Introduction

Emerging contaminants (EC) have been widely reported in wastewater in recent times (Naidu et al. 2016). Endocrine disruptive chemicals (EDCs) are regarded as EC which are widely used in anthropogenic activity (Kumar et al. 2022). Bisphenol A (BPA) is one of the most used EDCs. It is abundant in polycarbonate polymers (Abrams and Kremer 2003) and epoxy resins (Ferdosian 2015). Due to the diversified applications, they are in high demand. As a result, BPA production has reached thousands of tons per annum. The BPA can mimic the endocrine hormone leading to negative effects on human health and animals. It also reported that BPA can cause cancerous tumors (Erkekoglu and Koçer-Gümüşel 2017; Hafezi and Abdel-Rahman 2019), may lead to cardiovascular diseases (Chen et al. 2023) and birth defects, and mental disorders even at very low concentrations(Vandenberg et al. 2007; Chen et al. 2023). BPA exposure in the environment at a dose of 10−8 M can encourage the migration of cancer cells (Jia et al. 2019). For marine creatures, the presence of BPA in concentrations between 1000 and 10,000 µg/L can be extremely hazardous (Careghini et al. 2015). So, there is an acute need for the remediation of BPA in water and wastewater.

On the other hand, the use of dyes is growing due to the rapid industrial development to meet the demands of population expansion. Industries producing textiles, paper, wool, rubber, plastics, cosmetics, and medicines use cationic dyes (Mehrizad and Gharbani 2016). Cationic dyes are more poisonous than other type of dyes as they are capable of penetrating cell membrane (Sokolov et al. 2023). MB is one of those cationic dyes. It is carcinogenic and mutagenic, and it can even enter the path of photosynthesis, which can lead to its entry into to food chain (Ali et al. 2019), causing adverse effects on human health (Gharbani et al. 2022).

There are numerous methods, including adsorption, photocatalysis, Fenton’s process, electric coagulation, and biodegradation for wastewater treatment. Nevertheless, these methods do have several drawbacks, including insufficient removal, the formation of sludge, and high running costs (Tabatabaei et al. 2011; Mehrizad and Gharbani 2016; Tursi et al. 2018). Adsorption is a popular water treatment technology for removing waterborne pollutants (Mehrizad et al. 2012) due to its removal efficiency, ease of use, affordability (Mehrizad and Gharbani 2014), high rate of reusability, and minimal production of hazardous sludge. Adsorbents with high carbon contents are typically versatile and can be used effectively to remove both organic and inorganic contaminants (Gupta et al. 2009). These carbonaceous adsorbents have extensive surface area and well distribution of pore sizes (Mehrizad and Gharbani 2014). The activated carbon (AC) was chosen even though alternative materials such as metal–organic frameworks, magnetic nanocomposites, and graphitic material were there, as those materials require more cost, time, and laborious work. However, commercially available AC is expensive (San Miguel et al. 2001), leading to the use of alternative materials such as biowaste (Dural et al. 2011; Tokula et al. 2023), industrial waste (Ali et al. 2012; Soliman and Moustafa 2020), and other substances which can cause damage to the environment if they are not properly addressed.

One viable possibility is discarded tires, which pose a significant disposal challenge. Many of these scrap tires are being used for high-performance road bases and playground covers. These waste tires possess a high calorific value due to their elevated carbon-to-hydrogen ratio. Superior methods are those that conserve energy. Incineration and combustion may come under this, but they are unfit according to risk–benefit analysis as they can cause more damage to the environment (Machin et al. 2017). Numerous investigations and summarization of tire pyrolysis show that pyrolysis is an environment-friendly way to manage waste tires in a profitable way (Zerin et al. 2023). Tire pyrolyzed carbon is a promising source of activated carbon, which minimizes the production cost of activated carbon and reduces the landfill waste (Zerin et al. 2023).

Even though there are attempts to get AC from tire pyrolysis carbon for the removal of several pollutants such as remozal dye (Nogueira et al. 2019), atrazine and ibuprofen (Frikha et al. 2022), BPA (Acosta et al. 2018; Kuśmierek et al. 2020), and for heavy metals (Shahrokhi-Shahraki et al. 2021; Frikha et al. 2022), there is a lag in the study of the capability of the adsorbent of a single treatment in removing a wider range of pollutants.

This study aims to investigate the use of waste rubber tire carbon as an adsorbent with a high carbon content that could have significant adsorption capacity for waterborne pollutants. Activating the tire carbon was done using a chemical method by activating the tire pyrolyzed carbon with KOH at 750 °C for 5 h. Adsorbing model pollutants BPA, a neutral hydrophobic molecule, and MB, a cationic hydrophilic molecule, are used for evaluating the adsorbent. The different factors that affect adsorption, such as contact time, initial pollutant concentration, adsorbent dose, temperature, pH, presence of an electrolyte, and regeneration studies, are examined in a series of experiments to determine the adsorption efficiencies for BPA and MB. Utilizing the equilibrium data obtained to explore kinetics, thermodynamics, and adsorption isotherms.

