Water purification using an eco-friendly adsorbent Open Access

It is an undeniable fact the planet of Earth is facing a wide range of serious environmental problems, ranging from climate change (Salah et al. 2018, 2022) to air pollution (Al-Sareji et al. 2021; Grmasha et al. 2021) and production of solid wastes (Abdulredha et al. 2017). However, water pollution is the most significant dilemma due to the limited sources of fresh water on this planet (Al-Hashimi et al. 2021; Arab et al. 2022). The increasing utilisation of dyes in many industries, such as textile, pharmaceutical, food, paper, and leather, has resulted in serious environmental damage (Yagub et al. 2014; Hashim et al. 2019). Coloured effluents can harm water sources as dyes have a considerable solubility in water, even in low concentrations. In addition, dyes are undesirable in drinking water, minimise sunlight, and reduce photosynthesis and oxygenation in water bodies (Albadarin et al. 2014; Omran et al. 2019). Dye pollution of water bodies may be toxic to aquatic creatures, preventing organismal degradation and disrupting biological cycles. Furthermore, toxic dyes induce skin irritation, allergic dermatitis, genetic mutations, and cancer risk (Chung 2016). Dyes' removal from aqueous solutions is constrained by several factors, including high prices, the generation of hazardous substances, and high energy requirements. So to reduce dye concentration in wastewater, cost-effective, efficient, and ecologically friendly technology must be developed (Tahir et al. 2016; Abdulhadi 2022). Different approaches might be used to remove dye from coloured effluents. Biodegradation, solvent extraction, ion exchange, oxidation, and adsorption are some of the processes involved (Sivarajasekar & Baskar 2015; Sharma & Kaur 2018; Hashim et al. 2021b). For example, the electrocoagulation method is currently in use for dye removal from solutions (Emamjomeh et al. 2020; Abdulhadi et al. 2021; Hashim et al. 2021a). Hashim et al. (2019) used an electrocoagulation cell to remove black reactive dyes from water under different experimental conditions and found the removal efficiency reached 96%. Removal of dyes using plants is another common method nowadays; for example, Karaghool et al. (2022) used sweet basils to remove a textile dye from water. Although the sweet basils removed 93% of 25 mg/L of the dye, the required time was very long (about ten days). Adsorption is one of the most attractive methods for dye removal because it is simple, dependable, ecologically safe, and cost-effective. Activated carbon, which is described as a carbonaceous solid with a high surface area and substantial microporosity, has been used to filtrate dirty water for a few decades (Soonmin 2018). It is also being utilised as a conventional adsorbent more regularly (Dias et al. 2007). Artificially activated carbon, on the other hand, is prohibitively expensive. As a result, several studies have focused on the production of activated carbon from a variety of low-cost materials, including orange peel (Teka & Enyew 2014), solid pineapple waste (Mahamad et al. 2015), palm shell waste (Wong et al. 2016), and peanut shell (Georgin et al. 2016). In this context, this research aims to check the usability of a natural plant-activated carbon to remove methylene blue dye (MBD) from the aqueous medium in batch mode.

The biosorbent employed in this study was activated carbon produced from a native Iraqi plant (Schanginia/sp.) (SAC). This plant may be found in many locations in Iraq. The activation of the corresponding material by carbonisation was used to produce activated carbon. The plant was chopped into small pieces and carbonised for one hour at 250 °C in a muffle furnace (Thermo Scientific TY408X-2A, USA); the furnace was heated first to 250 °C before starting the carbonisation process, which means the plant pieces were subjected to a temperature of 250 °C for one hour. The level and duration of heating were determined through a series of tests with various degrees and intervals. As the level of heating and period were raised, the plant was totally carbonised to ash. The activated carbon was rinsed many times with distilled water before being baked in an electric oven at 100 °C to remove any excess moisture and stored in a desiccator. The biosorbent was finally crushed and sieved using a 100 mesh sieve.

