Radioactive contamination has drawn great attention with the development of nuclear science and technology. Especially for long-lived actinides and lanthanides, even little radioactivity can pose long-term radiation for water and hazard human health^[1,2,3]. Thus, it is vital to remove or immobilize the radionuclide from the radioactive wastewater as high-efficient as possible. Europium (Eu), as a homologue for the actinides and lanthanides due to their similar physicochemical properties and similar ionic radius, has been widely studied^[1,4-8]. The research on Eu(III) is helpful for us to understand other trivalent actinides and lanthanides in the environment. Currently, many technologies were applied to remove Eu(III), such as precipitation^[9,10], ion exchange^[11], adsorption^{[1,2-4,12-13]}, coagulation-flocculation^[14], and membrane filtration^[15]. Among these approaches, sorption is a cost-efficient and widely-used method.

Generally, the sorption behavior of a radionuclide on the solid-water interface is a crucial process mediating its diffusion, migration, mobility and bio-availability in the natural environment^[16,17]. For Eu(III), its sorption behavior on various absorbents such as compacted bentonite^[18], alumina^[19], carbon nanotubes^[20], titanate nanotubes^[4], manganese dioxide^[6], attapulgite^[21], γ-Al₂O₃^[16], grapheme oxide^[1,8,22] has been widely studied. However, biochar, as a low-cost and highly efficient adsorbent, is a relatively new practice. Furthermore, biochar has rarely been used to remove Eu(III). For example, Frišták et al.^[23] separated the Eu sorption on Fe-modified biochar in their recent work. Therefore, it is necessary to enhance its sorption research on biochar.

Biochar-based materials with excellent physicochemical properties and diversified functionalities are widely used in wastewater treatment fields^[24]. As one of diverse feedstocks, agricultural residues are generally employed to make biochar by thermal degradation in an oxygen- limited environment^[25]. Especially, herbaceous biochars usually have higher carbon contents, larger surface areas and strong aromaticity, high porosity, and abundant oxygen functional groups^[24], which benefits the removal of heavy metals. In recent decades, many researchers have proved sorption mechanisms of heavy metals on biochar, including physical sorption, ion exchange, electrostatic interactions, complexation, precipitation, which endows biochar with high-efficient sorption capability^[25]. For example, 99% Cd, Pb, and Zn could be removed from aqueous solutions by using biochar^[26]. Commercial biochar shows excellent sorption performances towards several heavy metal ions, such as Zn(II), Cu(II), Co(II), Pb(II), and Cd(II)^[27]. Currently, biochar is more widely researched on the sorption for some heavy metals, especially for cadmium ions, lead and chromium ions^[24]. For Eu(III), it is relatively scanty. Thus, biochar technology, as a promising and emerging waste water treatment technology, should be applied to explore its sorption capability for Eu(III) from water.

All properties of adsorbates, adsorbents, and aqueous solutions determine the sorption behavior of heavy metals on adsorbents. Effects of many environmental factors such as pH, ionic strength, humic substances, and temperature on Eu(III) sorption onto various adsorbents have been proved to be evident^{[1-5,16-19,26]}. It indicates that environmental factors of solid-liquid interface mediate the Eu(III) sorption process. For example, pH affects Eu(III) removal ratio through changing the speciation, surface charge and binding sites of the sorbents^[1,19]. Humic substances enhance or inhibit the Eu(III) sorption through different degrees of complexation with adsorbates and adsorbents^[5,19]. Higher temperature provides sufficient energies for the transfer of Eu(III) to the interior of the absorbent^[4,19]. However, the Eu(III) sorption behavior on biochar as a function of various environmental factors is still scarcely studied. Therefore, it is necessary to explore the relative mechanism to make sure the sorption capacitiy of biochar for Eu(III).

In present work, we selected the agricultural residual (rice straw) slowly pyrolyzed at 600 ℃ as the adsorbent to investigate the effects of contact time, pH, ionic strength, humic substances, and temperature on the sorption behavior of Eu(III). To illustrate the characteristics of the adsorbent, biochar samples derived from rice straw were detected using SEM, TEM, and FT-IR, respectively. To demonstrate the interaction mechanisms between Eu(III) and the biochar, three kinetic models and four isotherm models were built.