Experimental

Materials

Tire pyrolyzed carbon used for this study was purchased from India MART, and it is 99%. Potassium hydroxide (KOH) (85%) of SD Fine Chem limited, Sodium chloride (NaCl) (99.9%) (SRL), Sodium hydroxide (NaOH) (85%) (SRL), Bisphenol A (BPA with empirical formula C16H18O2) (97%) (Sigma-Aldrich), Methylene blue (MB with empirical formula C16H18ClN3S.3H2O), Nitric acid (HNO3) (65%) (Avra), Hydrochloric acid (HCl) (32%) (Avra), ethanol (C2H5OH) (99%) (Sigma-Aldrich), and Sodium nitrite (NaNO3) (99%) (Avra) were purchased from distributors and Millipore water was used for all the experiments. All chemicals were analytical grade and used without further purification.

Preparation of adsorbents

The initial treatment was utilizing varying molar concentrations of KOH (1 M, 2 M, 4 M, 6 M, and 8 M) in 50 mL volumes. A total of 10 g of TPC was sieved to obtain particles with a size ranging from 150 to 75 µm. These particles were then combined with a solution of KOH and well-mixed using a magnetic stirrer at 300 rpm. The mixture was heated until the entire water evaporated, and a dried powder was crushed again. The dried powder was thermally treated in a tubular furnace in the nitrogen atmosphere furnace at a temperature of 750 °C for 5 h. Subsequently, the resulting carbon materials are designated as K-TPAC-1, K-TPAC-2, K-TPAC-4, K-TPAC-6, and K-TPAC-8 (Fig. 1).

Fig. 1
figure 1

Preparation of activated carbon from TPC with KOH treatment

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Characterization techniques

The X-ray diffraction (XRD) technique was used to investigate the phase purity and transformation of activated carbon produced from pyrolyzed tires. The diffraction patterns were recorded using Cu Kα radiation (wavelength of 1.5418 Å). A Ni filter was employed to limit the range of 2θ to 5–80°. The step size for scanning was 0.0167°, and the scan rate was 0.0301°/s. The analyses were performed using the MalvernPANalytical, Empyrean (Netherland) equipment. Scanning electron microscopy (SEM) was done to obtain high-resolution images and sample surface information. The instrument used to take the SEM images is the ZEISS, Sigma 360 VP (Germany). Surface area measurements were conducted using a Quanta chrome, Nova 2200e (USA) surface area analyzer, employing the BET (Brunauer, Emmett, and Teller) method. Before the studies, the samples were subjected to degassing at a temperature of 573 K for 3 h to remove any contaminants present on the sample surface. The temperature-programmed desorption of CO2 and CO was carried out with the assistance of a chemisorption system (Microtrac BEL Corporation, BELCAT II (Japan)) equipped with a TCD detector. Experiments with He-TPD were conducted to investigate the amount and type of oxygen groups on the adsorbent surface. For this analysis, samples were carried out at a linear heating rate of 10 K/min with a targeted temperature of 1173 K and a constant helium flow of 30 mL/min. The quantification of the CO and CO2 gases evolved was done by analyzing the signal and the type of oxygen group is predicted based on the temperature at which the gas is evolving. Fourier-transform infrared spectroscopy was used to understand the phase transformation and the functional groups of prepared adsorbents with JASCO, FT/IR 4600 (Japan). X-ray photoelectron spectroscopy (XPS) of prepared adsorbents was done using an Al-K source at 1489.5 eV, emission current, and voltage of 10 mA and 15 kV, respectively. The instrument SCHIMADZU, Axis Supra (Japan), was used for these XPS studies. To test the pHPZC, sodium nitrate solutions (0.1 M) were made with varied initial pH values: 2 to 12 of 25 mL each. To prepare these solutions, the pH is adjusted with 0.05 M aqueous solution of HNO3 or NaOH. For these solutions, 25 mg of adsorbent was added. Those containers are sealed and kept in an orbital shaker for 24 h and then measured for final pH. The plot of initial pH versus change in pH was drawn. Systronics µ pHSystem 36 l (India) digital pH meter is used to measure the pH of the solution. The T90 + UV–Vis spectrometer MB, manufactured by PG Instruments Ltd. in India, was utilized for the quantification of BPA and pollutants.

Adsorption studies

The initial batch studies were conducted using K-TPAC-1, K-TPAC-2, K-TPAC-4, K-TPAC-6, and K-TPAC-8 for BPA adsorption by taking 50 mL of ppm BPA solution and adsorbent dosage 50 mg, to know the best adsorbent among them. Further experiments were continued with K-TPAC-6 adsorbent for BPA and MB adsorption. The influence of initial pollutant concentrations is measured by taking BPA with initial concentrations of 5, 20, 50, and 75 ppm, while MB concentrations were 20, 50, 100, and 200 ppm. The effect of the adsorbent dose was tested at 0.5, 1, and 1.5 g/L. The impact of pH on BPA and MB adsorption is examined in the pH range of 2 to 10. The effect of ions presence on the adsorption of BPA and MB is examined with 0, 1, 5, 10 mM NaCl present in the pollutant solution. The effects of temperatures 298 K, 308 K, 318 K, and 328 K were investigated for MB and BPA adsorption.

