Removal and Adsorption Mechanism of Tetracycline Using Manganese-modied Cotton Straw Biochar in an Aqueous Solution

Farming in China’s rural areas leads to antibiotic pollution in waterbodies making it a grave issue. Cotton straw biochar (CSBC) was prepared by oxygen-limited pyrolysis at 400 °C (CSBC400) and 600 °C (CSBC600); and Mn-modied CSBC (MCSBC) was produced by the KMnO 4 wrapping method for tetracycline (TC) removal from aqueous solutions. The effects of temperature, initial solution concentration, pH, ion type, and ionic strength on TC adsorption were investigated. The adsorption process of the biochars achieved an equilibrium state after 360 min, and the highest equilibrium adsorption amount (13.254 mg/g) was found for MCSBC. The kinetic adsorption process, which was dominated by chemisorption, was well-described by the pseudo-second-order kinetic model. The adsorption was a non-homogeneous heat absorption process, and the adsorption isotherm data tting was compatible with the Freundlich model. A better adsorption effect of MCSBC was observed when the pH was < 4. Monovalent cations (Na + , K + , NH 4+ , and Ca 2+ ) had a facilitative effect on the adsorption process. The adsorption mechanisms of TC by MCSBC included pore diffusion, H bonding, electrostatic interactions, and π–π accumulation. Therefore, MCSBC has a good adsorption capacity for TC and can be used for the treatment of TC-based pollutants in aqueous environments.


Introduction
Antibiotics are chemical substances produced by microorganisms or higher plants and animals with antipathogen or other active cell development functions (Grenni et al. 2018). They not only inhibit the growth of bacteria in animals, but also promote animal feed utilization and growth rates. With the rapid development of large-scale breeding technologies, use of antibiotics have increased in the breeding industry. However, using antibiotics can have adverse impacts on humans and introduces new challenges to the treatment of emerging environmental pollutants. A Study by Wei et al. (2016) reported that only a small fraction of antibiotics can be absorbed after being used by humans and animals; the majority is discharged into water or soil via human and animal excrement, and may enter the human body via the food chain. Zhao et al. (2018) reported that many unabsorbed veterinary antibiotics are introduced by animal feces and retained in the farmland soil environment for a long time. Moreover, the order of magnitude of the concentration of antibiotics detected in soil has increased from ng/kg to mg/kg (Li et al. 2015). Therefore, attention should be paid to antibiotic pollution in the environment. Tetracycline (TC) is the second most common antibiotic produced and used worldwide; its production and consumption in China is one of the highest worldwide (Grenni et al. 2018). Similar to other antibiotics, TC is extremely harmful to human and animal health because of its low metabolic e ciency. Approximately 90 % of the ingested TC can be discharged into the environment via urine; TC ingested by pigs can be rapidly eliminated via urine and feces (Daughton and Ternes, 1999). A study analyzed the in uence of poultry activities consuming TC; it concluded that the concentration of TC in the environment owing to animal activities was 30-497 ng/L (Winckler and Grafe, 2001). One study reported that the maximum total concentrations of three types of TC (i.e., TC, aureomycin, and oxytetracycline) in vegetable soils with manure application were 242.6 µg/kg and 415.0 µg/kg in southern and central China, respectively (Wei et Loading [MathJax]/jax/output/CommonHTML/jax.js al. 2019). Therefore, it is important to explore effective TC removal technologies to control antibiotic pollution in soil and water environments.
Traditional methods for removing antibiotics from water include adsorption and xation, biodegradation, membrane separation, coagulation, and advanced oxidation (Anthony et al. 2021). The adsorption and xation method is widely used for the removal of TC from water because of its advantages, including simple operation, low price and adsorbent concentration, good removal performance, and non-toxic byproducts. Among all the adsorbents, biochar (BC) a new material, is highly e cient and has been widely studied in recent years. It is a highly aromatic and refractory solid substance formed by the pyrolysis and carbonization of biomass materials in the absence of oxygen (Lehmann et al. 2006). BC application has advantageous effects in water environment remediation, C xation, and waste reuse. BC is expected to become an appropriate alternative material for activated C adsorption. BC prepared with wheat straw, sawdust, plant residues, municipal sludge, and algae as precursor materials has been studied for the removal of TC in water environments (Hao et  In this study, cotton straw (CS) BC (CSBC) was prepared by lower limit oxygen pyrolysis at 400°C (CSBC400) and 600°C (CSBC600) using CS as the precursor material. Mn-modi ed BC was prepared by the KMnO 4 coating method with CSBC600 as the precursor material, and its related structures were characterized. The kinetic adsorption and isothermal adsorption processes of TC by the three BCs were analyzed, and the effects of pH, ionic strength, and ambient temperature on TC removal were investigated. The adsorption of TC in wastewater by the three BCs was investigated using a kinetic adsorption model and thermodynamic adsorption models. The functional groups on the surface of the Mn-modi ed CSBC (MCSBC) before and after adsorption were characterized by Fourier transform infrared spectroscopy (FTIR) and X-ray diffraction (XRD), and the adsorption mechanism was speculated. This study may provide new ideas for the development of BC materials and further promote the application of BC materials in water treatment.

