1 Introduction

The discharge of dyes containing wastewater from the leather, textile, paper, and plastics industries has emerged as a serious environmental concern due to their toxic and carcinogenic nature (Sharma et al. 2010). Methylene blue (MB) dye is a cationic dye generally used for dying cotton, wood, and silk (Deng et al. 2011). MB dye exposure can adversely affect human health, including skin and eye irritation, inhaled respiratory discomfort, and toxic reactions if ingested (Hameed & El-Khaiary 2008). It is crucial to handle it with care and avoid misuse. In medical settings, MB is used as a treatment for specific conditions under professional supervision, but improper exposure can lead to health issues. Additionally, its improper disposal can harm the environment, particularly aquatic life. Therefore, it is important to address the issue of wastewater pollution resulting from the discharge of dyes and identify an efficient removal technology for its suppression from industrial effluent.

Several separation technologies are employed to remove dyes from wastewater. The separation technologies for the removal of dyes from wastewater are classified under the categories of physical, chemical, and biological methods (Slokar & Majcen Le Marechal 1998). It can be separated by the commonly used method of coagulation and flocculation, where chemicals are added to form flocs that trap the dye particles. Among them, adsorption is preferred owing to its high removal efficiency and cost-effectiveness (Kulkarni et al. 2017). Adsorption involves materials like activated carbon or certain clays to adsorb dye molecules from the water. Also, it is an attractive method due to its design simplicity, ease of operation, insensitiveness towards toxic pollutants, and a smaller number of harmful substances (Rafatullah et al. 2010; Sharma et al. 2010). Membrane filtration, including techniques like ultrafiltration and reverse osmosis, can effectively remove dyes by physically blocking their passage. Biological treatments, such as activated sludge or constructed wetlands, use microorganisms to break down dyes. Each method has its advantages and limitations, and the choice depends on factors like the type of dye, its concentration, and the specific requirements of the wastewater treatment process.

Biochar has been recognized as a versatile and valuable material with a wide range of environmental applications. It is a byproduct obtained by the carbonization of biomass, which has emerged as an effective biosorbent for eradicating organic contaminants from the aqueous solution (Malik et al. 2017; Yang et al. 2019a, b; Yang et al. 2017a, b; Yang et al. 2017a, b). The physical and chemical properties of biochar are mainly governed by the feedstock type and pyrolysis temperature (Lehmann & Joseph 2009; Ok et al. 2015; Yang et al. 2019a, b). Feedstock can be obtained from agriculture, forestry, industrial waste, and sewage sludge (Shakya & Agarwal 2017). However, the physicochemical properties such as surface area, pH, and functional groups of biochar are highly dependent on the pyrolysis temperature (Ding et al. 2014; Yang et al. 2019a, b). The release of volatile matter may be intensified at the higher pyrolysis temperature, which ultimately may create more internal pores (Shaaban et al. 2014). Interestingly, the potential applications of biochar also include addressing global environmental issues such as climate change mitigation, carbon sequestration, contaminant immobilization, energy production, greenhouse gas reduction, soil amendment, and water purification (Abhishek et al. 2022; Ahmad et al. 2014; Lehmann & Joseph 2009). The unique properties of biochar such as large surface area having various functional groups, porous structure and mineral composition help to effectively remove the contaminants from the wastewater (Kumar et al. 2022; Nartey & Zhao 2014; Sinha et al. 2022a, b; Sinha et al. 2022a, b).