The aboveground biomass of rice straw was harvested from Shaoxing city (Zhejiang province, China). Specifically, the rice straw was firstly obtained in the field, washed with distilled water and then dried at 105 ℃ in the lab. The dried biomass was crushed and ground to <1.0 mm particle size using a fodder grinder and then was sieved by less than 250 μm (60 mesh) sieve. Subsequently, the sieved particles were pyrolyzed at 600 ℃ with a supply of N₂ at a rate of 7 ℃/min. And to keep complete carbonization, the pyrolysis time was set at 4 h.

Eu(III) stock solution was prepared by dissolving, evaporating, and re-dissolving Eu₂O₃ in 1.0 mmol/L HClO₄. Humic substances,i.e., humic acid (HA) and fluvic acid (FA), extracted from the soil in Hua-Jia county (Gansu province, China), were adequately studied previously and proved to have strong effects on interactions between adsorbate and adsorbent^[3,28]. All chemicals employed in this experiment were purchased in analytical purity and used without any purification. All solutions were prepared with Milli-Q water.

A field emission scanning electron microscope (JSM- 6360LV, Japan) and a transmission electron microscope (JEM-1011, Japan) instrument were used to observe the microstruture of rice straw-derived biochar samples, respectively. The FT-IR spectra of the biochar was recorded with a FT-IR spectrometer (NEXUS, America) in the wavelength range of 4000-400 cm^-1 to characterize its surface functional groups. In addition, Brunauer-Emmett- Teller (BET) and Barrett-Joyner-Halenda (BJH) models were used to evaluate specific surface area and pore volume of the rice straw-derived biochar samples, respectively.

The batch experiments of Eu(III) on biochar were carried out at three temperatures, i.e., 298, 318, and 338 K. Specifically, the stock suspensions of Eu³⁺ solution, biochar, NaClO₄, HA, FA, and Milli-Q water were added in polyethylene tubes in order to get the desired concentrations of different components. The ratio of Eu(III) to the solution volume was 0.45 g/L. The pH of sorption systems were adjusted through adding inappreciable volumes of 0.01 or 0.1 mol/L HClO₄ or NaOH solutions. The suspensions were stirred for 24 h to reach the sorption equilibrium, and then the solid and liquid phases were separated by centrifugation at 9000 r/min for 30 min.

The Eu(III) concentration in the supernatant was measured by liquid scintillation counting using a Packard 3100TR/AB Liquid Scintillation analyzer (PerkinElmer) with ULTIMA GOLD ABTM scintillation cocktail. The Eu(III) removal percent [Eu(III) sorption=(C₀-C_e)/C₀× 100%], distribution coefficient [K_d=(C₀-C_e)/C_e×V/m], and sorption amount onto biochar [q_e=(C₀-C_e)×V/m] were calculated from the initial Eu(III) concentration (C₀, mg/L), the final or equilibrium Eu(III) concentration (C_e, mg/L), the biochar mass (m, g), and the suspension volume (V, L).

All experimental data were subjected to the averages of duplicate or triplicate experiments to ensure the experimental repeatability and improve the data accuracy. The relative errors of the experimental data were less than 5%.

Fig. 1 shows SEM image, TEM image, and FT-IR spectrum of rice straw-derived biochar. SEM image suggests the stratified structure is formed and there are large gaps between two stratums (Fig. 1(a)). Furthermore, both surface and edge of each stratum are relatively rough. TEM image shows that the biochar sample has two kinds of inner structures: one is loose and rough, the other is sheet and smooth, both of which may endow the biochar with excellent sportive capability (Fig. 1(b)). FT-IR spectrum indicates functional groups on biochar (Fig. 1(c)). The band at 3444 cm^-1 is ascribed to both free and H-bonded -OH stretching vibrations of biochar^[29,30]. The peaks at 1385 and 1037 cm^-1 are attributed to aromatic carboxyl O=C-O and alkoxy C-O stretching vibrations, respectively^[1]. The bands at 795, 521, and 467 cm^-1 are attributed to the integrated crystalline structure of biochar^[31]. These oxygen functional groups may effectively bind with heavy metals. According to the results shown in our previous study^[32], the average pore diameter is 45.43 nm, and the specific surface area of the biochar is 61.11 m²/g.