The BET surface area of SAC was calculated using nitrogen adsorption-desorption isotherms acquired at a relative pressure of 0.35 p/p^o using a Thermo Finnegan analyser (Qsurf-9600, USA). Different assays were conducted before and after the biosorption procedure to validate the biosorption of MBD onto SAC. Using a KBr tablet, Fourier-transform infra red (FTIR) spectroscopy (SHIMADZU, IRPrestige-21, Japan) was used to detect the presence of functional groups on the SAC surface in the range of 4,000–400 cm⁻¹. The changes in SAC surface and form before and after MBD removal were studied using scanning electron microscopy (SEM) (TESCAN, Mira3, France), which produces high-resolution images and elemental analysis using energy dispersion spectroscopy (EDS). SPM was used to investigate the topography of SAC (SPM AA300, Advanced Angstrom Inc., USA, with contact style of AFM). The pH value when the end and the initial pH are the same and the biosorbent surface charge is neutral at this value (point zero charges, pH_pzc) is an essential feature for determining which ionic types may be absorbed by the biosorbent at the desired pH. The pH_pzc of SAC is calculated using the batch equilibrium approach, as described before (Mohseni-Bandpi et al. 2016). SAC (0.5 g) was put in conical flasks containing 50 ml 0.1M NaCl solution. The main pH (pH_i) values (2–9) were arranged using a 0.1 M NaOH or HCl solution. After that, the suspensions were agitated for 24 hours at 150 rpm. After that, SAC was separated, and the solution's pH value was measured. The pH_pzc is calculated by graphing the pH vs pH_i curve from the point of interception zero.

In this research, analytically ranked reagents were employed. The adsorbate, MBD (Basic blue 9), was obtained from Sigma Aldrich. [7-(Dimethylamino) phenothiazine-3-ylidene]- chloride of dimethylazanium is the IUPAC systemic name for this compound, CI 52,015, molecular mass 319.85 with a maximum absorption wavelength of 664 nm; molecular formula = C₁₆H₁₈N₃ClS. Batch mode was used to research MBD biosorption. The pH was regulated with a 0.1 M solution of NaOH and HCl.

EDS analysis was used to get the elemental analysis of SAC, which is presented in Table 1.

The SAC sample has an extremely high carbon concentration compared to other elements. As a result, it can be concluded that carbon is the main component in the sample, indicating that it is carbonaceous and suitable for adsorption. S had a very low extent, indicating that just a few contaminants were present. Table 2 lists the various physicochemical features of SAC.

FTIR is a spectroscopy method used to evaluate functional groups located on the surface of biosorbents that may be involved in MBD interactions and biosorption (Giannakoudakis et al. 2018). Figure 1(a) and 1(b) show the FTIR spectra of SAC in the 400–4,000 cm⁻¹ range before and after MBD removal to explain the positions of active sites of these functional groups and the shifting of their locations after MBD removal. The presence of a significant number of IR bands, which appear to be crucial to many functional groupings, in agreement with their own wavelengths, demonstrates the complex character of SAC in Figure 1(a) (cm⁻¹). The –OH stretching vibrations of the hydrogen-bound hydroxyl group of polymeric substances such as carboxylic acids, phenols, and alcohols in the group of pectin and cellulose located on the surface of the biosorbent are characterised by high and wide absorption peaks in the range of 3,414–3,236.55 cm⁻¹. The beneficial oxygen groups on the surface of the SAC significantly boost its hydrophilic properties and serve as binding sites for contaminating organic molecules (Chen et al. 2015). In this area, the –NH asymmetric stretching of amino groups was also seen (Latif et al. 2019). The peaks at 2,924.09 and 2,852.72 cm⁻¹ are symmetric and asymmetric, respectively, for the C-H stretching vibrations of aliphatic acids (Latif et al. 2019). The peaks at 1,637.56 and 1,618.28 cm⁻¹ correspond to C = O stretching associated with carboxyl organic groups and bend vibrating of the functional group –OH, respectively (Silva et al. 2019). The peak at 1,460.11 cm⁻¹ reveals the stretching vibrations of aromatic C = C groups in aromatic rings (Sever et al. 2019). The –COO- symmetric stretching vibrations and C-N stretching vibrations have maxima at 1,375.25 and 1,323.17 cm⁻¹, respectively (Das et al. 2020). N-O stretching vibrations of aliphatic amines (Ansari et al. 2016). IR bands show stretching vibrations of C-O and C-O-C groups of carboxylic acids, phenols, alcohols or esters are shown by IR bands at 1,000–1,300 cm⁻¹ (Jawad et al. 2018). C-N stretching vibrations are represented by the peaks at 785.03 and 719.45 cm⁻¹ (Zhang et al. 2013). Furthermore, the presence of alkyl halide compounds with a C-Br (stretched vibration) bond is indicated by the peaks below 700 cm⁻¹ (Kanthasamy et al. 2020). As shown in Figure 1(a), the SAC surface before MBD biosorption is rich in hydroxyl, carbonyl, and carboxylic groups, which act as proton donors to bind cationic MBD as other medium and weak functional groups that contribute to dye removal. New peaks were formed, which indicates the possible involvement of alkyne and ester groups in the adsorption process during the biosorption of MBD onto the surface of the biosorbent surface (Figure 1(b)) (Rao 2021). These differences in peaks before and after adsorption suggested that these functional groups play a major role in MBD biosorption via electrostatic interaction with positively charged MBD molecules. H-bonding interactions between H atoms on the SAC surface and N atoms in the MBD structure are also part of the biosorption mechanism. Other studies have given similar reasons for MBD biosorption employing coal-activated carbon (Jawad et al. 2019).