Lantana camara L., commonly known as Lantana, is an invasive weed (family Verbenaceae) primarily native to subtropical and tropical America except for a few taxa indigenous to tropical Asia and Africa (Negi et al. 2019). It has become a significant problem in many regions around the world after this plant was introduced to various countries as an ornamental garden plant due to its colorful flowers. In 1808, this plant species was also introduced to India as an ornamental plant at the National Botanical Garden (Hakimuddin 1930; Kohli et al. 2006). But later, it became a potential invader and adversely impacted the native plant species diversity and ecosystem functioning (Sharma et al. 2005; Sharma & Raghubanshi 2009). The aggressive growth and the ability to produce toxic compounds of Lantana deter herbivores and prolific seed production and allow them to escape cultivation and establish themselves in natural ecosystems. Lantana outcompetes native plants, disrupts ecological balance, and leads to habitat degradation. Controlling its spread is challenging, making it a serious concern for biodiversity conservation and land management efforts in affected areas. The invasion of Lantana in the forest ecosystems has adversely impacted the biodiversity via soil erosion (Day et al. 2003), nurturing vectors carrying communicable diseases (Syed & Guerin 2004), fostering fire fortuity (Hiremath & Sundaram 2005), and loss of native biodiversity due to allelopathic effect (Sharma et al. 2005). In order to manage the growth of the plant species, numerous control measures such as mechanical, chemical and biological have been employed (Negi et al. 2019), but none of them have been able to eradicate its invasion completely. Despite all, this notorious weed finds application in herbal medicine, industrial use, agricultural practices, phytoremediation and phytoextraction of heavy metals (Kumar et al. 2017; Sharma & Sharma 1989). Also, activated carbon synthesized from Lantana was effectively used as an adsorbent in order to remove tartrazine (Gautam et al. 2015), MB dye (Amuda et al. 2014), and phenol (Girish & Ramachandra Murty 2014) from aqueous solution.

The conversion of invasive species into biochar can significantly improve invasive plant management (Dong et al. 2013). It can also valorize its biomass for various environmental applications. Using biochar synthesized from the stem of Lantana invasive plant, MB dye removal was presented (Amuda et al. 2014), but a comparative analysis using their stem and leaf was not studied. Therefore, the main objective of this study was to examine the potential of biochar prepared from the leaf and stem of Lantana for the removal of MB dye from the aqueous solution as a function of solution chemistry, temperature, retention time, and absorbent doses. This was supplemented by kinetic isotherm, and thermodynamic modelling to better understand the correlation of MB dye interaction with biochar particles.

2 Materials and methods

2.1 Materials

MB dye with the molecular formula of C16H18ClN3S and molar mass of 319.85 \({\text{g mol}}^{-1}\) was procured from “Sigma-Aldrich, India”. A stock solution of 1.0 \({\text{g L}}^{-1}\) of MB dye was prepared in deionized (DI) water and the working solutions were diluted with DI water using the stock solution. The concentration of the dye was determined at 610 nm wavelength using a UV–Vis spectrophotometer (Agilent Cary 100, USA). All chemicals utilized in the experiments were of analytical grade procured from “Sigma-Aldrich, India”.

2.2 Preparation of leaf and stem biochar of Lantana

Lantana was collected from Tathagat Residential Hall (25°00′44.4"N 85°24′37.1"E), Nalanda University, Rajgir, Bihar, India. The plant materials were washed with tap water followed by DI water thrice. It was sun-dried for 10 h and crushed down to smaller pieces followed by washing with DI water thrice. The final powdered form of Lantana biomass was oven-dried (125 L-PID, Icon Instrument Company, India) at 120 ℃ for 24 h. The oven-dried biomass was pyrolyzed at 600 ℃ using a muffle furnace (Icon Instrument Company, India). The chosen temperature is within the temperature range reported for the preparation of lantana camara biochar (Chen et al. 2022; Ganesan et al. 2021). The BC samples prepared after the pyrolysis of Lantana were named BCL600 and BCS600 for leaves and stems, respectively. The samples were ground further using a ball mill to make the particle size uniform and stored in an air-tight glass bottle for batch experiments. Stock suspension of BCL600 and BCS600 were prepared by suspending the required amount of BC in double DI water. The BC suspensions were sonicated for 5 min using an Ultrasonic Cleaner (Olympus 10 × 50 DPS1, India) to completely homogenize the solution; however, the working concentrations were prepared daily with the requisite dilution.