The initial concentration of pollutant used is 20 ppm, and the volume taken is 50 mL. The adsorbent dosage used was 1 g/L, and the agitation speed was set at 200 rpm. The adsorption experiments are conducted at room temperature and the agitation speed is set at 200 rpm. Then, the samples collected were analyzed for BPA & MB concentrations with UV–visible spectrophotometers at wavelengths of 276 and 663 nm, respectively. The adsorption efficiency (%) and adsorption capacity qt (mg/g) are calculated with the following formulae (1) and (2).

$$\text{Adsorption efficiency} \left(\%\right)=\frac{{\text{C}}_{0}-{\text{C}}_{t}}{{\text{C}}_{0}} X 100$$
(1)
$$\text{Adsorption capacity} \left({q}_{t}\right)=\frac{{(\text{C}}_{0}-{\text{C}}_{t})V}{\text{m}}$$
(2)

where Co and Ct are the initial and final concentrations of pollutant in ppm, V is the volume of solution in mL, and m is the mass of adsorbent in g. All the experiments are done thrice and meen of values are considered.

Desorption tests were conducted to assess the potential for reuse, regeneration, and interaction between the adsorbate and the adsorbent. For these studies, 0.5 g of adsorbent was initially mixed with 20 ppm 500 mL of pollutant solution and agitated till the equilibrium was reached. The adsorbent was collected by centrifugation. The spent adsorbent is regenerated with ethanol for BPA and with 0.5 M HCl for MB. The spent adsorbent is treated with respective eluting solution for 15 min followed by centrifuging to separate the solvent, which is analyzed for pollutant concentration. This process continued till the maximum amount of pollutant was removed. Then the adsorbent is dried in an oven at 110 °C and is used for the next cycle.

The desorption efficiency (d) of the solvent and regeneration efficiency or reuse capability (r) of the adsorbent can be measured by following Eqs. (10 and11) (Mpatani et al. 2020):

$$\text{Desorption efficiency} \left(d\right)=\frac{\text{quantity of pollutant desorbed}}{\text{total quanity of pollutant on spent adsorbent before desorption}}\text{X 100}$$
(3)
$$\text{Regeneration efficiency}\left( r\right)=\frac{\text{adsorption capacity in }{\text{n}}^{\text{th}}\text{ cycle }({q}_{n})}{\text{adsorption capacity in }{1}^{\text{st}}\text{ cycle}({q}_{1})} X 100$$
(4)

Results and discussion

Characterization of adsorbent

The X-ray diffraction plots of TPC and all activated tire pyrolyzed carbons K-TPAC 1–6 can be seen in Fig. 2a. The broad peaks at 2θ values of 25.4 and 47.6 are due to the response of carbon (0 0 2), (1 0 1) planes. These peaks are more pronounced in activated carbons, which show a more amorphous nature of carbon in the respective adsorbents (Dandekar et al. 1998). There are several small sharp peaks at 2θ values of 28.6, 30.6, and 33.2 are the peaks belonging to (1 0 2), (1 0 3), and (1 0 4) planes of ZnS (JCPDS#892739) along with amorphous carbon peak. The peaks at 2θ values of 31.7, 34.3, and 36.2 are because of (1 0 0), (0 0 2), and (1 0 1) planes of ZnO (JCPDS#891397). The peak at 2θ value of 20.2 is because of (0 4 1) plane is SiO2 (JCPDS#891813).

Fig. 2
figure 2

a X-ray diffraction pattern and b FTIR of TPC and K-TPAC1-8

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The strong and broad FTIR peak in the region of 3550 to 3200 cm−1 is due to intermolecular hydrogen-bonded -O–H stretching, the peak at 1720 cm−1 is C = O stretching carboxylic acid, the medium peak at 1640 cm−1 is conjugated C = C stretching, 1543 cm−1 is cyclic alkene C = C stretching, 1380 cm−1 is O–H or C-H bending vibrations, and the strong and broad peak at 1037 cm−1 is anhydride (O = C–O–C = O)functional group (Zhang et al. 2017; Mopoung and Dejang 2021) (Fig. 2b).

The scanning electron microscope (SEM) (Fig. 3a, b) images clearly show the formation of cavities within the surface; the voids on the surface of the carbons appear to be the consequence of these chemical reagents evaporating during carbonization and its reaction with surface carbon, leaving the area that they had previously filled (Hu et al. 2022), which also explains the increase in surface area with the chemical activation of TPC.

Fig. 3
figure 3

SEM of images of a TPC and b K-TPAC6

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The Brunauer–Emmett–Teller (BET) surface area of the K-TPAC-1, K-TPAC-2, K-TPAC-4, K-TPAC-6, and K-TPAC-8 are found to be 46, 45, 72, 154, 160, and 141 m2/g respectively. An increase in the surface area of the adsorbent is noted up to treatment with 6 M KOH, after which it began to decrease. Other researchers also noticed a similar pattern in their investigations (Sirimuangjinda et al. 2013; Acosta et al. 2018). The reason might be that at very high concentrations of KOH, the wall between adjacent pores may rupture, leading to a decrease in the surface area.

