varies and contact time, adsorbent dose, and initial concentration are kept constant. It can be observed from the graph that the removal efficiency has a decreasing pattern with respect to the solution pH. From a pH of 1 to 3, it slowly decreases from 86 to 83.5% and sharply drops down to 51.8% as the pH further increases to 7. Since the surface of the adsorbent is positive at lower pH and the adsorbate has a negative nature at all times, it seems quite evident that the removal efficiency is higher at lower pH values. At relatively higher pH values, the adsorbent becomes negatively charged surfaces where both adsorbate and adsorbent assume similar charges. Vunain et al. (2021) indicated a high removal efficiency of chromium at pH in the range of 2–4 and showed a decline in efficiency at a higher pH solution.

The removal efficiency as a function adsorbent dose at constant contact time, pH, and initial concentration is demonstrated in Figure 7(c). As the adsorbent was changed from 1 to 3 g/L step by step, the removal efficiency was found to rise quickly from 60 to 83.8%. When the adsorbent was further increased to 7 g/L with an interval of 1 g/L, however, the efficacy of the activated sludge adsorbent in removing the phosphate molecules from the synthetic solution increased only by a small margin. It seems that it tends towards a constant value, in the vicinity of 85% or so. Both the rapid increment and the slow change of removal efficiency with a linear variation of adsorbent dosage are perceivable. For a given 30 mg/L initial concentration, the removal efficiency is expected to rise sharply as there is enough room for adsorption. At higher adsorbent doses, the adsorbate might be taken up or equilibrium may be attained that cause the removal performance to tend to a constant value. The result is in line with the result found by Panda et al. (2017) where the authors prepared an adsorbent by varying the contact time to 0–70 min, pH to 2–7.5, adsorbent dose to 1–10 g/L, and initial concentration to 10–50 mg/L. It was indicated that the removal efficiency increased and then approached constant values as the adsorbent dose increased (Panda et al. 2017).

At a constant contact time of 70 min, a pH of 3, and an adsorbent dose of 3 g/L, the effect of initial concentration on the performance of the activated sludge is illustrated in Figure 7(d). As with the pH, the general pattern of removal efficiency decreases as the initial concentration increases. It slowly decreases from 86 to 84% when the initial concentration increases from 10 to 30 mg/L, and thereafter it falls sharply to 50.2% as the initial concentration changes to 70 mg/L. It is apparent that the decrement in removal efficiency is with respect to the initial concentration. For a given 3 g/L, removal efficiency was expected to decrease as the initial concentration increased. Being constant implies that there is a fixed amount of surface area to adsorb phosphate molecules. When the initial concentration increases, it means that the phosphate molecules would be way more for the given adsorbent amount used. As a result, the removal efficiency declines sharply when the initial phosphate concentration upsets. Yang et al. (2022) and Chakraborty et al. (2022) showed that removal efficiency decreased when initial concentration varied in the range of 20–60 mg/L.

Combined effect of parameters on the adsorption process and model validation

Model validation and ANOVA: Based on the preliminary study, the range of the variable was framed and 30 runs were conducted. The RSM–CCD approach outlines to what extent a pair of independent variables influence the removal efficiency of the activated sludge adsorbent in the event of removing synthetic phosphate solutions. Given the experimental data, a quadratic mathematical equation (model) was developed that predicts the removal efficiency as a function of the four independent variables. The model and the four parameters were statistically evaluated, and all were significant with a p-value of less than 0.0001. In addition, most of the interactive terms were found to be statistically significant (p-value < 0.05). The mathematical model that was obtained from the regression of experimental data and used draw plots to optimize the process is given in Equation (9).
(9)
where RE (%) is the removal efficiency of the activated sludge adsorbent, A is the contact time, B is the pH, C is the adsorbent dose, and D is the initial concentration.
In addition to the p-value, the model was also statistically evaluated using the coefficient of determination, R2 value. The R2 was 0.9776, which is close enough to unity and implies that the experimental data and those data determined by the mathematical model developed are more or less similar. The deviation between the actual data generated by conducting 30 experiments and the equivalent data predicted by a model developed are shown in Figure 8.
Figure 8

Actual vs model predicted.

Figure 8

Actual vs model predicted.