2.3 Characterization of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)

The Infrared spectra of the unloaded \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) was obtained by using Fourier Transform Infrared Spectroscopy (Perkin Elmer-RZX, USA) to identify the surface functional groups in the wavenumber range of 400–4000 \({\text{cm}}^{-1}\). The Zeta potential and particle size (Zeta sizer Nano, ZS90, Malvern Instruments, USA) were determined in order to measure the magnitude of charges on the surface and external dimension of leaf and stem BC. The elemental composition (C, H, O, and N) of the samples was assessed by an elemental analyzer (Vario EL III CHNOS Elemental Analyzer, USA). X-ray diffraction (XRD) analysis was performed to determine the sample’s composition (Bruker, Germany). Scanning Electron Microscopy (SEM) was performed to understand the spatial variations in the chemical composition of BC and Energy Dispersive X-ray analysis (EDAX) was performed to find the presence of metals in addition to C and O in BC (FESEM-EDAX, JEOL, Japan).

2.4 Batch sorption experiments

The adsorptive removal efficiency of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) prepared from Lantana for MB dye removal was examined by batch sorption experiments. Standard dye solutions of desired concentrations were prepared by diluting the stock solution. The pH of the MB dye solution was adjusted using 0.1 M HCl or 0.1 M NaOH solutions. Batch sorption experiments were performed by taking 50 mL of different dye concentrations (5, 10, 15 and 20 \({\text{mg L}}^{-1}\)) in 250 mL of Erlenmeyer conical flasks at desired pH, contact time, and adsorbate concentrations. An absorbent dose of 250 \({\text{mg L}}^{-1}\) and agitation rate of 150 rpm on a digital orbital shaker (WW5190020-Cole Parmer, USA) were used in the batch experiments. The sorption experiments were conducted at 303, 313 and 323 K temperature. The conical flasks were kept on the magnetic stirrer with a hot plate (Spinot-MC-02, India) along with the stirring bars. The temperature of the dye solution was maintained using a thermometer. The filtration and centrifugation techniques are common for separation of absorbent from the solution (Ganguly et al. 2020; Lyu et al. 2018). In this study, the adsorbent was segregated from the aqueous solution after equilibrium by centrifugation at 10,000 rpm for 10 min (Tarsons MC-01 Micro-centrifuge, India). All experiments were conducted in duplicate, and the averages of those results are presented in the manuscript.

3 Results and discussion

3.1 Characterization of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)

The particle size and elemental composition of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) are presented in Table 1. The particle size of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) was 517.2 ± 40.2 and 401.6 ± 4.2 nm, respectively. The elemental analysis of \({\text{BC}}_{\text{L}}600\) demonstrated that the percentage of oxygen is higher than that of carbon. Whereas the percentage of carbon in \({\text{BC}}_{\text{S}}600\) is exceptionally high as compared to the percentage of oxygen, which indicated the carbonaceous nature of the sorbent. The percentage of nitrogen is low in \({\text{BC}}_{\text{S}}600\) as compared to \({\text{BC}}_{\text{L}}600\). The zeta potential values of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) were -31.7 ± 1.6 and -41.3 ± 1.4 mV at about neutral pH conditions. The zeta potential is an indicator of stability of colloidal dispersion. When the zeta potential value is low then the attractive forces overcome the repulsion leading to flocculation. A high value of zeta potential indicates electrically stabilized colloids resisting the flocculation. The stability of the colloids having zeta potential values exiting between 0 to ± 40 mV indicates low stability, while zeta potential values existing between ± 40 to ± 60 mV and above show good stability of the colloids (Greenwood & Kendall 1999). In the present study, the zeta potential value of \({\text{BC}}_{\text{L}}600\) is less than ± 40 mV indicating low stability of BC solution. Whereas \({\text{BC}}_{\text{S}}600\) solution is comparatively stable having zeta potential value slightly greater than ± 40 mV.