The Helium Temperature-Programmed Decomposition (TPD) studies demonstrate the oxygen group contents on the surface of the adsorbent. Here, it is done to compare the functional groups on TPC and K-TPAC-6. Based on previous studies, it is assigned that carboxylic acids (will generate CO2 gas below 673 K) and lactone (will generate CO2 gas below 920 K), anhydrides (will generate CO2 and CO above 873 K); ethers, phenols, quinones, and carbonyls (are responsible for CO evolution above 1000 K) (Boudou and Prent 2006; Manoj Kumar Reddy et al. 2015; Ibrahim Abouelamaiem et al. 2018). The TPD profiles of the adsorbents are shown in Fig. 4a and the data is given in Table 1. FTIR of K-TPAC-6 after He-TPD is taken to show that the oxygen functional groups have been eliminated due to the evolution of CO and CO2 during TPD studies (Fig. S1). The total quantity of gas desorbed from the surface of K-TPAC-6 is more than the gas evolved from the TPC, which means that K-TPAC-6 has more oxygen functional groups than TPC. The increase in oxygen content can also be supported by FTIR and XPS.

Fig. 4
figure 4

a Temperature programmed decomposition studies. b XPS-Survey wide scan. c X-ray photoelectron spectra C1s deconvolution of TPC and K-TPAC-6. d pHPZC of TPC. e pHPZC of K-TPAC-6

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Table 1 Temperature programmed decomposition studies and the quantity of CO or CO2 gases evolved
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According to the X-ray photoelectron spectroscopy (XPS) survey-wide scan, the oxygen atomic percentages in the TPC and K-TPAC-6 are 7.5% and 12.5% respectively, demonstrating oxygen content increase after activation (Fig. 4b). The defectiveness and oxidation status of the carbon on the adsorbent surface is assessed using XPS deconvolution peaks of C 1 s. The deconvolution peaks of carbon of the adsorbent in Fig. 4a show the peaks at different binding energies 284.1, 285.5, 286, and 288.6, and 290 eV. The peak at 284.1 eV belongs to C–C, C = C bonded carbons; the peak at 285.5 illustrates the defect carbon (Datsyuk et al. 2008; Shao et al. 2021). The peaks at 286, 288.6, and 290 eV belong to C–OH, C = O, and O-C = O (Shao et al. 2021). After activation, both the defect-rich carbon and carbon with oxygen functional group have been increased as shown in Fig. 4c.

The point of zero charge (pHPZC) is the pH where the surface of the adsorbent will be with no charge and above that pH adsorbent surface will have a negative charge and below that it will have a positive charge (Iheanacho et al. 2023). The pHPZC of TPC and K-TPAC-6 are 6.89 and 7.21 respectively (Fig. 4d); such type of increase in pHPZC in KOH-treated tire carbon was also observed in other studies (Acosta et al. 2018).

Adsorption studies

Comparison of adsorption efficiency for BPA removal by all the tire pyrolyzed activated carbons

The evaluation of the adsorption efficacy of all synthesized adsorbents is listed in Table 2. Removal efficiency of 84% was shown by K-TPAC-6, which was superior to that of TPC and other KOH-treated TPC adsorbents. The following studies were conducted using K-TPAC-6 adsorbent. The increased surface area of K-TPAC-6 led to an enhancement of defect-rich carbon in tire pyrolyzed carbon after KOH treatment, resulting in a larger adsorptive capacity.

Table 2 Adsorption efficiency (%) for BPA by various adsorbents prepared
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To understand the equilibrium of adsorption, a series of adsorption experiments were conducted, whereby the adsorbent dosage and starting concentration of BPA and MB were varied. As the contact time increases, there is a corresponding increase in the amount of BPA and MB adsorbed on the surface of the adsorbent until an equilibrium is reached. The adsorption using K-TPAC-6 has a higher efficiency, with equilibrium being achieved within 20 min. The results of the adsorption studies indicate that the initial concentration of the pollutant and the dose of the adsorbent significantly influence the efficacy of adsorption.

Effect of initial pollutant concentration and adsorbent dosage on adsorption efficiency of K-TPAC-6 for BPA and MB

Effect of initial pollutant concentration

The adsorption efficiency for BPA by K-TPAC-6 is 91% with a BPA solution with an initial concentration of 5 ppm, and it decreased to 57% as the concentration of pollutants increased from 5 to 75 ppm (Fig. 5a). The adsorption capacity has increased from 4.5 to 43.3 mg/g (Fig. 5b) (Table 3). Similarly, when the initial concentration of MB was 20 ppm, the effectiveness of MB removal is 99%.