CSBC and MCSBC preparation
CS was collected from a cotton eld at the experimental station of Tarim University in Alar City, China (81°16'53"N, 40°36'28"E). The CS biomass materials were collected, washed with distilled water, dried at 60 °C, and pulverized using a small high-speed and multi-function pulverizer. The material was then loaded into a preheated quartz-tube furnace (KPX-9162 MBE, China) and incubated at 20 °C/min to 400 or 600 °C for 4 h. After air-cooling to 25 ℃, the products were washed using deionized water and subsequently oven-dried at 80 °C for 12 h. The ne pulverized products were further separated by passing them through a 60 mesh sieve. The CSBC products (CSBC400 and CSBC600) were packed in a sealing bag and stored for further use. The adsorption speed was set at 180 rpm and the centrifuge tube was oscillated in the dark at constant temperature for 6 h (the preliminary experiment showed that the adsorption reached equilibrium after 6 h). After reaching the equilibrium time, the samples were removed, left for 2 h, and centrifuged at 4000 rpm for 20 min. The supernatant was then ltered through a 0.45 μm membrane.

Adsorbent characterization
The C, N, H, and O contents of the BC were determined using an elemental analyzer (Elementar Micro Cube, Germany). The speci c surface area, pore volume, and pore size of the BC were determined using a Brunauer-Emmett-Teller (BET) speci c surface area analyzer (ASAP 2460, McInstruments, USA). The mineral composition of the functional groups was determined using FTIR (Perkin Elmer Spectrum 100, USA). The mineral composition of the BC was determined using XRD (Bruker D8 Advance, Germany).

Data analysis
The TC adsorption capacity (q e ) and removal rate (R) were calculated as follows: where q e (mg/g) is the equilibrium adsorption capacity, C 0 (mg/L) is the initial concentration of TC in the solution, and C e (mg/L) is the equilibrium concentration of TC in the solution. V (L) is the volume of the solution, m (g) is the amount of BC added, and R (%) is the removal rate.
The pseudo-rst-order kinetics model, pseudo-second-order kinetics model, and intra-particle diffusion model were used to t the adsorption kinetics process of BC for TC, as follows: where q e (mg/g) is the equilibrium adsorption capacity, q t (mg/g) is the adsorption capacity (t/min), k 1 (min − 1 ) is the pseudo-rst-order kinetic rate constant, k 2 (g mg − 1 min − 1 ) is the pseudo-second-order kinetic rate constant, k d (mg g − 1 min − 0.5 ) is the rate constant of the diffusion model within the particles, and C (mg/g) represents the marginal layer effect.
Langmuir and Freundlich isothermal adsorption equations were used to t the adsorption of TC by BC, as follows: q e = q max k L C e 1+k L C e 6 ( ) where q e (mg/g) is the equilibrium adsorption capacity, q max (mg/g) are constants and the maximum adsorption capacities related to the adsorption capacity and adsorption strength in the Langmuir model, respectively. k F (L/mg) and k L (L/mg) are the Freundlich constants, and the size of 1/n is related to the nonlinearity of the adsorption isotherm.