Table 1 Particle size and elemental composition of unloaded \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)
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FTIR results played a vital role in elucidating the mechanisms of adsorption and provided valuable insights into the chemical structure of various compounds. The FTIR spectra of unloaded \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) is shown in Fig. 1 (A) and (B), respectively. In \({\text{BC}}_{\text{L}}600\) spectra, the band around 3300–3500 \({\text{cm}}^{-{1}}\) and 1582 cm−1 were attributed due to surface with alcoholic or phenolic \({\text{O}}-{\text{H}}\) stretching and bending mode of vibration. respectively. Due to the presence H-bond, the O–H stretching band is broad enough. Similar O–H stretching and bending modes of vibration were observed at around 3600–3200 cm−1 and 1570 cm−1, respectively in the FTIR spectra of \({\text{BC}}_{\text{S}}600\) (Fig. 1B). From the broadness of IR spectra, the O–H bonds present in \({\text{BC}}_{\text{S}}600\) are more H-bond as compared to \({\text{BC}}_{\text{L}}600\). The bands observed between 1400–1600 \({\text{cm}}^{-{1}}\) confirm the presence of \(\text{C=C}\) bond in both \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) spectra. The C \(-\) H bending vibrations were observed around 876.1, 820.9, 757.5 \({\text{cm}}^{-{1}}\) in \({\text{BC}}_{\text{L}}600\) spectra and 874.5, 796.5, and 712.8 \({\text{cm}}^{-{1}}\) in \({\text{BC}}_{\text{S}}600\) spectra. The C-H stretching vibrations are not observed in both FTIR spectra which is due to the predominant effect of O–H bands. The band around wavelength 1054.5 \({\text{cm}}^{-{1}}\) in the \({\text{BC}}_{\text{S}}600\) spectra was associated with C \(-\) O stretching, which is very weak in case of \({\text{BC}}_{\text{L}}600\).

Fig. 1
figure 1

FTIR spectra of (A) unloaded \({\text{BC}}_{\text{L}}600,\) (B) unloaded \({\text{BC}}_{\text{S}}600\), (C) loaded \({\text{BC}}_{\text{L}}600\) and and (B) loaded \({\text{BC}}_{\text{S}}600\)

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XRD pattern of samples BCL600 and BCS600 are shown in Fig. S1 within the o2θ range of 0–40°. The XRD patterns of both samples are different. In the XRD pattern of \({\text{BC}}_{\text{L}}600\), the numbers of peaks are larger than the XRD pattern of BCS600 within the o2θ range of 0–40°. The peaks showed characteristic amorphous carbon XRD patterns. However, in \({\text{BC}}_{\text{L}}600\), additional peaks are observed due to the presence of crystalline phases of some oxides of different metals like Ca, Mg, Si (i.e., CaO, MgO, SiO2, respectively).

The SEM images elucidate the morphology and physical characteristics of the adsorbent. Figure 2 (A-F) shows the SEM images of BCL600 and BCS600 in two different magnifications. These images reveal the surface morphology of BCL600 and BCS600 samples. It is observed that the sizes of particles present in BCL600 are higher than those in BCS600 and the size distribution in BCL600 is more homogeneous than BCS600. The EDAX analysis explains the elemental composition of adsorbent and stipulate the types and amounts of elements present in the adsorbent, including the main constituents and impurities on the surface. It is useful for understanding the chemical nature of the adsorbent and its potential application in adsorption processes. The EDAX spectrums of both samples are shown in Fig. S2 (A, B). The EDAX spectrum of BCL600 shows that the sample contains not only C and O but also Ca, Mg, and Si. The EDAX spectrum of BCS600 shows that it contains only C and O.