Fig. 5
figure 5

Pollutant initial concentration effect on a, b adsorption efficiency and unit adsorption capacity for BPA removal and c, d adsorption efficiency and unit adsorption capacity for MB removal by K-TPAC-6

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Table 3 Unit adsorption capacity (qe) and adsorption efficiency (%) for BPA and MB adsorption on K-TPAC-6 with varying initial concentrations of pollutants
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However, this efficiency gradually reduced to 71% at a concentration of 200 ppm. Additionally, the qe rose from 19.9 to 70.8 mg/g as the initial concentrations increased, as shown in Fig. 5d and Table 3. At high levels of BPA and MB (75 ppm and 100 ppm), there is a decrease in adsorption efficiency with the increase in pollutant concentration for both pollutants. As the adsorbent surface is overloaded at higher concentrations, the pollutant experiences incomplete adsorption, leftover with more unadsorbed molecules which decreases the adsorption efficiency. The adsorption capacity (qe) is more at higher concentrations of pollutants. This is due to the excess availability of pollutant molecules at higher pollutant concentrations (Ezeh et al. 2017; Gorzin and Bahri Rasht Abadi 2018; Awad et al. 2019).

Effect of adsorbent dosage

As shown in Fig. 6a, the adsorption efficiency for BPA improved from 79 to 89% when the adsorbent dose was raised from 0.5 to 1.5 g/L. The unit adsorption capacity for 0.5 g/L of adsorbent was 31.5 mg/g, but it decreased to 11.8 mg/g for 1.5 g/L, as shown in Fig. 6b. Similarly, the adoption efficiency of MB increased from 86 to > 99% (Fig. 6c), while the adsorption capacity dropped from 34.4 to 13.3 mg/g for adsorbent dosages of 0.5 g/L and 1.5 g/L, respectively, as shown in Fig. 6d and Table 4.

Fig. 6
figure 6

Adsorbent dose effect on a, b adsorption efficiency and unit adsorption capacity for BPA removal and c, d adsorption efficiency and unit adsorption capacity for MB removal by K-TPAC-6

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Table 4 Unit adsorption capacity (qe) and adsorption efficiency (%) for BPA and MB adsorption on K-TPAC-6 with varying adsorbent dosage
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As the total available surface area is more in the case of high adsorbent dosage, the adsorption efficiency is more (Gharbani et al. 2015). At the higher adsorbent dosage, the concentration of pollutant will become a limiting factor, leaving most of the active sites unsaturated or unadsorbed at equilibrium on the surface of the adsorbent; as a result, adsorption capacity qe of the adsorbent decreases with increasing adsorbent dosage (Ezeh et al. 2017; Gorzin and Bahri Rasht Abadi 2018).

Kinetic study of adsorption

Studying kinetics is an important step to understand the adsorption of pollutants by adsorbent with respect to time. We can determine how quickly or slowly the adsorption is occurring. Three kinetic models are selected here; those are pseudo-first order, pseudo-second order, and intra-particle diffusion model.

The linear form of a pseudo-first-order can be expressed as follows with the Eq. (5):

$$\text{log}\left({\text{q}}_{e}-{\text{q}}_{t}\right)=\text{log}{q}_{e}-\frac{{k}_{1}}{2.303}(t)$$
(5)

Here, qe and qt are the amount of pollutant adsorbed in mg/g of adsorbent at equilibrium and at time t respectively. The \({k}_{1}\) is the rate constant of adsorption (min−1) (Lagergren 1898; Ho and McKay 1998; Yuh-Shan 2004; Huang et al. 2014; Gómez-Serrano et al. 2021). The plots of \(\text{log}\left({\text{q}}_{e}-{\text{q}}_{t}\right)\) vs. t for various BPA (5, 20, 50, and 75 ppm) (Fig. 7a) and MB (20, 50, 100, and 200 ppm) (Fig. 7d) concentrations were used to compute the values of \({k}_{1}\)(Table S1).

Fig. 7
figure 7

a Pseudo-first order, b pseudo-second order, and c intra particle diffusion model fits for BPA adsorption with initial concentrations 5, 20, 50, and 75 ppm. d Pseudo-first-order, e pseudo-second-order, and f intra particle diffusion model fits for MB adsorption with initial concentrations 20, 50, 100, and 200 ppm by K-TPAC-6

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The equation for linear form for pseudo-second-order is represented with Eq. (6):

$$\frac{t}{{\text{q}}_{t}}=\frac{1}{{k}_{2}{q}_{e}}+\frac{1}{{q}_{e}}(t)$$
(6)

Here, qe and qt are the amount of pollutant adsorbed in mg/g of adsorbent at equilibrium and at time t respectively. The k2 is the rate constant of second-order adsorption (g/mg/min) (Ho and McKay 1998, 1999; Gómez-Serrano et al. 2021).

The pseudo-2nd-order kinetics model exhibited the highest coefficient values, with an R2 (correlation coefficient) value 0.99. The findings presented in this study align with previous research on the adsorption of BPA (Javed et al. 2018; Ndagijimana et al. 2019) and MB (Hameed et al. 2007; Do et al. 2021) onto adsorbents. The theoretical values of k2 and qe were determined by analyzing the intercept and slope of the linear plot (t/qt versus t for various concentrations of BPA (5–75 ppm) and MB (20–200 ppm) depicted in Fig. 7b and e. The results revealed a strong correlation between the calculated values and the experimental qe values, as presented in Table 5.