Close modal
Interaction effect of parameters: As can be seen from Figure 9(a)–9(d), the removal efficiency was found to vary from 66.7 to 85.6% owing to the interaction effect of the variables. The removal efficiency as a function of contact time and pH at 3 g/L of adsorbent and 30 mg/L of initial phosphate concentration is depicted in Figure 9(a). As can be observed from the graph, the removal efficiency increases from 77 to 83% as the contact time increases from 50 to 70 min and the pH decreases from 4 to 3. The response, thereafter, slowly increases to 85.4% when the contact time and pH further change to 90 min and 2, respectively. The removal efficiency is directly proportional to contact time and inversely proportional to changes in pH. When the adsorption was allowed to take place for a long time, it was apparent that more adsorbate would be adsorbed on the surface of the adsorbent and more phosphate molecules could be removed from the solution. This may enhance the removal efficiency to any extent possible. At a prolonged time, however, the adsorption process might approach equilibrium, a point where no mass transfer or equal mass transfer in both ways is attained, which slows down the removal efficiency. Conversely, low pH favors more adsorption as the surface of the adsorbent becomes more positively charged, which attracts the negatively charged phosphate ions.
Figure 9

3D surface plot of removal efficiency as a function of (a) contact time and pH, (b) contact time and adsorbent dose, (c) initial concentration and contact time, (d) pH and adsorbent dose, (e) pH and initial concentration, and (f) adsorbent dose and initial concentration.

Figure 9

3D surface plot of removal efficiency as a function of (a) contact time and pH, (b) contact time and adsorbent dose, (c) initial concentration and contact time, (d) pH and adsorbent dose, (e) pH and initial concentration, and (f) adsorbent dose and initial concentration.

Close modal

The response surface plot of the removal efficiency at various contact times and adsorbent dose at a given pH of 3 and initial concentration of 30 mg/L is demonstrated in Figure 9(b). A quick rise from 75 to 83.3% in removal efficiency was observed as the contact time and adsorbent dose were increased from 50 to 71 min and 2 to 3 g/L, respectively. Afterward, only a 2.5% increment was found even though both variables were allowed to increase to 90 min and 4 g/L, respectively. In line with the optics that the adsorbent dose increases the room availability for adsorbate, putting aside the mass transfer barrier it causes when piled up, the performance of the activated sludge adsorbent was boosted pretty well with the dose. It seems that both variables were highly interactive (p-value 0.0014) in the phosphate adsorption process on the activated sludge adsorbent.

Figure 9(c) illustrates the impact of contact time and initial concentration on the removal efficiency both at once while pH and adsorbent were kept at 3 and 3 mg/L in that order, respectively. As predicted by the mathematical model developed, there was a substantial change from 76.5 to 83.4% in removal efficiency when the adsorbate was allowed to contact for 22 min from 50 to 72 min with the adsorbent and as the initial concentration of the phosphate was changed between 40 and 30 mg/L. Furthermore, the removal efficiency increased to 85.6% due to the change in contact time of 90 min and when the initial concentration declined to 20 mg/L. The result suggested that higher contact time and less initial concentration are favorable for removing the adsorbate with higher efficacy.

The 3D graph of the removal efficiency of the activated sludge adsorbent, pH, and adsorbent dose is shown in Figure 9(d). The figure was plotted using the data from the mathematical model of the removal efficiency by varying the pH and adsorbent dose from 2 to 4 units and, yet, keeping the other two parameters, contact time and initial concentration, at 70 min and 30 mg/L, respectively. It can be seen from the graph that the response increases with the adsorbent dose and declines when the pH of the solution within which the adsorption was taking place increases. On the other hand, when pH declines from 4 to 2 and adsorbent increases from 2 to 4 g/L, the removal efficiency shows a general increment pattern from 76.4 to 84.4%, but at a different rate.

The interaction effect of both pH and initial concentration during phosphate molecules being adsorbed on the surface of the activated sludge adsorbent is illustrated in Figure 9(e). While studying these variables' effects using the model, contact time and adsorbent dose remain fixed at 70 min and 3 g/L, respectively. It is evident from the graph that removal efficiency increases from 78 to 85% as the pH of the solution declines from 4 to 2 and the initial concentration decreases from 40 to 20 mg/L. It is apparent that low pH and concentration of adsorbate increase the efficiency. This may be attributed to the availability of relatively enough surface in 3 g of adsorbent in a solution for the low, say 30 mg, of adsorbate in a solution. If this holds, then the results can highlight the extent to which the adsorbent capacity would be in removing phosphate molecules.

Based on the model developed from the 30 experimental data inputs, the graph of removal efficiency, adsorbent dose, and initial concentration at a given contact time and pH is shown in Figure 9(f). The graph reveals that the removal efficiency increases proportionally with the adsorbent dose and decreases with the adsorbate concentration. As can be observed from the figure, the removal efficiency increases from 76.5 to 83.8% as the adsorbent dose increases from 2 to 3.3 g/L, and the initial concentration decreases from 40 to 28 mg/L at 70 min and a pH of 3. The removal efficiency slightly increases to 85% when both parameters further change to 4 and 20 mg/L in the same order. The results unfold that the parameters are interactive in the adsorption process (with a p-value of 0.02), and both have more effect together than individually.

