Removal of 241Am from Aqueous Solutions by Adsorption on Sponge Gourd Biochar

Luffa cylindrica biomass was converted to biochar and the removal of 241Am by pristine and oxidized biochar fibers was investigated in laboratory and environmental water samples. This species has the added advantage of a unique microsponge structure that is beneficial for the production of porous adsorbents. The main purpose of this study was to valorize this biomass to produce an efficient adsorbent and investigate its performance in radionuclide-contaminated waters. Following the preparation of Am3+ solutions at a concentration of 10−12 mol/L, the adsorption efficiency (Kd) was determined as a function of pH, adsorbent mass, ionic strength, temperature, and type of aqueous solution by batch experiments. At the optimum adsorbent dose of 0.1 g and pH value of 4, a log10Kd value of 4.2 was achieved by the oxidized biochar sample. The effect of temperature and ionic strength indicated that adsorption is an endothermic and entropy-driven process (ΔH° = −512 kJ mol−1 and ΔS° = −1.2 J K−1 mol−1) leading to the formation of inner-sphere complexes. The adsorption kinetics were relatively slow (24 h equilibrium time) due to the slow diffusion of the radionuclide to the biochar surface and fitted well to the pseudo-first-order kinetic model. Oxidized biochar performed better compared to the unmodified sample and overall appears to be an efficient adsorbent for the treatment of 241Am-contaminated waters, even at ultra-trace concentrations.


Introduction
Americium is a man-made, radioactive metal, and its most common isotopes are 241 Am and 243 Am. In small amounts, americium is present in uranium minerals due to nuclear reactions that may occasionally occur. Most americium is produced in nuclear reactors when neutrons are captured by uranium or plutonium, and about 100 g of americium is contained in one ton of spent nuclear fuel. No natural sources of americium exist, therefore, its presence in the environment is mainly due to nuclear weapons testing or accidental releases from nuclear power plants [1,2]. The americium isotope 241 Am (t 1/2 = 432.2 y) can be easily produced in pure form and therefore has several applications. It is used in smoke detectors, lightning rods, and portable sources of alpha and gamma rays, which have found application in a number of medical and industrial uses. Particularly, the 60 keV gamma ray emission of 241 Am is used in radiography and X-ray fluorescence spectroscopy for material analysis and quality control [1,2]. The appropriate handling and disposal of such sources is critical because americium released in the environment may cause health effects to living organisms including humans. After uptake, americium is rapidly transported to the bones where it can be stored for a long period of time. There, the radionuclide decays slowly, emitting alpha-particles and gamma rays, which can alter genetic material and cause bone cancer [1,2]. Hence, it is obligatory to remove americium from contaminated environments including natural waters and wastewaters.
Although several oxidation states are known, ranging from +2 to +7, 241 Am in solutions exists predominantly in the trivalent oxidation state (Am 3+ ). The speciation of Am 3+ in

Effect of Contact Time on the 241 Am Adsorption
Adsorption kinetics describes the relationship between the mass transfer of the adsorbate and time in the adsorption process, which is helpful to determine the time needed for the system to reach equilibrium as well as the kinetic parameters associated with the rate of the adsorption process. The time needed to reach equilibrium in the system has been investigated by determining the relative quantity of Am 3+ adsorbed by the oxidized biochar fibers as a function of contact time. The associated kinetic data are summarized in Figure 1 and reveal that equilibrium was reached after 24 h. Therefore, subsequent experiments have been carried out allowing for 24 h contact time to assure equilibrium. In the sub-picomolar concentration range, equilibrium was reached after 24 h contact time, which is a significantly higher contact time compared to the few hours required at increased radionuclide/metal ion concentrations [7][8][9]24,38]. This is because at very low concentration levels, the diffusion of the radionuclide cations to the biochar surface is the adsorption rate determining step, and the mass transfer phenomena are reduced. Shorter equilibrium times in the range of 30-180 min have been reported in the literature for 241 Am, however, using much higher Am 3+ concentrations and highly modified, silica-based adsorbents [38,39]. Furthermore, the kinetic data have been fitted with the pseudo-first-order (PFO) and pseudo-second-order (PSO) kinetic models and the results indicate that the data were better fitted with the PSO (R = 0.999) compared to the PFO (R = 0.952) kinetic model. values; and (c) compare the performance of the biochar adsorbent in three distinctly different aqueous environments: seawater, groundwater, and wastewater.