The initial solution pH importantly affects the sorption behavior of heavy metals on biochar^[1,19]. The influence of pH on the Eu(III) sorption on biochar in the range of 1.9-10.2 is shown in Fig. 2(a). The sorption percentage increases from 9% to 97% with the increase of pH in the range of 1.9-8.4. In this process, the sorption increases rapidly in the pH range of 1.9-7.2, but becomes slowly at pH>7.2. At last, the highest sorption level is maintained at pH>8.4. It indicates that Eu(III) sorption is strongly dependent on pH, which is consistent with the Eu(II) sorption on some other adsorbents such as alumina oxides^[33], multiwall carbon nanotubes^[34], hydrous alumina^[35], and graphene oxide^[1]. This phenomenon can be explained by the microstructures of Eu(III) on an adsorbent calculated from the EXAFS spectra and the species of Eu(III) on biochar surface. For example, Sheng et al.^[4,5,6] found that the (co)precipitation of Eu(III) could form at high pH, which can partially explain why the sorption maintains in high level at high pH. In addition, the relative distribution of Eu(III) species at different pH calculated from the hydrolysis constants of Eu(III) is shown in Fig. 2(b). The ion Eu³⁺ is the main species at pH<7.3, and Eu(OH)₂⁺ is the dominant species at pH 7.4-8.4. Thereby, it is reasonable that the sorption rapidly increases at pH 1.9-7.2 and slowly increases at pH 7.3-8.4. Besides, the precipitation Eu(OH)₃ is the dominant species at pH>8.4, so that the sorption maintains the high level. In fact, the variation of Eu(III) species at different pH is not the only reason. The surface charge and binding sites may be other reasons. According to protonation reaction (i.e., SOH+H⁺ «SOH₂⁺), the electrostatic repulsion can occur between Eu(III) and the positively- charged sites on biochar in the low uptake of Eu(III) due to large number of binding sites at low pH. Nevertheless, the biochar surface becomes negatively-charged at high pH due to the deprotonation reaction (i.e., SOH«SO^-+ H⁺). Thereby, electrostatic repulsion significantly decreases with pH raising, which enhances the Eu(III) sorption until the formation of the Eu(III) precipitation.

Fig. 2(a) also shows the influence of three NaClO₄ concentrations (i.e., 0.1, 0.01, 0.001 mol/L) on sorption. The sorption of Eu(III) on biochar is independent on ionic strength. In fact, the responses of the sorption to pH and ionic strength generally determine the interaction mechanism between adsorbent and adsorbate. If the sorption is affected by both pH and ionic strength, the interaction is dominated by outer-sphere complexation or ions exchange; if the sorption is affected by pH but not ionic strength, inner-sphere complexation dominates^[21]. Our results show that the pH effects are significant while the effects of different ionic strength are not obvious. Thereby, inner-sphere surface complexation predominated in the sorption of Eu(III) on the biochar.

Both FA and HA can control sorption and transport due to their strong complexation ability with metal ions^[16]. Here, the presence of HA/FA enhances Eu(III) sorption on biochar at pH<7.0, but inhibits Eu(III) uptake at pH>7.0 (Fig. 3). And there is no obvious gap between HA and FA addition at the whole range of pH. The results are similar to effects of HA/FA on the sorption of Eu(III) on GO^[1], γ-Al₂O₃^[16], and compacted bentonite^[18]. HA/FA can mediate Eu(III) sorption by changing the adsorbent surface charges and/or the sites availability^[36]. The Zeta potentials for both FA and HA are negative at pH>2^[35]. Interestingly, the complexation ability of surface adsorbed HA/FA with Eu(III) is stronger than that with adsorbents. Thereby, the negatively-charged HA/FA is easier adsorbed on the positively-charged biochar at low pH. While the positively-charged Eu(III) ions easily form complexes with HA/FA, which enhances Eu(III) sorption on HA/FA-biochar hybrids. At high pH, the solubility of HA/FA increases and the biochar becomes negatively-charged, leading to the electrostatic repulsion. In addition, a part of free stable HA/FA-Eu(III) complexes form in solutions. Thus, the negatively-charged HA/FA is difficult to be adsorbed on the biochar surface, resulting in the decrease of Eu(III) sorption on biochar^[1,16-18].