Optimized parameters for adsorption of phosphate using activated WTPS

As is disclosed in the interaction effect of variables, the removal efficiency increases with an increment of contact time and adsorbent dose, and with pH and initial concentration, it was observed that the removal efficiency declines. The results imply that there is a trade-off between variables to get higher removal efficiency of the activated sludge adsorbent. As a consequence, a numerical optimization of the adsorption process was made that takes the positive and negative effects of all variables into consideration. The optimum operating conditions for the phosphate adsorption process by the activated sludge adsorbent are a contact time of 70 min, a pH of 3, an adsorbent dose of 3 g/L, and an initial concentration of 30 mg/L with 83% efficiency. To see if this model-predicted optimum condition is consistent with experimental data, an experiment was conducted at these values of the parameters. The experiment was done three times, and average results showed a more or less similar removal efficiency of 83.4%, with only a 1.2% deviation from the model-predicted result. The results were somehow consistent with the result found by Ahmad et al. (2022) in which case the authors got 80% removal efficiency using activated bentonite. In another scenario, however, the removal efficiency would go up to 89% as can be seen from the work of Yapo et al. (2022). Other authors such as Ahmadi & Igwegbe (2018) and Padmavathy et al. (2016) showed 40 mg/L and 2 g/L of optimum initial concentration and adsorbent dose, respectively. In that regard, the results obtained in this study are a bit different. Contact time, however, varies from 50 to 120 min as is shown in the literature (Wen et al. 2019; Elkarrach et al. 2023).

Adsorption kinetics and isotherm study

Adsorption isotherm: The study of adsorption isotherm models for the phosphate as adsorbate and activated sludge as adsorbent unfold which model best describes the adsorption process and the maximum amount of phosphate that the adsorbent can adsorb when equilibrium is attained at a best case scenario. The experimental data were regressed against the Freundlich and Langmuir isotherm model and the regression yielded R2 values of 0.936 and 0.999, respectively (Figure 10). The model constants and adsorption intensity (n) or adsorption capacity (KF) are shown in Table 5. Based on the data fitting and R2 values, it can be concluded that the adsorption process obeys a monolayer pattern phosphate molecule on the surface of the activated sludge adsorbent. As a result, the Langmuir isotherm model was the best fit in the experimental data. The Langmuir isotherm model is a robust model that can handle most adsorption processes mathematically (Pedrosa et al. 2022).
Table 5

Summary of isotherm constants and R2 values

ModelsLangmuir isotherm
Freundlich isotherm
Constantsqm (mg/g)KL (L/mg)R2KF (mg/g)nR2
Values 13.4 0.41 0.999 6.53 5.84 0.9366 
ModelsLangmuir isotherm
Freundlich isotherm
Constantsqm (mg/g)KL (L/mg)R2KF (mg/g)nR2
Values 13.4 0.41 0.999 6.53 5.84 0.9366 
Figure 10

Experimental data fitting with adsorption isotherm models.

Figure 10

Experimental data fitting with adsorption isotherm models.

Close modal
Study of adsorption kinetics: The kinetics of phosphate adsorption of the activated sludge adsorbent were studied to determine the best-fit model. The experimental data were fitted to two adsorption kinetics models, namely, PFO and PSO models (Figure 11). The coefficient of determinations from the regression result of the kinetics models was 0.8502 for first-order kinetics and 0.9882 for second-order kinetics. The model constants (rate constant) with R2 values are shown in Table 6. From the experimental data fitting, the phosphate adsorption using activated WTPS was best studied with a PSO kinetics model.
Table 6

Summary of the adsorption kinetics models

ModelsPFO
PSO
Constantsqe (mg/g)K1 (min−1)R2qe (mg/g)K2 (g(mg)−1min−1)R2
Values 8.7 0.02234 0.8502 8.7 0.1022 0.9882 
ModelsPFO
PSO
Constantsqe (mg/g)K1 (min−1)R2qe (mg/g)K2 (g(mg)−1min−1)R2
Values 8.7 0.02234 0.8502 8.7 0.1022 0.9882 
Figure 11

Adsorption kinetics models: (a) PFO and (b) PSO kinetics.

Figure 11

Adsorption kinetics models: (a) PFO and (b) PSO kinetics.

Close modal

Repeatability capacity of activated sludge adsorbent

The number of times that the activated sludge adsorbent could be used repeatedly was studied using adsorption–desorption processes. As illustrated in Figure 12, the removal efficiency of the activated sludge adsorbent declined from 83 to 70% when it was used three times in a row. When activated sludge adsorbent was further reused five times, the removal efficiency fell to 20%. It can be said that the activated sludge adsorbent could be reused for only two, if not three times, before its removal efficiency falls substantially. A more or less similar result can be seen from the study of Kenneth et al. (2015). The authors showed that the removal efficiency of activated carbon fell from 65 to 18% as the cycle ran from 1 to 5 times.
Figure 12

Reusability test of activated WTPS adsorbent materials.