Fig. 2
figure 2

SEM image of (A, B, E) BCL600 and (C, D, F) BCS600 with different resolutions

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3.2 Effect of initial dye concentration and contact time

Figure 3 illustrates the effect of “initial dye concentration” and “contact time” on the adsorption of MB dye onto \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\). The amount of MB dye adsorbed on the surfaces of the adsorbent increased with increase in contact time due to the availability of number of active sites on the surfaces of adsorbents. As shown in Fig. 3, the dye solution treated with \({\text{BC}}_{\text{L}}600\) had attained equilibrium after ~ 100 min of contact time, whereas dye solution treated with \({\text{BC}}_{\text{S}}600\) attained equilibrium after ~ 60 min of contact time at constant pH 8. The adsorption equilibrium capacity had increased from 17.49 to 41.77 \({\text{mg g}}^{-1}\) and 25.95 to 64.93 \({\text{mg g}}^{-1}\) as the initial dye concentration was increased from 5 to 20 \({\text{mg L}}^{-1}\) at 303 K with \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\), respectively. After the diffusion of the dye molecules from the bulk solution to the surface of adsorbent particles, the adsorption process gradually slowed down until it attained equilibrium. This can be explained as a prominence of saturation of functional sites on the surface of adsorbents with the dye molecules and the penetration of dye molecules into the pores of the adsorbent structure. The interaction of dye molecules with the adsorbent was enhanced by increasing the dye concentrations as it overcomes the mass transfer resistance between the solid and aqueous phase (Gautam et al. 2015). Similar trend was reported on the sorption of MB on periodic mesoporous titanium phosphonate materials (Ren et al. 2013).

Fig. 3
figure 3

Effect of initial dye concentration and contact time on adsorption of MB dye with (A) \({\text{BC}}_{\text{L}}600\) and (B) \({\text{BC}}_{\text{S}}600\)(Temperature = 303 K, pH = 8, Volume of MB dye = 40 ml, Contact time = 100 min, BC dose = 10 ml of 250 mg/L)

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3.3 Effect of pH on the adsorption process

The effect of initial solution pH on the adsorption of MB dye by \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) was studied within the pH range of 3–12 and shown in Fig. 4. It was observed that the % removal of dye from aqueous solution increased rapidly with the increasing pH from 3 to 12 by keeping all other experimental conditions constant. The results derived from the experimental data exhibit that the maximum uptake of MB dye occurred for 5 \({\text{mg L}}^{-1}\) of BC at pH 12 with 69.02 and 69.07% using \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\), respectively. The percentage removal of dye was 48.45, 52.13, and 65.72% and 38.38, 49.15, and 54.94% at pH 3, 6 and 9 for \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\), respectively. The sorption mechanism is likely affected as a consequence of the electrostatic attraction and repulsion of dye molecules in the aqueous solution. At low pH, the sorption capacity of both the BC samples are low due the repulsive interaction between positively charged sorbent and positively charged sorbate. MB dye is a cationic dye (pKa = 3.80) and at low pH the BC surface remains predominantly cationic due to sorption of H+ ions. With increase of solution pH, the sorbent surface become less and less positive. As a result, the sorption capacities of cationic MB dye increases. At high pH (~ 12), the surface of sorbent become negative, and a columbic attraction occurs between negatively charged sorbent and positively charged dye molecules (Hameed & Ahmad 2009). Raising the pH generally causes adsorbents to have more hydroxide ions and increases negative zeta potentials which facilitates the adsorption the cationic dye (i.e., MB dye) and removal from the media (Tarekegn et al. 2021). As a result, both the BC samples show very high dye sorption capacity at high pH. Similar results have been reported in adsorption of MB dye on steam-activated carbon prepared from Lantana stem (Amuda et al. 2014).

Fig. 4
figure 4

Effect of initial solution pH of the sorption of MB dye onto (A) BCL600 and (B) \({\text{BC}}_{\text{S}}600\)(Temperature = 303 K, Volume of MB dye = 40 ml, Contact time = 100 min, BC dose = 10 ml of 250 mg/L) at different dye concentration

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3.4 Effect of adsorbent dosage