Effect of Contact Time on the 241 Am Adsorption
Adsorption kinetics describes the relationship between the mass transfer of the adsorbate and time in the adsorption process, which is helpful to determine the time needed for the system to reach equilibrium as well as the kinetic parameters associated with the rate of the adsorption process. The time needed to reach equilibrium in the system has been investigated by determining the relative quantity of Am 3+ adsorbed by the oxidized biochar fibers as a function of contact time. The associated kinetic data are summarized in Figure 1 and reveal that equilibrium was reached after 24 h. Therefore, subsequent experiments have been carried out allowing for 24 h contact time to assure equilibrium. In the sub-picomolar concentration range, equilibrium was reached after 24 h contact time, which is a significantly higher contact time compared to the few hours required at increased radionuclide/metal ion concentrations [7][8][9]24,38]. This is because at very low concentration levels, the diffusion of the radionuclide cations to the biochar surface is the adsorption rate determining step, and the mass transfer phenomena are reduced. Shorter equilibrium times in the range of 30-180 min have been reported in the literature for 241 Am, however, using much higher Am 3+ concentrations and highly modified, silicabased adsorbents [38,39]. Furthermore, the kinetic data have been fitted with the pseudofirst-order (PFO) and pseudo-second-order (PSO) kinetic models and the results indicate that the data were better fitted with the PSO (R = 0.999) compared to the PFO (R = 0.952) kinetic model.

Effect of pH on 241 Am Adsorption
The solution pH strongly affects the sorption efficiency (Kd values), since both the Am 3+ species distribution in solution and the protonation/dissociation of the surface-active moieties of the biochar materials depend on the proton concentration. The latter applies to the π system-proton interaction and carboxylic dissociation for unmodified and oxidized biochar, respectively [20][21][22][23][24]. A schematic illustration of the interaction of Am 3+ cations with the aromatic (π-system) and the carboxylic groups on the biochar surface is given in Figure 2.