Figure 12

Reusability test of activated WTPS adsorbent materials.

Close modal

An activated sludge adsorbent was synthesized from wastewater sludge at different conditions. Thermal followed by acid activation produced a cheap and promising adsorbent. The characterization techniques used revealed that the activated sludge is a potential candidate to be used as an adsorbent for phosphate impurities in water. The efficacy of the activated sludge adsorbent was examined by allowing it to adsorb phosphate molecules. It was observed that contact time and adsorbent affected positively the phosphate adsorbing process by the activated sludge, while pH and initial concentration had a negative impact. The adsorption process was scrutinized to pinpoint the capacity of the adsorbent for wastewater treatment. To that end, the process of adsorption was optimized and 70 min contact time, pH of 3, 3 g/L adsorbent dose, and 30 mg/L initial concentration with 83% removal efficiency were obtained as optimal conditions. Adsorption kinetics and isotherm investigation show that the adsorption of phosphate on the surface of activated sludge adsorbent follows the PSO kinetics model and the Langmuir isotherm model. In addition, the reusability test revealed that the activated sludge adsorbent could be reused three times before its removal efficiency falls below 70%. Altogether, the activation and the adsorption experiments revealed that the easily available material that is being disposed of from every water treatment plant can be used for water purification purposes. It can then be concluded that WTPS can be activated and used as an effective and low-cost adsorbent for the phosphate removal at pilot or even large-scale applications with some additional scale upping experiments and modification.

The Department of Chemical and Food Engineering, Faculty of Food and Chemical Engineering, Wollo University, Kombolcha, Ethiopia, Bahir Dar University, Bahir Dar, Ethiopia, and Addis Ababa Science Technology University, Addis Ababa, Ethiopia are all acknowledged by the authors for the assistance in the research laboratory work.

ES contributed to the conceptualization, methodology, writing the draft, formal analysis, and investigation. TST contributed to the resources, data curation, software, and validation. BG contributed to reviewing and editing the writing. TTN and AGA contributed to visualization and supervision. DAM contributed to manuscript editing. After reading the final draft of the manuscript, all authors gave their approval for publication.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.