Figure 5 shows the effect of adsorbent dose on the adsorption of MB dye over the \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) surfaces. The effect of adsorbent dose on the adsorption process was examined by ranging the adsorbent dose from 100 to 400 \({\text{mg L}}^{-1}\) by keeping all other experimental condition constant at 303 K. The maximum amount of dye adsorbed on \({\text{BC}}_{\text{L}}600\) took place for initial dye concentration of 5 \({\text{mg L}}^{-1}\) at pH 8 from 49.54 to 64.78% with an increase in adsorbent dose from 100 to 400 \({\text{mg L}}^{-1}\), respectively. Whereas the adsorption of MB dye on \({\text{BC}}_{\text{S}}600\) was maximum for initial dye concentration of 10 \({\text{mg L}}^{-1}\) at pH 8 with 20.69 to 34.46% for adsorbent dose of 100 to 400 \({\text{mg L}}^{-1}\), respectively. The increase in removal efficiency can be attributed to the increase in available sorption sites (Ahmad et al. 2014). The adsorption of dye onto the surface of adsorbent is highly concentration dependent. The concentrations of MB dye used in the experiment are considerably low in the current study compared to the dosages of the adsorbent. It is possible that there could be more vacant sites available for adsorption until the reaction attains equilibrium. Therefore, the dye concentrations can be increased further until all the vacant sites are occupied by the dye molecules and become saturated. Similar trend was observed for MB dye adsorption with BC prepared from anaerobic digestion residue using palm bark, and eucalyptus (Sun et al. 2013), and steam-activated Lantana camara stem (Amuda et al. 2014). From Fig. 5, it is also observed that \({\text{BC}}_{\text{L}}600\) is an effective adsorbent at low dye concentration, whereas \({\text{BC}}_{\text{S}}600\) is effective at comparatively high dye concentration.

Fig. 5
figure 5

Effect of adsorbent dose on the adsorption of MB dye with (A) \({\text{BC}}_{\text{L}}600\) and (B) \({\text{BC}}_{\text{S}}600\) (Temperature = 303 K, pH = 8 Volume of MB dye = 40 ml, Contact time = 100 min)

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3.5 Adsorption kinetics

The sorption kinetic data were analyzed by nonlinear pseudo 1st order and pseudo 2nd order kinetic model equations (Eq. S3 and S5) and shown in Fig. 6. The kinetic parameters obtained are summarized in Table 2. From R2 or χ2values, it is clear that the sorption kinetics follows pseudo 2nd order kinetic model for both \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\). From the rate constant values (k2), it is shown that with increase of initial dye concentration rate of sorption reaction gradually decreases for both the adsorbent. This is due to the fact that with increase of initial dye concentration, the number of active sites on the sorbent per adsorbate molecules has decreased. Rate constant values are almost within the similar range for both the sorbents. Similar studies were reported for MB dye sorption kinetics on different adsorbents follows the pseudo 2nd order kinetic model (Kannan & Sundaram 2001; Rafatullah et al. 2010; Tran et al. 2020). Since present sorption reaction follows the pseudo 2nd order model for both the sorbents therefore the sorption process must be multistep process in which some diffusion steps must be involved.

Fig. 6
figure 6

Kinetic model fitting of adsorption of MB dye with (A) \({\text{BC}}_{\text{L}}600\) and (B) \({\text{BC}}_{\text{S}}600\)(Temperature = 303 K, pH = 8 at different concentration

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Table 2 Parameters of two kinetic models for MB dye adsorption onto \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)
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3.6 Diffusion kinetics

The present kinetic data were analyzed by intraparticle diffusion model (Eq S6) and shown in Fig. S3. The results obtained are summarized in Table 3. From the Fig. S3 A and B, it is clear that for both \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) sorbents the qt vs t0.50 plot shows linearity from 5 min of contact time at all the concentrations studied. None of the linear plot passes through the origin. Thus, for the present sorption process, the intraparticle diffusion mechanism is operative but it is not involved in the rate limiting step (Suteu et al. 2015).