The SEM method was used to give pictures at different magnifications for investigating the morphology of the SAC surface before and after MBD biosorption to explore the textural morphology of the surface. Before MBD biosorption, the SEM images (Figure 2(a)) revealed an uneven structure with cracks and many non-homogeneous ravines formed on the SAC surface, which give a significant surface area for dye removal. Several noticeable changes occurred after MBD biosorption (Figure 2(b)), such as a smooth surface with glossy and white patches. The appearance of white patches and dots could indicate that some of the dye was removed through physical adsorption (Oliva et al. 2018). Furthermore, the surface of the SAC seemed to be coated in a thick layer of sediments due to absorbed MBD completely filling the pores. This observation was supported by the literature (Jabar et al. 2020).

As shown in Figure 3(a) and 3(b), the SPM test was used to characterise and assess the topography of the SAC before and after MBD biosorption. According to Figure 3(a), the SAC before MBD biosorption exhibits an uneven and deformed surface with considerable strata of uneven heterogeneous holes, indicating a high possibility of MBD molecules being absorbed. On the other hand, Figure 3(b) shows a flatter surface with reduced pore structures following MBD biosorption, implying biosorption and establishment of MBD molecules by the readily reachable pore surrounding regions of the SAC surface (Jawad et al. 2018). Table 3 summarises the most important findings from this study.

The pH of the MBD solution is the most crucial element determining the biosorption process when compared to all other parameters (Vijayalaks et al. 2019). MBD biosorption by SAC is quite sensitive to the pH of the solution. The biosorption of an adsorbent is influenced by the chemistry of the adsorbent surface, which may be altered by altering the primary pH value. Figure 4 depicts the effects of pH on MBD biosorption. SAC's dye removal effectiveness peaks at pH 9, but no significant gains in adsorbent removal performance have been found above that pH; therefore, pH 9 is regarded to be the optimal pH for SAC. Other research findings have also been published related to this one (Mahmoudi et al. 2015). The fact that the surface of the biosorbent is positively charged at lower pH values explains the pH impact on MBD biosorption since the quantity of MBD biosorbed ions rises as the pH value increases.

The effect of pH on MBD removal was investigated based on the zero charge point (pH_pzc) of SAC. The pH at which the adsorbent's surface has a net zero charge is known as pH_pzc. SAC's pH_pzc was determined to be 4.4, as illustrated in Figure 5. This indicates that the SAC surface is protonated by H+ ion absorption at pHpzc, resulting in electrostatic repulsion with the cationic MBD. Furthermore, when pH > pH_pzc, the SAC surface has a negative charge, creating an electrostatic interaction with the cationic MBD. This clearly illustrates that the pH of the reaction mixture must be higher than pH_pzc for the cationic dye to biosorb (Shakoor & Nasar 2017).

The effect of contact time on biosorption has been explored in the design of batch biosorption studies. Contact time between the adsorbent and the adsorbate is important in the biosorption process (Saleem et al. 2016). This period will be useful in determining the reasonableness of the binding and the best time for contaminant confiscation. Interaction time is another key parameter that impacts the biosorption technique's cost-effectiveness by controlling the biosorption operation's kinetics and supervising the motivational practice of a biosorbent for actual application (Mashkoor & Nasar 2019). Figure 6 depicts the impact of contact time on MBD removal by SAC as a biosorbent. It was discovered that increasing the contact duration (75 minutes) increased the removal effectiveness, which remained constant in the 75–120 minutes range, which corresponded to 100% biosorption of the biosorbent. After that, the best time of 75 minutes was chosen for future investigation. It means that the elimination rate slows with time, and finally, a saturation level is reached (ElSayed 2018).