Acelas
N. Y.
,
Martin
B. D.
,
López
D.
&
Jefferson
B.
2015
Selective removal of phosphate from wastewater using hydrated metal oxides dispersed within anionic exchange media
.
Chemosphere
119
,
1353
1360
.
https://doi.org/10.1016/j.chemosphere.2014.02.024
.
Adhikari
R. A.
,
Bal Krishna
K.
&
Sarukkalige
R.
2016
Evaluation of phosphorus adsorption capacity of various filter materials from aqueous solution
.
Adsorpt. Sci. Technol.
34
,
320
330
.
https://doi.org/10.1177/0263617416653121
.
Afolabi
I. C.
,
Popoola
S. I.
&
Bello
O. S.
2020
Modeling pseudo-second-order kinetics of orange peel-paracetamol adsorption process using artificial neural network
.
Chemom. Intell. Lab. Syst.
203
,
104053
.
https://doi.org/10.1016/j.chemolab.2020.104053
.
Afzaal
M.
,
Hameed
S.
,
Liaqat
I.
,
Ali Khan
A. A.
,
Abdul Manan
H.
,
Shahid
R.
&
Altaf
M.
2022
Heavy metals contamination in water, sediments and fish of freshwater ecosystems in Pakistan
.
Water Pract. Technol.
17
,
1253
1272
.
https://doi.org/10.2166/wpt.2022.039
.
Ahmad
A. Y.
,
Al-Ghouti
M. A.
,
Khraisheh
M.
&
Zouari
N.
2022
Insights into the removal of lithium and molybdenum from groundwater by adsorption onto activated carbon, bentonite, roasted date pits, and modified-roasted date pits
.
Bioresour. Technol. Reports
18
,
101045
.
https://doi.org/10.1016/j.biteb.2022.101045
.
Ahmadi
S.
&
Igwegbe
C. A.
2018
Adsorptive removal of phenol and aniline by modified bentonite: Adsorption isotherm and kinetics study
.
Appl. Water Sci.
8
,
1
8
.
https://doi.org/10.1007/s13201-018-0826-3
.
Ali
H.
,
Khan
E.
&
Ilahi
I.
2019
Environmental chemistry and ecotoxicology of hazardous heavy metals: Environmental persistence, toxicity, and bioaccumulation
.
J. Chem.
2019
,
1
14
.
https://doi.org/10.1155/2019/6730305
.
Al Tahmazi
T.
2017
Characteristics and Mechanisms of Phosphorus Removal by Dewatered Water Treatment Sludges and the Recovery
.
Banks
D.
,
Jun
B.-M.
,
Heo
J.
,
Her
N.
,
Park
C. M.
&
Yoon
Y.
2020
Selected advanced water treatment technologies for perfluoroalkyl and polyfluoroalkyl substances: A review
.
Sep. Purif. Technol.
231
,
115929
.
https://doi.org/10.1016/j.seppur.2019.115929
.
Behboudi-Jobbehdar
S.
,
Soukoulis
C.
,
Yonekura
L.
&
Fisk
I.
2013
Optimization of spray-drying process conditions for the production of maximally viable microencapsulated L. acidophilus NCIMB 701748
.
Dry. Technol.
31
,
1274
1283
.
https://doi.org/10.1080/07373937.2013.788509
.
Belachew
N.
&
Hinsene
H.
2020
Preparation of cationic surfactant-modified kaolin for enhanced adsorption of hexavalent chromium from aqueous solution
.
Appl. Water Sci.
10
,
38
.
https://doi.org/10.1007/s13201-019-1121-7
.
Bensitel
N.
,
Haboubi
K.
,
Azar
F.-Z.
,
El Hammoudani
Y.
,
El Abdouni
A.
,
Haboubi
C.
,
Dimane
F.
&
El Kasmi
A.
2023
Potential reuse of sludge from a potable water treatment plant in Al Hoceima city in northern Morocco
.
Water Cycle
4
,
154
162
.
https://doi.org/10.1016/j.watcyc.2023.07.002
.
Berkessa
Y. W.
,
Mereta
S. T.
&
Feyisa
F. F.
2019
Simultaneous removal of nitrate and phosphate from wastewater using solid waste from factory
.
Appl. Water Sci.
9
,
28
.
https://doi.org/10.1007/s13201-019-0906-z
.
Calderón-Franco
D.
,
Apoorva
S.
,
Medema
G.
,
van Loosdrecht
M. C. M.
&
Weissbrodt
D. G.
2021
Upgrading residues from wastewater and drinking water treatment plants as low-cost adsorbents to remove extracellular DNA and microorganisms carrying antibiotic resistance genes from treated effluents
.
Sci. Total Environ.
778
.
https://doi.org/10.1016/j.scitotenv.2021.146364
.
Chakraborty
R.
,
Asthana
A.
,
Singh
A. K.
,
Jain
B.
&
Susan
A. B. H.
2022
Adsorption of heavy metal ions by various low-cost adsorbents: A review
.
Int. J. Environ. Anal. Chem.
102
,
342
379
.
https://doi.org/10.1080/03067319.2020.1722811
.
Drechsel
P.
,
Qadir
M.
&
Wichelns
D.
2015
Wastewater
.
Springer Netherlands
,
Dordrecht
.
https://doi.org/10.1007/978-94-017-9545-6