Table 3 Intraparticle diffusion model parameters of MB dye sorption on \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)
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3.7 Adsorption isotherms

The adsorption equilibrium data at three different temperatures of MB dye for \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) were analyzed using non-linear Langmuir and Freundlich isotherm models equations (Eqs S6 and S8). The fitted curves are shown in Fig. 7. The isotherm parameters obtained by this analysis are summarized in Table 4. Comparison of R2 or χ2 values clearly indicates that sorption of MB dye on both \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) follows Langmuir sorption isotherm at all the temperature studied. It indicates a homogeneous distribution of active sites on the surface of BC resulting in the formation of monolayer coverage of MB dye at a given temperature. Maximum sorption capacity (qm) of \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) are 64.99 and 15.39 mg g−1, respectively at 303 K. Thus, it is clear from qm values that \({\text{BC}}_{\text{L}}600\) shows a fairly high sorption capacity than \({\text{BC}}_{\text{S}}600.\) From Table 4 it is seen that maximum sorption capacity gradually decreases with increase of temperature for both the sorbents. This indicates the exothermic nature of the sorption process. The separation factor (\({\text{R}}_{\text{L}})\) values are calculated (Eq S7) for MB dye adsorption on both the sorbents at all the temperatures studied and presented in Table 4. The RL values for all the sorption processes lie within the range > 0 and < 1 which indicates favorable adsorption conditions (Bulut & Aydın, 2006).

Fig. 7
figure 7

Isotherm study of MB dye on (A) \({\text{BC}}_{\text{L}}600\) and (B) \({\text{BC}}_{\text{S}}600\)( pH = 8) and Freundlich, Langmuir sorption isotherm fitting

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Table 4 Parameters of adsorption isotherms of MB dye onto \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)
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A comparative study of maximum Langmuir sorption capacities of different adsorbents applied for MB dye sorption with the adsorbents had been performed and presented in Table 5. Comparing the sorption capacity of different adsorbents with the \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\) further determines the potential of using \({\text{BC}}_{\text{L}}600\) to produce BC for the effective removal of MB. Adsorbents such as activated carbon developed from Ficus carica bast (Pathania et al. 2017), Polyaniline and Polypyrrole Macro– nanoparticles (Maruthapandi et al. 2018) and Dunino raw halloysite mineral (Filice et al. 2021) have reported low qm values than \({\text{BC}}_{\text{L}}600\) for the effective removal of MB dye.

Table 5 Comparison of maximum Langmuir sorption capacity (qm) of different adsorbents
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3.8 Thermodynamics of the adsorption process

The spontaneity of the adsorption process was determined by thermodynamic studies. Gibb's free energy change (ΔG0) indicates the spontaneity of any process. At a given temperature, the adsorption processes occur spontaneously if ΔG0 is negative. It is required to take into account both the enthalpy (ΔH0) and the entropy (ΔS0) while computing ΔG0. An exothermic adsorption is shown by a negative value for ΔH0, whereas an endothermic adsorption is indicated by a positive value. Different thermodynamic parameters are calculated using Eqs S9 and S10. The plot of lnK vs 1/T is shown in Fig. S4. Thermodynamic parameters of MB dye sorption on BCL600 and BCS600 are presented in Table 6. In the present study, the negative values of \({\Delta {\text{G}}}^{0}\mathrm{ at all the three}\) temperatures studied shows that, the dye adsorption onto BCL600 and BCS600 are favorable under the experimental conditions (Sumalinog et al. 2018). Decrease of \({\Delta {\text{G}}}^{0}\) values with increase in temperature for both sorbents indicates that the sorption is more favorable at higher temperature conditions. The negative values of \({\Delta {\text{H}}}^{0}\) confirmed the exothermic nature of the sorption process. The estimated values for change in enthalpy are lower than 40 kJ mol−1for both sorbents. This clearly indicates that the forces involved between the adsorbate and absorbent is van der Waals’ type resulting in physical adsorption (Fan et al. 2016). Thus, it can be inferred that the interaction mechanism between the MB dye and BC is mainly physical adsorption in the current study. The positive \({\Delta {\text{S}}}^{0}\) values suggest an increase in the degree of randomness at the solid/solution interface, which is comparatively higher in case of dye adsorption onto BCS600 than BCL600 during the sorption process. This observation is in accordance with MB dye adsorption onto activated BC produced from municipal solid waste (Sumalinog et al. 2018).