Table 5 Pseudo-second-order kinetic parameters for BPA and MB removal by K-TPAC-6 with various initial concentrations
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To learn more about the rate-limiting stage in the adsorption process, the intra-particle diffusion model was utilized. Adsorption generally involves three stages (Rout et al. 2014): pollutant is transported by (1) film diffusion from the boundary film to the adsorbent’s outer surface; (2) intra-particle diffusion from the adsorbent’s surface to its pores; and (3) equilibrium stage adsorption of pollutant on the adsorbent’s inner surface (Tang and Zhang 2016). The linear form of the Morris-weber Eq. (7) is:

$${q}_{t ={k}_{p}{t}^{1/2} + C}$$
(7)

where qt (mg/g) is the amount adsorbed within time t (min), \({k}_{p}\) is the intra-particle diffusion rate constant (mg/g min1/2), and C is the constant indicating the thickness of the boundary layer. In the plot of qt vs. t1/2, \({k}_{p}\) is slope and C is the intercept (Weber and Morris 1963; Patawat et al. 2020). The higher the slope value, the more is the rate of adsorption (Huang et al. 2018). The intra-particle model fit for the adsorption kinetics with initial concentrations of 5, 20, 50, and 75 ppm of BPA and 20, 50, 100, and 200 ppm of MBwere drawn. The plots of qt vs. t1/2 (Fig. 7c and f) for both BPA and MB with all the concentrations are multilinear with different slopes, as these plots are not a straight single line indicate that intra-particle diffusion is not only rate-limiting step (Wang and Guo 2022). The first step slope has more in each case, which indicates that bulk diffusion is showing more adsorption rate than the intra-particle diffusion step (Huang et al. 2018).

Adsorption isotherms

The amount of pollutant adsorbed onto the adsorbent at equilibrium was studied and plotted as a function of the equilibrium concentration of a pollutant at constant temperature is considered as the adsorption isotherm. In this experiment, Langmuir and Freundlich’s isotherms were used to evaluate experimental results.

Langmuir (1918) was first to propose this theory. It is based on certain assumptions: homogenous and fixed number of active sites on the surface of the adsorbent, mono-layer adsorption of adsorbent molecules, and the adsorbed molecules do not interact (Do 1998; Ong et al. 2014).

Langmuir isotherm is expressed as in Eq. (8):

$${q}_{e}={q}_{max}\frac{ {K}_{L} {C}_{e}}{1+{K}_{L}{C}_{e}}$$
(8)

It can be rearranged as Eq. (9)

$$\frac{{C}_{e}}{{q}_{e}}=\frac{1}{{q}_{max{K}_{L}}}+\frac{{C}_{e}}{{q}_{max}}$$
(9)

Here, \({q}_{\text{max}}\) (mg/g) is the maximum adsorption capacity and \({K}_{L}\) (L/mg) Langmuir constant is related to the rate of adsorption. With the plot of \(\frac{{C}_{e}}{qe} \text{vs} {C}_{e}\) (Fig. 11a), we will get an isotherm with the slope of \({(1/q}_{\text{max}})\) and intercept of (1/(\({q}_{\text{max}}\)*KL)). With this, we can predict \({q}_{\text{max}}\) (mg/g) (Langmuir 1918; Elgeundi 1991; Zhuang et al. 2020; Mehrizad 2017).

Freundlich adsorption isotherms can be applied to multilayer adsorption on heterogeneous sites, in contrast to the Langmuir isotherm. It is predicted on the idea that affinities and the heat distribution of adsorption toward the heterogeneous surface are not uniform (Foo and Hameed 2010) and is represented with the Eq. (10).

$${{q}_{e}=K}_{F} {C}^\frac{1}{n}$$
(10)

It can be rewritten as Eq. (11)

$$log{q}_{e}=log{K}_{F}+1/nlog{C}_{e}$$
(11)

Here, the \({K}_{F}\) is the Freundlich constant in mg/g and \(n\) is the Freundlich exponent. A higher n (n > 1) value indicates favorable adsorption (Mehrizad 2017; Zhuang et al. 2020).

The experimental data fits into Langmuir isotherm for BPA and MB adsorption with K-TPAC-6. Figure 8a and c (Table 6) is for linear Langmuir fits for BPA and MB and Fig. 8b and d (Table 6) is for linear Freundlich fits for BPA and MB. The qmax obtained from these Langmuir isotherms is 49.2 mg/g (with initial concentrations of 5, 20, 50, and 75 ppm of BPA), and 72.09 mg/g (with initial concentrations of 20, 50, 100, and 200 ppm of MB) at 298 K. These calculated values are similar to experimental values and the Langmuir isotherm plot exhibited the highest coefficient values than that of Freundlich isotherm.