.
Elkady
M. F.
,
Ibrahim
A. M.
&
El-Latif
M. M. A.
2011
Assessment of the adsorption kinetics, equilibrium and thermodynamic for the potential removal of reactive red dye using eggshell biocomposite beads
.
Desalination
278
,
412
423
.
https://doi.org/10.1016/j.desal.2011.05.063
.
Elkarrach
K.
,
Omor
A.
,
Atia
F.
,
Laidi
O.
,
Benlemlih
M.
&
Merzouki
M.
2023
Treatment of tannery effluent by adsorption onto fly ash released from thermal power stations: Characterisation, optimization, kinetics, and isotherms
.
Heliyon
9
,
e12687
.
https://doi.org/10.1016/j.heliyon.2022.e12687
.
Han
S.
,
Zang
Y.
,
Gao
Y.
,
Yue
Q.
,
Zhang
P.
,
Kong
W.
,
Jin
B.
,
Xu
X.
&
Gao
B.
2020
Co-monomer polymer anion exchange resin for removing Cr(VI) contaminants: Adsorption kinetics, mechanism and performance
.
Sci. Total Environ.
709
,
136002
.
https://doi.org/10.1016/j.scitotenv.2019.136002
.
Harsha Hebbar
H. R.
,
Math
M. C.
&
Yatish
K. V.
2018
Optimization and kinetic study of CaO nano-particles catalyzed biodiesel production from Bombax ceiba oil
.
Energy
143
,
25
34
.
https://doi.org/10.1016/j.energy.2017.10.118
.
Inglezakis
V. J.
,
Balsamo
M.
&
Montagnaro
F.
2020
Liquid–solid mass transfer in adsorption systems – An overlooked resistance
.
Ind. Eng. Chem. Res.
59
,
22007
22016
.
https://doi.org/10.1021/acs.iecr.0c05032
.
Jung
K.-W.
,
Hwang
M.-J.
,
Park
D.-S.
&
Ahn
K.-H.
2016
Comprehensive reuse of drinking water treatment residuals in coagulation and adsorption processes
.
J. Environ. Manage.
181
,
425
434
.
https://doi.org/10.1016/j.jenvman.2016.06.041
.
Kang
G.
&
Cao
Y.
2014
Application and modification of poly(vinylidene fluoride) (PVDF) membranes – A review
.
J. Memb. Sci.
463
,
145
165
.
https://doi.org/10.1016/j.memsci.2014.03.055
.
Kassahun
E.
,
Fito
J.
,
Tibebu
S.
,
Nkambule
T. T. I.
,
Tadesse
T.
,
Sime
T.
&
Kloos
H.
2022
The application of the activated carbon from Cordia africana leaves for adsorption of chromium (III) from an aqueous solution
.
J. Chem.
2022
,
1
11
.
https://doi.org/10.1155/2022/4874502
.
Kenneth
A.
,
Emmanuel
G. C.
,
Edith
A. B.
,
Stephen
A. E.
,
Omenesa
H.
,
Titus
Y. M.
&
Nwankwere
E.
2015
Regeneration and reuse of neem husk activated carbon in hospital wastewater treatment
.
Int. J. Sci. Technoledge
3
,
154
157
.
Khosravi
R.
,
Moussavi
G.
,
Ghaneian
M. T.
,
Ehrampoush
M. H.
,
Barikbin
B.
,
Ebrahimi
A. A.
&
Sharifzadeh
G.
2018
Chromium adsorption from aqueous solution using novel green nanocomposite: Adsorbent characterization, isotherm, kinetic and thermodynamic investigation
.
J. Mol. Liq.
256
,
163
174
.
https://doi.org/10.1016/j.molliq.2018.02.033
.
Kiprono
P.
,
Kiptoo
J.
,
Nyawade
E.
&
Ngumba
E.
2023
Iron functionalized silica particles as an ingenious sorbent for removal of fluoride from water
.
Sci. Rep.
13
,
8018
.
https://doi.org/10.1038/s41598-023-34357-8
.
Kowalski
M.
,
Kowalska
K.
,
Wiszniowski
J.
&
Turek-Szytow
J.
2018
Qualitative analysis of activated sludge using FT-IR technique
.
Chem. Pap.
72
,
2699
2706
.
https://doi.org/10.1007/s11696-018-0514-7
.
Krotz
A. L.
2019
Elemental Analysis: N, NC, CHNS/O and TOC Characterization of Sewage Sludge by the FlashSmart Elemental Analyzer. Thermo Fisher Scientific. Available from: https://files.yamato-net.co.jp/product/flashsmart/AN-42342.pdf.
Kumar
P.
&
Chauhan
M. S.
2019
Adsorption of chromium (VI) from the synthetic aqueous solution using chemically modified dried water hyacinth roots
.
J. Environ. Chem. Eng.
7
,
103218
.
https://doi.org/10.1016/j.jece.2019.103218
.
Kumari
A.
,
Singh Maurya
N.
,
Kumar
A.
,
Kant Yadav
R.
&
Kumar
A.
2023
Options for the Disposal and Reuse of Wastewater Sludge, Associated Benefit, and Environmental Risk
.
Intech
, p.
13
.
https://doi.org/10.5772/intechopen.109410
.
Lu
Y.
,
Song
S.
,
Wang
R.
,
Liu
Z.
,
Meng
J.
,
Sweetman
A. J.
,
Jenkins
A.
,
Ferrier
R. C.
,
Li
H.
,
Luo
W.
&
Wang
T.
2015