Table 6 Thermodynamic parameters of sorption of MB dye onto \({\text{BC}}_{\text{L}}600\) and \({\text{BC}}_{\text{S}}600\)
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3.9 Mechanisms of interaction

The possible mechanism of dye adsorption on BC prepared from Lantana camara L. can be determined by combining the results of adsorption and infrared spectroscopy analyses. The interaction of hydroxyl-group presents on the surface of BCL600 and BCS600 with MB dye molecules in the solution is related to intermolecular hydrogen bonding. In alkaline conditions, the hydroxyl-group increases in the solution by enhancing the interaction of dye molecules with the BC. The presence of aromatic carbon on both BCL600 and BCS600 indicated the hydrophobic nature of the adsorbents. The presence of C = C and C \(-\) H groups indicated the presence of benzene like structure on the surface of BCL600 and BCS600. The structure of MB dye contains three aromatic rings which allow the π-π interaction with both BCL600 and BCs600 since their surfaces also contains benzene like structures. The cation of the thiazine group is present in the middle of two aromatic rings comprising of dimethyl amide in the MB dye structure, which possibly have interacted with the π system of the BC. The presence of C \(-\) O on the surface of BCS600 can interact by developing an electrostatic bond with the cationic dye molecules. The N atoms of two 2° amine groups present in the structure of MB dye can form H-bonds with the surface O–H groups of BC samples. As a result of the fact the intensities of O–H stretching mode of vibrations are lowered to a large extent (Fig. 1 C, D). Similar interactions have been reported previously for MB dye interaction with activated carbon, graphene oxide, and carbon nanotubes (Li et al. 2013). It is likely that the dye molecules can physically interact with the surface moieties of both the adsorbents prepared from Lantana in the current study. A schematic representation of different interactions of MB dye with BC samples are presented in Fig. 8. It highlights the hydrogen bonding, electrostatic interaction, and stacking interaction, respectively over the BC surfaces.

Fig. 8
figure 8

Different types of interactions between methylene blue and biochar surface

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4 Conclusions

Through the adsorption potential of BC prepared from Lantana biomass, the adsorption of dye onto BCL600 and BCS600 was found to be highly dependent on the pH of the solution. It was observed that the adsorption of dye molecules onto BCL600 and BCS600 had significantly increased with the increase in pH from 3 to 12. The amount of dye adsorbed by BCL600 was found to be high as compared to that of BCS600 possibly due to the presence of surface –O–H and C–O functional groups. It was also noted that the dye analyzed with BCL600 acquired equilibrium in ~ 100 min only whereas, the dye treated with BCS600 attained equilibrium in about 60 min. The rate of sorption for both BCL600 and BCS600 was following the “pseudo-second-order” kinetics. It implies that the adsorption of dye molecules onto the surface of BC is a multi-step process involving external adsorption as well as diffusion of the dye molecules into the interior of the adsorbent. The equilibrium data fitted very well in the Langmuir isotherm equation for both the adsorbents, confirming multilayer adsorption capacity with BCL600 and BCS600. The negative values of both ΔG0 and ΔH0 for the sorption processes suggest that the adsorption processes are spontaneous and exothermic in nature. The positive values of ΔS0 signified a higher degree of randomness in the system comprising BC and dye molecules.

Thus, this study goes a step further from previous works carried out by Amuda et al. (2014), as both the biomass of this invasive species has been studied and dye adsorption properties of the biochar derived from stem was also ascertained. BCL600 prepared from leaves had higher dye absorption capacity although it reached equilibrium at ~ 100 min where as BCS600 reached it at about 60 min. This implies effective utilization of the entire biomass of an invasive species. The surface of BC can be further modified depending on the types of contaminants present in their surroundings. Therefore, it can be anticipated that the BC particles obtained from Lantana camara L. biomass can be applied as a potential adsorbent for water and wastewater treatment.