To find out how the biosorbent dosage affects biosorption capacity and MBD removal performance in the aqueous medium, SAC biosorbent dosages varied from 0.1 to 1.5 mg/100 ml, with an initial MBD concentration of 50 mg/l. In terms of biosorbent dosage, Figure 7 depicts MBD ions' percent removal and absorption. The removal effectiveness of MBD improved considerably from 85.92 to 98.68 percent when the dosage of biosorbent increased from 0.1 to 0.8 g, as shown in Figure 7. Further increasing the biosorbent dose over 0.8 g does not result in a significant improvement in MBD biosorption; hence the optimum SAC quantity was determined to be 0.8 g/100 ml. This periodic rise in removal effectiveness can be accompanied by an increase in the accessibility of additional active adsorbent sites for MBD biosorption with an increase in the SAC dosage. However, it is worth noting that biosorption ability has followed a remarkable decline pattern in biosorption capacity for SAC from 42.96 to 3.31 mg g⁻¹ with a rise in biosorbent dose from 0.1 to 1.5 g/100 ml (Mashkoor & Nasar 2020). This might be because increasing the SAC amount increases the sorbent's interactions, such as accumulation, resulting in a considerable decrease in the sorbent's total surface area and, as a result, a decrease in its biosorption capacity (Ahmed et al. 2020).

In biosorption experiments, the initial dye concentration is a critical element. The effect of the main concentration of MBD solution (10–100 mg/l) on removal efficiency and biosorption capacity in this sample is shown in Figure 8. As MBD strength grew from 10 to 100 mg/l, the efficacy of dye removal decreased, with a drop in the range of 100–96.8%. On the other hand, the biomass removal potential rose in direct proportion to the starting concentration of MBD solution. The absorption rose from 1.25–12.1 mg/g when the MBD concentration rose from 10 to 100 mg/l. The clearance effectiveness was also decreased as the dye concentration was increased in several tests. As the MBD strength increases, the number of active removal sites on the biomass surface decreases, and there are no more areas to adsorb dye onto SAC (Geetha et al. 2015). In other investigations, biomass adsorption increased with increasing starting dye concentrations. This may be related to the increased driving force (Zango 2018).

The data of MBD biosorption onto SAC at equilibrium was examined using two commonly used isothermal simulations to describe the interaction between the biosorbent and the adsorbate, with R² and SE being used to assess the consistency of the fitting.

The Langmuir isothermal model assumes that monolayer adsorption occurs, that all adsorption sites at the adsorbent are structurally similar, and that adsorbed molecules at neighbouring sites do not interact (Gimbert et al. 2008). The adsorbent has limited adsorption ability, beyond which no further adsorption can occur when a molecule filled a site and reached equilibrium saturation (Langmuir 1918).

The isothermal version of R_L might be unfavourable (R_L > 1), linear (R_L = 1), favourable (0 < R_L < 1), or irreversible (R_L = 0) (Dargo et al. 2014). From the graph in Figure 10, the Langmuir variable K_L was calculated to be 0.2, and the qm associated with biosorption capacity was calculated to be 33.34 mg/g. Under the aforementioned experiment circumstances, R_L was also discovered to reflect the Langmuir isotherm applicability within the range of (0.0476–0.3333), as shown in Figure 10.

The Freundlich isothermal constants for uptake and surface heterogeneity are K_F (mg/g) (l/mg)1/n and 1/n, respectively (Alquzweeni & Faisal 2020). The slope and intercept of the log q_e vs log C_e graph in Figure 11 may be used to compute the n and K_F values, which are then arranged in Table 4. The component (1/n) is a function of the sorbent surface's heterogeneity, and as 1/n approaches zero, it changes more heterogeneously.

The ability of SAC to absorb MBD was compared to that of other biosorbents. The largest quantity of sorbed MBD was 34.33 mg/g. In addition, SAC had a higher adsorption capacity than the other biosorbents examined, as seen in Table 5 for MBD.

Figure 12 shows the comparisons of the MBD uptake potential between experimental data and those obtained using the Langmuir and Freundlich representations. The Langmuir model, which has a high R² and a low standard error (SE) value, better simulates the experimental data.

Naturally occurring plant-based activated carbon was employed as a low-cost resource to remove MBD from aqueous solutions. S_BET, FTIR, SEM, and SPM examinations were used to characterise the surface of SAC before and after MBD removal. For biosorption tests, several factors such as the initial pH, agitation time, SAC mass, and initial MBD strength were studied. The Langmuir isotherm model was used to predict the SAC equilibrium biosorption of MBD, which resulted in a homogenous monolayer coating of MBD on the SAC surface. The maximal biosorption capacity was observed to be 33.34 mg/g. The MBD biosorption capability of SAC is greater than that of other agricultural biomass, and it may be easily manufactured without any chemical treatment.

For future studies, the new adsorbent could be applied to remove other pollutants, such as heavy metals and other types of dyes.

Data cannot be made publicly available; readers should contact the corresponding author for details.