Impacts of soil and water pollution on food safety and health risks in China
.
Environ. Int.
77
,
5
15
.
https://doi.org/10.1016/j.envint.2014.12.010
.
Monvisade
P.
&
Siriphannon
P.
2009
Chitosan intercalated montmorillonite: Preparation, characterization and cationic dye adsorption
.
Appl. Clay Sci.
42
,
427
431
.
https://doi.org/10.1016/j.clay.2008.04.013
.
Mor
S.
,
Chhoden
K.
&
Ravindra
K.
2016
Application of agro-waste rice husk ash for the removal of phosphate from the wastewater
.
J. Clean. Prod.
129
,
673
680
.
https://doi.org/10.1016/j.jclepro.2016.03.088
.
Nadew
T. T.
,
Keana
M.
,
Sisay
T.
,
Getye
B.
&
Habtu
N. G.
2023
Synthesis of activated carbon from banana peels for dye removal of an aqueous solution in textile industries: Optimization, kinetics, and isotherm aspects
.
Water Pract. Technol.
18
,
947
966
.
https://doi.org/10.2166/wpt.2023.042
.
Neolaka
Y. A. B.
,
Riwu
A. A. P.
,
Aigbe
U. O.
,
Ukhurebor
K. E.
,
Onyancha
R. B.
,
Darmokoesoemo
H.
&
Kusuma
H. S.
2023
Potential of activated carbon from various sources as a low-cost adsorbent to remove heavy metals and synthetic dyes
.
Results Chem.
5
,
100711
.
https://doi.org/10.1016/j.rechem.2022.100711
.
Nguyen
M. D.
,
Thomas
M.
,
Surapaneni
A.
,
Moon
E. M.
&
Milne
N. A.
2022
Beneficial reuse of water treatment sludge in the context of circular economy
.
Environ. Technol. Innov.
28
,
102651
.
https://doi.org/10.1016/j.eti.2022.102651
.
Nguyen
M. D.
,
Donaldson
D.
,
Adhikari
S.
,
Amini
N.
,
Mallya
D. S.
,
Thomas
M.
,
Moon
E. M.
&
Milne
N. A.
2023
Phosphorus adsorption and organic release from dried and thermally treated water treatment sludge
.
Environ. Res.
234
,
116524
.
https://doi.org/10.1016/j.envres.2023.116524
.
Nobaharan
K.
,
Bagheri Novair
S.
,
Asgari Lajayer
B.
&
van Hullebusch
E.
2021
Phosphorus removal from wastewater: The potential use of biochar and the key controlling factors
.
Water
13
,
517
.
https://doi.org/10.3390/w13040517
.
Nohl
J. F.
,
Farr
N. T. H.
,
Sun
Y.
,
Hughes
G. M.
,
Cussen
S. A.
&
Rodenburg
C.
2022
Low-voltage SEM of air-sensitive powders: From sample preparation to micro/nano analysis with secondary electron hyperspectral imaging
.
Micron
156
,
103234
.
https://doi.org/10.1016/j.micron.2022.103234
.
Ooi
T. Y.
,
Yong
E. L.
,
Din
M. F. M.
,
Rezania
S.
,
Aminudin
E.
,
Chelliapan
S.
,
Abdul Rahman
A.
&
Park
J.
2018
Optimization of aluminium recovery from water treatment sludge using response surface methodology
.
J. Environ. Manage.
228
,
13
19
.
https://doi.org/10.1016/j.jenvman.2018.09.008
.
Panda
H.
,
Tiadi
N.
,
Mohanty
M.
&
Mohanty
C. R.
2017
Studies on adsorption behavior of an industrial waste for removal of chromium from aqueous solution, South African
.
J. Chem. Eng.
23
,
132
138
.
https://doi.org/10.1016/j.sajce.2017.05.002
.
Pedrosa
M.
,
Maldonado-Valderrama
J.
&
Gálvez-Ruiz
M. J.
2022
Interactions between curcumin and cell membrane models by Langmuir monolayers
.
Colloids Surf. B Biointerfaces
217
.
https://doi.org/10.1016/j.colsurfb.2022.112636
.
Puri
S.
&
Kumar
V.
2019
Adsorption kinetics and isotherms for the removal of rhodamine B dye and Pb +2 ions from aqueous solutions by a hybrid ion-exchanger
.
Arab. J. Chem.
12
,
316
329
.
https://doi.org/10.1016/j.arabjc.2016.11.009
.
Rai
M. K.
,
Shahi
G.
,
Meena
V.
,
Meena
R.
,
Chakraborty
S.
,
Singh
R. S.
&
Rai
B. N.
2016
Removal of hexavalent chromium Cr (VI) using activated carbon prepared from mango kernel activated with H3PO4
.
Resour. Technol.
2
,
S63
S70
.
https://doi.org/10.1016/j.reffit.2016.11.011
.
Rehman
K.
,
Fatima
F.
,
Waheed
I.
&
Akash
M. S. H.
2018
Prevalence of exposure of heavy metals and their impact on health consequences
.
J. Cell. Biochem.
119
,
157
184
.
https://doi.org/10.1002/jcb.26234
.
Renu
B.
,
Agarwal
M.
&
Singh
K.
2017
Heavy metal removal from wastewater using various adsorbents: A review
.
J. Water Reuse Desalin.
7
,
387
419
.
https://doi.org/10.2166/wrd.2016.104
.
Ros