BC characterization
To accurately analyze the micro-morphology of the material and the distribution of MnO x , the original BC and modi ed BC were observed by scanning electron microscopy-energy dispersive X-ray spectrometry (SEM-EDS) (Figure 1a, 1b, and 1c). The surface of the CSBC400 was smoother than CSBC600 and MCSBC, with uneven pleats, no signi cant pores, many particles deposited on the surface, and a nonporous structure, thereby indicating that the biomass was not completely cleaved and the pores were not completely opened. The surface of CSBC600 was rough, with irregular pleats, large protrusions, and some pores due to the escape of small molecules during pyrolysis ( could have caused an increase in the pore diameter, thereby increasing the speci c surface area and pore volume. Furthermore, an increase in the pore volume contributed to the entry of TC into the internal pore system (Shen et al. 2020).

Adsorption effect of BC
The adsorption effect of BC prepared under different conditions was evaluated so as to optimize the adsorption material for water treatment. The adsorption performance of CSBC600 at high pyrolysis temperatures was signi cantly better than that of CSBC400 at lower pyrolysis temperatures ( Figure 2). Thus, CSBC600 was selected as the modi ed precursor material in subsequent experiments, and the adsorption mechanism was explored. The best adsorption performance was achieved with  weakened the driving force of adsorption, thereby resulting in a slow increase in the adsorption amount ). The adsorption of TC by BC gradually stabilized after 360 min.
To better explain the adsorption mechanism, pseudo-rst-order and pseudo-second-order adsorption models were used to t the kinetic experimental data. The tting results of the pseudo-rst-order and pseudo-second-order adsorption models are listed in Table 2. The pseudo-second-order adsorption model accurately described the TC adsorption behavior of the original BC and modi ed BC (R 2 > 0.93). The TC adsorption amount obtained by tting the pseudo-second-order kinetics equation was closer to the experimentally measured value, thereby indicating that chemisorption was predominant throughout the adsorption process (Shepherd et al. 2017). The adsorption mechanism and rate-limiting steps of TC on all the BCs were further examined using the intra-particle diffusion model (Jiang et al. 2017). Figure 4 shows the tting curve of intra-particle diffusion is composed of two linear segments: the rst stage is straight but not at the origin, thereby indicating that TC adsorption is a multi-step process and that intraparticle diffusion is not the only rate-limiting step (Xiong et al. 2018). In different diffusion stages, the linear relationship between the adsorption amount of TC and t 0. . The diffusion rate of MCSBC was approximately six times that of CSBC600 and CSBC400, thereby indicating that the diffusion rate in the particles decreased sharply with the complete adsorption and lling of the pores in MCSBC; however, TC molecules continued to migrate slowly on the surface until they entered the pores of the particles due to activation (Song et al. 2014). The intercept C represents the range of the boundary layer thickness, that is, the larger the intercept, the greater the boundary layer effect (Limousin et al. 2007). As shown in Table 3, the value of C 2 was greater than that of C 1 , and the diffusion boundary effect was more signi cant in the second layer, thereby indicating that the total adsorption rate was controlled by both liquid lm diffusion and intra-particle diffusion (Limousin indicated that the TC adsorption force of MCSBC was stronger than that of CSBC400 and CSBC600. One study (Liu et al. 2019) showed that the Langmuir model could better describe the TC adsorption process of BC, and the adsorption was mainly single-layer adsorption. However, another study (Jang et al. 2018) showed that the Freundlich model could better describe the adsorption of TC by loblolly pine BC, and the adsorption process was heterogeneous. The equilibrium adsorption capacity of the Mn-modi ed BC prepared in this study was 42.076 mg/g, which was higher than that of CSBC400 (26.936 mg/g) and CSBC600 (28.038 mg/g). Therefore, MCSBC can potentially be applied as a water treatment material for the adsorption and removal of TC.