Fig. 8
figure 8

a, b Langmuir and Freundlich fit for BPA adsorption at 298 K. c, d Langmuir and Freundlich for MB adsorption at 298 K

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Table 6 Adsorption constants of Langmuir and Freundlich isotherms for BPA and MB removal by K-TPAC-6
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Thermodynamic parameters

The study of the impact of temperature on the adsorption process can provide key information on entropy and enthalpy changes. The enthalpy change in chemisorption typically ranges from 80 to 200 kJ/mol, whereas the energy produced by physisorption is 2.1–20.9 kJ/mol (Üzek et al. 2022). The experiments were to know the effect of varying temperatures 298, 308, 318, and 328 K on the adsorption of BPA (5–75 ppm) and MB (20 ppm) by K-TPAC-6. To know the thermodynamic parameters (ΔG°, ΔH°, and ΔS°) are computed from the graph ln kd vs. 1/T Eqs. (12, 13, and14) (Fig. 9a and Table 7).

$${k}_{d}=\frac{{q}_{e}}{{C}_{e}}$$
(12)
$${\Delta G}^{o}=RT ln{ k}_{d}$$
(13)
$$ln {k}_{d}=-\frac{{\Delta H}^{o}}{R}\left(\frac{1}{T}\right)+\frac{{\Delta S}^{o}}{R}$$
(14)

where \({k}_{d}\) is the equilibrium constant, qe pollutant amount adsorbed, and Ce concentration of pollutant at equilibrium (Ratkowsky and Giles 1990; Mehrizad 2017).

Fig. 9
figure 9

Thermodynamic parameters for adsorption of a BPA with initial concentrations 5, 20, 50, and 75 ppm. b MB with initial concentration 20, 50, 100, 200 ppm at 298, 308, 318, and 328 K

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Table 7 Thermodynamic parameters for BPA and MB removal by K-TPAC-6 at various temperatures with different initial pollutant concentrations
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The negative ΔH° value (− 42.3 kJ/mol) for 5 ppm BPA adsorption indicates it is an exothermic process, and this change in enthalpy is found to decrease (− 38.7, − 25.1, and − 11.5 kJ/mol) with the increased initial concentrations (20, 50, and 75 ppm). Thus, the exothermic nature is more for lower concentrated BPA solution as the most energetic sites are indeed filled first at low BPA concentrations (Acosta et al. 2018). Such type of enthalpy change range is also observed by many other researchers for different pollutants (Saleh et al. 2017; Tran et al. 2017; Hu et al. 2022). The values of entropy ΔS° changes are negative indicating that the randomness has been decreased as the pollutant particles moment is arrested as they were adsorbed on the surface of adsorbents. The negative value of Gibb’s free energy ΔG° gives information about the spontaneous process.

The positive ΔH° (15.5 kJ/mol) value for 20 ppm MB solution indicates the endothermic nature of MB adsorption on K-TPAC-6. The values of enthalpy change ΔH° and the entropy change ΔS° for all initial concentrations of MB solution are positive and shown in Fig. 9b and Table 7. The positive values of enthalpy is because of the dissociation of dye to give monomer from dimer or aggregates before it get adsorbed by adsorbent (Dural et al. 2011). In the process of MB adsorption, the positive value of ΔS° indicates that there is a rise in randomness at the interface between the adsorbent and the liquid, which may be accompanied by structural changes or interactions between the molecules of MB and activated Carbon (Bedin et al. 2016). The positive values of ΔH° and ΔS° are also observed by other scientists (Bhattacharyya and Sharma 2005; Hong et al. 2009; Makrigianni et al. 2015; Bedin et al. 2016; Egbosiuba et al. 2020). The Gibbs free energy change ΔG° values are negative indicating the spontaneous nature of adsorption. The thermodynamic parameters of MB adsorption are shown in Fig. 9b and Table 7.

Effect of pH on adsorption of BPA and MB

The K-TPAC-6 has better adsorption efficiency for BPA for a wider range of pH until pH 9. The adsorption efficiency from pH 2 to 9 is shown as 78% to 84% from pH 2 to 9. However, at higher pH-10, the adsorption is eventually decreased to 65%. This fashion is also found by Bautista-Toledo et al. ( 2005). One plausible explanation for this phenomenon is that BPA is in its molecular state while the pH was below 8.0, so there is no influence of pH for the maximum range of pH as shown in (Fig. 10). Then, the process of deprotonation occurred initially, and subsequently at about pH levels 8.0 and 9.0 respectively, resulting in the formation of mono or divalent anions. The decrease in BPA removal was found at higher (pH-10) levels, possibly because of the repulsive electrostatic interactions between the negatively charged surface of the adsorbents and the bisphenolate anion (Bhatnagar and Anastopoulos 2017).

Fig. 10
figure 10

Effect of initial pH on adsorption of BPA and MB by K-TPAC-6

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There is a significant variation in the adsorption of MB at different pH levels. At a lower pH below 7, the adsorption efficiency of MB is less, which indicates the influence of pH on its adsorption. This may be because the surface of the adsorbent and MB molecule both will have a positive charge at lower pH. The removal of MB increases drastically at higher pH (Wang et al. 2005; Malarvizhi and Ho 2010) and attained more than 99% removal. It is because the adsorbent surface has a negative charge which favors the adsorption of MB (Fig. 10). That means there exists electrostatic interactions between pollutant and adsorbent (Kavitha and Namasivayam 2007; Ramaraju et al. 2014; Manoj Kumar Reddy et al. 2015).