The contact time of Eu(III) uptake on biochar at 4 pH is shown in Fig. 4(a). The sorption of Eu(III) is fast at the first 2 h but becomes slowly until stable. The possible explanation is that the reactive sites on the adsorbent surface have large binding energies. At beginning, all surface sites of biochar are totally exposed resulting in an rapid increase of Eu(III) sorption. After the sites are almost occupied by Eu(III) ions, Eu(III) sorption becomes slow and diffuses into the micropore sites^[37]. Notably, the contact time at 4 pH is similar, which applies that all the sorption processes under 4 pH conditions are dominated by chemical rather than physical sorption^[38,39]. However, the adsorption content is enhanced by pH, which is coincident with the effects of the above-mentioned pH.

Both surface characteristics and diffusion resistance of sorbents control the sorption rate for metals. Thus, common appropriate kinetic models can uncover underlying uptake mechanisms^[40]. To study the specific rate constant of Eu(III) sorption on biochar, pseudo-second order model and intra-particle diffusion model are used to fit the sorption data. The linear form of the pseudo- second order kinetic is calculated as follows^[40,41]:

where k [g/(mg∙h)] is the rate constant; q_e and q_t represent the sorption ability of Eu(III) (mg/g) at equilibrium time and time t (h), respectively. The three dimensional linear plots of t/q_t vs. t at 4 pH are shown in Fig. 4(b). All corresponding kinetic parameters are shown in Table 1, in which the kinetic sorption process at each pH can be well described by the pseudo-second-order rate equation, because all the correlation coefficients (r²) are close to 1.

In the processes of Eu(III) sorption on biochar, the Eu(III) shifting from biochar surface to intra-particle active sites may limit the sorption rate^[42,43]. Thereby, we fit the relative intra-particle diffusion model. It can be expressed as follows^[25]:

where k_int is the intra-particle diffusion rate constant, and C represents a constant. In Table 1, r² for the particle diffusion model are more than 0.65, suggesting that this model fit the experimental data well. And the 3D plots of q_t vs. t^1/2 for intra-particle diffusion model simulation at 4 pH show that the lines do not pass through the origin (Fig.4(c)). Thus intra-particle diffusion is not the only rate control step but still exert influence on the sorption behavior of Eu(III) on biochar.

Fig. 5 shows that the sorption isotherms of Eu(III) on biochar. The amount of Eu(III) sorption increases with temperature elevating, indicating that the temperature can promote Eu(III) sorption. To explore the sorption mechanism, we simulate sorption isotherms using 4 equilibrium models, i.e., Langmuir, Freundlich, Temkin, and Dubinin-Radushkevich (D-R) model. Their linear forms are expressed as:

where q_max(mg/g) represents Eu(III) maximum sorption ability, K_L (L∙mg^-1) is the Langmuir affinity parameter; K_F (mg^1-n∙Lⁿ∙g^-1) is the Freundlich sorption capacity parameter; n is the degree of sorption dependence on equilibrium concentration; R is the universal gas constant (8.314 J∙K^-1∙mg^-1); T (K) is the Kelvin temperature; b is the sorption heat; A is the binding constant; β and ε are the D-R activity constant and Polanyi potential, ε is calculated by RTln(1+1/C_e). In D-R model, we also calculate the bonding energy (E, kJ/g) of the ion-exchange mechanism using the following equation:

Fig. 6 shows the linear plots of 4 isotherm models at 3 temperatures. And their parameters are listed in Table 2. The values of r² imply that the four models can fit the experimental data well. From the Langmuir model, the maximum sorption amount increases from 34.626 mg/g to 40.717 mg/g with the temperature increasing (Table 2), indicating that high temperature benefits Eu(III) sorption. The n values from the Freundlich model are less than 1, and corresponding r² values are more than 0.97. The two results indicate that Eu(III) sorption is a nonlinear process taking place on heterogeneous biochar surface^[44]. The K_F magnitude increases with the temperature elevating, suggesting that high temperature benefits Eu(III) sorption. The linear plots for the Temkin and D-R sorption isotherm also fit favorably with r²>0.801 at 3 temperatures (Fig. 6 & Table 2). The result further supports the findings that Eu(III) sorption is a chemisorption process and enhanced by the high temperature.

Among the 4 isotherm models, the Freundlich model fits better than the other models due to the highest r², which implies that the Eu(III) sorption completes on heterogeneous biochar surface. The magnitude of q_max derived from Langmuir isotherm and D-R isotherm has a large gap, which may be arise from the different assumptions between the two isotherm formulations. Totally, these results suggest that the rice straw-derived biochar can be considered as a novel and effective adsorbent for the removal of Eu(III) from water.

To investigate the thermodynamic properties, the thermodynamic parameters (∆H^θ, ∆S^θ, and ∆G^θ) were calculated from sorption isotherms at 3 temperatures. The free energy change (∆G^θ) of specific sorption is calculated as^[45]:

where K^θ is the equilibrium constant of sorption reaction. The values of lnK^θ are obtained by plotting lnK_dvs. C_e (Fig. 7(a)). Standard entropy (∆S^θ) and the average stan dard enthalpy (∆H^θ) are calculated from the slope and intercept of the plot lnK^θversus 1/T (Fig. 7(b)), respectively. The equation is as follows:

ln(K^θ) = ∆S^θ/R - ∆H^θ/RT(9)

Table 3 lists the thermodynamic parameters, which is beneficial to uncover interactive mechanisms of Eu(III) and biochar. Specifically, the negative values of ∆G^θ indicate the spontaneous process of Eu(III) sorption under the experimental conditions. More negative values of ∆G^θ at higher temperature (338 K) suggest that high temperature can provide more energy for favorable sorption. The reason may be that Eu(III) are readily desolvated and its sorption are promoted at high temperature^[1,13]. The positive values of ∆H^θ indicates the endothermic process of Eu(III) sorption on biochar. Also, there are strong interactions between Eu(III) and biochar. One reason is that Eu(III) has favorable solubility in solutions, and the hydration sheath of Eu(III) needs to be stripped out before its sorption on biochar. High temperature can provide more energy to favor the dehydration process. The positive value of ∆S^θ implies partial structural variations in Eu(III) and biochar during sorption process, resulting in randomness increasing of the biochar-solution interfacial system. Overall, the thermodynamic analysis illustrates that the sorption process of Eu(III) on biochar is a spontaneous and endothermic, and the sorption can be enhanced by high temperature.

Here, the sorption behavior of Eu(III) on rice straw-derived biochar was strongly dependent on various environmental factors, including pH, humic substances, contact time, temperature. Interestingly, the effect of each environmental factor is affected by solution pH, implying that the sorption process is complicated in natural polluted water systems, and we should consider multi-factors when remove toxic metals. According to the performance of the sorption of biochar for Eu(III), we found that surface (co)precipitation or inner-sphere surface complexation dominates the Eu(III) sorption process in the whole pH range; this chemical sorption rate was limited by intra-particle diffusion process. Furthermore, the Eu(III) sorption takes place on a heterogeneous surface of biochar which is a spontaneous and endothermic process. The research helps us to understand the sorption properties of the rice straw-derived biochar for Eu(III) and its potential application value. In future, some microscopic technologies such as XPS, EXAFS should be used to uncover the underlying mechanisms.