Effect of temperature on TC adsorption
As shown in Figure 6, as the temperature increased from 25 to 45 ℃, the TC adsorption capacity of the three BC materials gradually increased. The TC adsorption capacity of BC increased as the temperature and speed of molecular movement increased (Shen et al. 2020 The effect of solution pH on the removal of TC by different BCs is shown in Figure 7. As the pH of the solution increased from 3 to 9, the TC adsorption capacity of CSBC600 and MCSBC decreased gradually, whereas the adsorption capacity of CSBC400 increased gradually, thereby indicating that the TC adsorption capacity of BC depends on both the pH and BC properties. Under different pH values, TC showed different species distributions. As the pH of the solution increased from 3 to 9, the TC adsorption capacity of CSBC600 and MCSBC decreased gradually because of the strong electrostatic repulsion between the TC molecules (H 4 TC + and H 2 TC -) and the positive and negative charges on the surface of the adsorbent (Zhu et al. 2014). When the pH of the solution was lower than 3.3, the main type of TC was protonated TC (TC + ) and the BC surface was negatively charged. Therefore, the interaction between TC and BC is mainly electrostatic, and the adsorption effect is the best (Zhang et al. 2019a). When the pH was between 3.4 and 7.7, TC molecules (i.e., H 3 TC) had no net charges, the electrostatic attraction was weak, and the adsorption e ciency decreased (Liu et al. 2021). In contrast, the TC adsorption capacity of CSBC400 showed an upward trend, which also veri ed that pH can affect the adsorption capacity of BC, thereby indicating that adsorption mechanisms such as surface complexation and cation bridging might have existed . The study showed that ash affected the speci c surface area and porosity of BC (Li et al. 2017), and thus affected its adsorption effect. The difference in ash content is a key factor affecting pH (Li et al. 2017). The ash content of the CSBC differed and was the key factor affecting the pH and zeta potential. Therefore, the difference in the adsorption effect at different pH values might have been caused by ash ). In conclusion, changes in pH affect the surface charge and TC type distribution on BC, which further shows that the electrostatic interaction between BC and TC is not the only in uencing factor and mechanism.
In uence of ionic strength on adsorption capacity The effects of salt ions (Na + , K + , Ca 2+ , and NH 4 + ) and the TC removal intensity using the various BCs are shown in Figure 8. As the concentration of monovalent cations in the solution increased, BC had a slight promoting effect on TC removal, which might have occurred because the low Na + concentration can improve the ionization level of TC molecules (Xiang et al. 2020), thereby resulting in a strong electrostatic interaction between the charged TC and the surface of the BC adsorbent. Related studies (Ersan et al. 2017; Gao et al. 2012) have reported that Na + and K + have a slight promoting effect on TC adsorption by BC; moreover, these studies speculate that the mechanism was attributed to salting-out. The studies also indicated that the non-electrostatic force could offset part of the electrostatic repulsion force, thereby increasing the adsorption capacity. Compared with the above salt ions, the bivalent cation Ca 2+ also had a promoting effect on TC adsorption, which might have been due to the electrostatic interaction between

XRD comparison before and after adsorption
The crystal structure and phase composition characteristics of the BC samples were characterized using XRD, as shown in Figure 10. In the CSBC400, CSBC600, and MCSBC structures, the diffraction peaks at 28.0°, 41.0°, and 25.9° could be assigned to SiO 2 , which has also been reported in previous studies (Lin et al. 2017 indicating that there was a chemical effect between CSBC600 and KMnO 4 in the modi cation process.
The sharp diffraction peak of Mn 3 O 4 in the MCSBC structure was signi cantly weakened after adsorption, which proved that TC adsorption had an effect on the phase of the MCSBC material structure.
However, MCSBC still had a diffraction peak similar to that of the original CSBC600 after adsorption, thereby indicating that MCSBC maintained a high degree of crystallinity after adsorption.