A.
,
Lillo-Ródenas
M. A.
,
Fuente
E.
,
Montes-Morán
M. A.
,
Martín
M. J.
&
Linares-Solano
A.
2006
High surface area materials prepared from sewage sludge-based precursors
.
Chemosphere
65
,
132
140
.
https://doi.org/10.1016/j.chemosphere.2006.02.017
.
Saad
D. M. G.
,
Cukrowska
E. M.
&
Tutu
H.
2012
Phosphonated cross-linked polyethylenimine for selective removal of uranium ions from aqueous solutions
.
Water Sci. Technol.
66
,
122
129
.
https://doi.org/10.2166/wst.2012.133
.
Saadatkhah
N.
,
Carillo Garcia
A.
,
Ackermann
S.
,
Leclerc
P.
,
Latifi
M.
,
Samih
S.
,
Patience
G. S.
&
Chaouki
J.
2020
Experimental methods in chemical engineering: Thermogravimetric analysis – TGA
.
Can. J. Chem. Eng.
98
,
34
43
.
https://doi.org/10.1002/cjce.23673
.
Shobier
A. H.
,
El-Sadaawy
M. M.
&
El-Said
G. F.
2020
Removal of hexavalent chromium by ecofriendly raw marine green alga Ulva fasciata: Kinetic, thermodynamic and isotherm studies
.
Egypt. J. Aquat. Res.
46
,
325
331
.
https://doi.org/10.1016/j.ejar.2020.09.003
.
Sumari
M.
,
Sholihah
N.
,
Aisiyah
M. M.
,
Oktaviani
I.
,
Khilmi
N.
&
Prakasa
Y. F.
2018
Effectiveness of modified natural zeolite through acid activation as a catalyst on cellulose conversion into glucose from cotton assisted by ultrasonic
.
J. Phys. Conf. Ser.
1093
.
https://doi.org/10.1088/1742-6596/1093/1/012011
.
Turki
A.
,
El Oudiani
A.
,
Msahli
S.
&
Sakli
F.
2018
Infrared spectra for alfa fibers treated with thymol
.
J. Glycobiol.
07
,
1
8
.
https://doi.org/10.4172/2168-958x.1000130
.
Usman
M. O.
,
Aturagaba
G.
,
Ntale
M.
&
Nyakairu
G. W.
2022
A review of adsorption techniques for removal of phosphates from wastewater
.
Water Sci. Technol.
86
,
3113
3132
.
https://doi.org/10.2166/wst.2022.382
.
Volperts
A.
,
Plavniece
A.
,
Kaare
K.
,
Dobele
G.
,
Zhurinsh
A.
&
Kruusenberg
I.
2021
Influence of chemical activation temperatures on nitrogen-doped carbon material structure, pore size distribution and oxygen reduction reaction activity
.
Catalysts
11
,
1460
.
https://doi.org/10.3390/catal11121460
.
Vunain
E.
,
Njewa
J. B.
,
Biswick
T. T.
&
Ipadeola
A. K.
2021
Adsorption of chromium ions from tannery effluents onto activated carbon prepared from rice husk and potato peel by H3PO4 activation
.
Appl. Water Sci.
11
,
1
14
.
https://doi.org/10.1007/s13201-021-01477-3
.
Wen
T.
,
Zhao
Y.
,
Xu
Y.
,
Guo
J.
,
Fu
G.
,
Cheng
Y.
&
Zhong
Y.
2019
Optimization of process parameters and kinetics of adsorption treatment of thallium-containing wastewater
. In:
E3S Web Conference
,
Vol. 118, pp. 01025. https://doi.org/10.1051/e3sconf/201911801025
.
Xu
K.
,
Tao
H.
&
Deng
T.
2016
Removal of phosphate from coating wastewater using magnetic Fe-Cu bimetal oxide modified fly ash
.
J. Water Reuse Desalin.
6
,
430
436
.
https://doi.org/10.2166/wrd.2015.105
.
Yan
Y.
,
Sun
X.
,
Ma
F.
,
Li
J.
,
Shen
J.
,
Han
W.
,
Liu
X.
&
Wang
L.
2014
Removal of phosphate from wastewater using alkaline residue
.
J. Environ. Sci.
26
,
970
980
.
https://doi.org/10.1016/S1001-0742(13)60537-9
.
Yapo
N. S.
,
Aw
S.
,
Briton
B. G. H.
,
Drogui
P.
,
Yao
K. B.
&
Adouby
K.
2022
Removal of fluoride in groundwater by adsorption using hydroxyapatite modified Corbula trigona shell powder
.
Chem. Eng. J. Adv.
12
.
https://doi.org/10.1016/j.ceja.2022.100386
.
Zakaria
R.
,
Jamalluddin
N. A.
&
Abu Bakar
M. Z.
2021
Effect of impregnation ratio and activation temperature on the yield and adsorption performance of mangrove based activated carbon for methylene blue removal
.
Results Mater.
10
,
100183
.
https://doi.org/10.1016/j.rinma.2021.100183
.
Zhang
Y.
,
Wang
Y.
,
Zhang
H.
,
Li
Y.
,
Zhang
Z.
&
Zhang
W.
2020
Recycling spent lithium-ion battery as adsorbents to remove aqueous heavy metals: Adsorption kinetics, isotherms, and regeneration assessment
.
Resour. Conserv. Recycl.
156
,
104688
.
https://doi.org/10.1016/j.resconrec.2020.104688
.
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