Effect of added salt NaCl on the BPA and MB adsorption

The ion effect was conducted for the presence of 1, 5, and 10 mM NaCl along with pollutant to observe their interference with pollutants adsorption by adsorbent. These ions have the capability of hindering the electrostatic interactions of the pollutant with the adsorbent, as they can be involved in those interactions. These ions can also hinder the hydrogen bonding of pollutants with adsorbent by salting out the -OH groups of adsorbents. The K-TPAC-6 has shown an adsorption efficiency of approximately 84% (Fig. 11a) (Table S2) in the presence of all the different molar concentrations of NaCl. There is no effect of these ions presence on the adsorption of BPA. It may indicate that there is no major role of electrostatic interaction or H-bonding in the adsorption of BPA (Shao et al. 2021) by K-TPAC-6.

Fig. 11
figure 11

Adsorption efficiencies a for BPA and b for MB at different ionic concentrations

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The MB adsorption decreased with increased NaCl concentrations. As the NaCl concentration increased from 1 to 10 mM, its adsorption decreased from 99 to 87% as shown in Fig. 11b and Table S2. This fashion of decreasing adsorption of MB in the presence of electrolytes is also seen by Boumediene et al. (2018) and Ali et al. (2019) in their research. It shows that electrostatic interactions play a major role in the adsorption of MB.

Desorption and reusability

The desorption of BPA was done with ethanol as it is found to be a promising desorbing agent rather than HCl, NaCl, and NaOH solutions. The desorption was made with MB 0.5 M HCl, which was found to be more suitable than other solvents. In acidic conditions, the adsorbent surface becomes positive, where the MB can be effectively desorbed. The regenerated adsorbent shows considerable efficiency even in the fifth cycle for both BPA and MB adsorption independently (Fig. 12 a and b) (Table S3).

Fig. 12
figure 12

Adsorption efficiencies a for BPA and b for MB after each cycle by K-TPAC-6

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Plausible mechanism

The tests, including the presence of NaCl (Fig. 11a) and the influence of pH (Fig. 10) on BPA adsorption, suggest that there is no significant effect of electrostatic interactions or hydrogen bonding. From XPS, it was confirmed that KOH treatment leads to defective carbon improvement in the adsorbent and based on the literature, we assume hydrophobic interactions (Sun et al. 2020; Shao et al. 2021) or π-π interactions (Sun et al. 2020) play a significant role in the adsorption of BPA by K-TPAC-6 (Shao et al. 2021), whereas the hydrogen bonding contribution is minimal. In the SEM images of adsorbent after BPA adsorption, we can observe the aggregation on its surface (Fig. S2).

As the presence of NaCl in MB solution (Fig. 11b) and the initial pH (Fig. 10) can affect the MB adsorption efficiency, this indicates the major role of electrostatic interactions and hydrogen bonding between MB and K-TPAC-6 adsorbent (Abdulhameed et al. 2021; Hanjing Xue a,1 et al. 2021; Dolas 2023), along with a small portion of π-π interactions. This can be explained by analyzing the FTIR of the adsorbent before and after MB adsorption (Fig. S3). The band at 3420 cm−1 in activated carbon (K-TPAC-6) shifted to 3440 cm−1 after MB adsorption (K-TPAC-6-MB), as the stretching vibration of O–H combined with the N–H stretching vibration (Pezoti et al. 2014). The peak at 1630 cm−1 is assigned to N–H bending vibration coupled with C-N and the decrease in intensity of peaks in the region of 800 to 500 cm−1 indicates π-π interactions between adsorbent and MB (Fu et al. 2015; Bedin et al. 2016).

The comparative study for the adsorption capacities of BPA and MB by K-TPAC-6 with the literature is shown in Tables 8 and 9 respectively.

Table 8 Comparison with the literature for BPA adsorption
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Table 9 Comparison with the literature for MB adsorption
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Conclusions

The best adsorbent selected is K-TPAC-6 which has a surface area 160 m2/g. It has shown 84% and > 99% removal for BPA and MB respectively, with initial pollutant concentration of 20 ppm. The enhanced BPA removal can be explained by improved defect-rich carbon and surface area (BET and XPS), which provides sites for hydrophobic and \(\pi\)-\(\pi\) interactions between K-TPAC-6 adsorbent and BPA pollutant. The high adsorption of MB is because of improved oxygen-containing groups (XPS, FTIR, and He-TPD) on the K-TPAC-6 surface, which are helpful for MB adsorption with electrostatic interactions and hydrogen bonding. The adsorption of BPA is exothermic and that is endothermic for MB. Neither NaCl presence nor initial pH affects the BPA adsorption, but they can do so for MB adsorption. Adsorption of both the pollutants followed pseudo-second-order kinetics and Langmuir isotherm model. And efficiency of the adsorbent is notable even in 5th cycle. Hence, it is proved that the waste tire is a promising source to synthesize activated carbon, which can show good adsorption efficiency for both BPA and MB from aqueous streams.

The fundamental findings observed in this study can be the basis for future studies on the simultaneous adsorption of these two pollutants (BPA and MB), which is very much essential for present society as the co-existence of endocrine-disrupting chemicals (BPA) and cationic dye (MB) is common in wastewater.