Adsorption mechanism
Loading [MathJax]/jax/output/CommonHTML/jax.js BC adsorption mechanisms include pore diffusion, Lewis acid-base action, electrostatic action, ion exchange, surface complexation, a cationic-π mechanism, H bonding, and π-π stacking (Hoslett et al. ). In addition, the complete aromatic structure of high-temperature BC can form π-π stacking with the benzene ring structure of TC. In this study, the speci c surface area of the CSBC was in the order of MCSBC > CSBC600 > CSBC400. The TC adsorption capacity of CSBC was in the order of MCSBC > CSBC600 > CSBC400. Therefore, pore action is an important mechanism for TC removal by BC.
The effect of solution pH on the adsorption of TC by BC indicates that the effect of electrostatic action is limited because pH changes the surface charge of BC and the presence of TC, thereby affecting the interaction between BC and TC (Liu et al. 2021).
FTIR analysis showed that -OH, C-O-C, the aromatic structure, and -CH 2 were involved in TC adsorption, and the mechanisms involved included H bonding and π-π stacking. The reaction mechanism can be inferred from the changes in the XRD binding energies and relative contents of the functional groups. The O-containing functional groups of MCSBC were different from those of CSBC600 and CSBC400. XRD analysis showed that the changes in the O-containing functional groups in MCSBC after TC adsorption were related to π-π stacking and H bonding ). FTIR analysis and XRD showed that the aromatic structure and O-containing functional groups of BC played an important role in TC adsorption.
In conclusion, the main mechanisms of TC adsorption by MCSBC are pore diffusion, H bonding, electrostatic interactions, and π-π stacking.

Conclusion
In this study, a new type of MCSBC granular adsorbent with a porous and coarse structure was synthesized using the KMnO 4 coating method. The adsorption of TC by MCSBC was effective in the low pH range and was less affected by common coexisting cations (Na + , K + , Ca 2+ , and NH 4 + ) in aqueous solution. The kinetic results indicated that the TC adsorption process was chemisorption and that 360 min was an adequate time to reach equilibrium. The equilibrium data showed that the process represents Loading [MathJax]/jax/output/CommonHTML/jax.js a mixture of monolayer and multilayer adsorption, and an increase in temperature is bene cial to the process. Although the MCSBC presented in this study showed its potential as a valid adsorbent for TC from aqueous solutions, additional factors (e.g., microorganisms, coexisting anions, and methods) should be considered before its application in future works.

Declarations
Ethics approval and consent to participate Not applicable.

Consent for publication
All authors allow the publication of the paper.

Availability of data and materials
All data used to support the ndings of this study are included with in the article.

Competing interests
The authors declare that they have no competing interests. Notes: q e is the amount of tetracycline adsorption; K 1 is the pseudo-rst-order rate constant; K 2 is the pseudo-second-order rate constant. Table 3. Fitting parameters of intra-particle diffusion equation of cotton straw biochars (CSBCs) pyrolyzed at 400 ℃ (CSBC400) and 600 ℃ (CSBC600) and cotton straw biochar modi ed with 0.035 M KMnO 4 (MCSBC).
Adsorbent K d1 ( mg g -1 min -0.5 ) C 1 (mg/g) R 2 K d2 ( mg g -1 min -0.5 ) C 2 (mg/g) R 2 Notes: K d1 and K d2 are the intra-particle diffusion rate constant; C 1 and C 2 are a constant representing the thickness of boundary layer.

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