Effect of pH on 241 Am Adsorption
The solution pH strongly affects the sorption efficiency (K d values), since both the Am 3+ species distribution in solution and the protonation/dissociation of the surface-active moieties of the biochar materials depend on the proton concentration. The latter applies to the π system-proton interaction and carboxylic dissociation for unmodified and oxidized biochar, respectively [20][21][22][23][24]. A schematic illustration of the interaction of Am 3+ cations with the aromatic (π-system) and the carboxylic groups on the biochar surface is given in Figure 2.  Depending on the solution pH, Am 3+ exists in acidic solutions predomin form of the Am 3+ cation, at near neutral solutions predominantly as Am 3+ , Am Am(CO3) + , and under alkaline conditions, the stable Am(CO3)2 − is the domina [4][5][6]40]. Regarding the biochar surface, at pH = 2, the carboxylic moieties were protonated, because the mean value of the acid dissociation constant, pKa, w At pH 4, a significant number of the carboxylic moieties was deprotonated and = 7 and 9 did the deprotonated carboxylic groups dominate, and the biochar negatively charged [20][21][22][23].
The effect of pH on the Am 3+ adsorption by biochar fibers has been inves the corresponding log10Kd values are summarized in Figure 3. The oxidized bi (BC_ox) showed a much higher adsorption affinity for Am 3+ than their non-oxi terpart (BC), which may be attributed to the presence of a significantly higher tion of oxygen-containing moieties (e.g., carboxylic groups) on the BC-ox surf Oxygen-containing moieties (e.g., phenolic, carboxylic groups) are hard Lewi therefore exhibit a higher affinity for hard Lewis acids such as the Am 3+ catio Hence, the highest Kd values for BC_ox were observed in the acidic pH reg (BC_ox) = 4.1 ± 0.1 at pH = 2 and log10Kd (BC_ox) = 4.2 ± 0.2 at pH = 4, and the lo (log10Kd (ΒC) = 3.4 ± 0.3 at pH = 7 and log10Kd (ΒC_ox) = 3.5 ± 0.2 at pH = 9. The s higher affinity of the oxidized biochar for Am 3+ in the acidic pH region can b to the predominance of the Am 3+ cation in the respective pH region and th deprotonated, negatively charged surface carboxylic moieties, which strongly complex the Am 3+ cations, resulting in the formation of inner-sphere comple These observations agree well with Li et al. (2001) and Zhang et al. (2014), wh comparable Kd values at acidic pH [40,33]. In the neutral and alkaline pH r progressive formation of the Am(OH)2 + , Am(CO3) + , and Am(CO3)2 − species r decline in the electrostatic attraction and to some extent hindered the surfa formation. On the other hand, the pristine biochar fibers presented the lowest the strongly acidic pH region (log10Kd (ΒC) = 2.1 ± 0.3 at pH = 2), whereas the v 4, 7, and 9 were comparable to the values obtained by BC_ox. The lower sorpt of pristine biochar for Am 3+ at pH = 2 is related to the increased concentration which compete with the Am 3+ cations based on cation-pi interaction betwee tively charged species and the aromatic rings of the biochar [20][21][22][23][24]. Generall of pH on the Am 3+ adsorption by biochar fibers in the picomolar concentrati similar to the adsorption of its chemical analogue (Eu 3+ ) [39][40][41]. Depending on the solution pH, Am 3+ exists in acidic solutions predominantly in the form of the Am 3+ cation, at near neutral solutions predominantly as Am 3+ , Am(OH) 2 + , and Am(CO 3 ) + , and under alkaline conditions, the stable Am(CO 3 ) 2 − is the dominating species [4][5][6]40]. Regarding the biochar surface, at pH = 2, the carboxylic moieties were extensively protonated, because the mean value of the acid dissociation constant, pK a , was below 4. At pH 4, a significant number of the carboxylic moieties was deprotonated and only at pH = 7 and 9 did the deprotonated carboxylic groups dominate, and the biochar surface was negatively charged [20][21][22][23].
The effect of pH on the Am 3+ adsorption by biochar fibers has been investigated and the corresponding log 10 K d values are summarized in Figure 3. The oxidized biochar fibers (BC_ox) showed a much higher adsorption affinity for Am 3+ than their non-oxidized counterpart (BC), which may be attributed to the presence of a significantly higher concentration of oxygen-containing moieties (e.g., carboxylic groups) on the BC-ox surface [20][21][22][23][24]. Oxygen-containing moieties (e.g., phenolic, carboxylic groups) are hard Lewis bases and therefore exhibit a higher affinity for hard Lewis acids such as the Am 3+ cationic species. Hence, the highest K d values for BC_ox were observed in the acidic pH region log 10 K d (BC_ox) = 4.1 ± 0.1 at pH = 2 and log 10 K d (BC_ox) = 4.2 ± 0.2 at pH = 4, and the lowest values (log 10 K d (BC) = 3.4 ± 0.3 at pH = 7 and log 10 K d (BC_ox) = 3.5 ± 0.2 at pH = 9. The significantly higher affinity of the oxidized biochar for Am 3+ in the acidic pH region can be attributed to the predominance of the Am 3+ cation in the respective pH region and the partially deprotonated, negatively charged surface carboxylic moieties, which strongly attract and complex the Am 3+ cations, resulting in the formation of inner-sphere complexes [20][21][22][23][24]. These observations agree well with Li et al. (2001) and Zhang et al. (2014), who reported comparable K d values at acidic pH [33,40]. In the neutral and alkaline pH regions, the progressive formation of the Am(OH) 2 + , Am(CO 3 ) + , and Am(CO 3 ) 2 − species resulted in a decline in the electrostatic attraction and to some extent hindered the surface complex formation. On the other hand, the pristine biochar fibers presented the lowest K d value in the strongly acidic pH region (log 10 K d (BC) = 2.1 ± 0.3 at pH = 2), whereas the values at pH 4, 7, and 9 were comparable to the values obtained by BC_ox. The lower sorption affinity of pristine biochar for Am 3+ at pH = 2 is related to the increased concentration of protons, which compete with the Am 3+ cations based on cation-pi interaction between the positively charged species and the aromatic rings of the biochar [20][21][22][23][24]. Generally, the effect of pH on the Am 3+ adsorption by biochar fibers in the picomolar concentration range is similar to the adsorption of its chemical analogue (Eu 3+ ) [39][40][41].

Effect of Ionic Strength on 241 Am Adsorption
At increased metal ion concentrations, spectroscopic characteriza form, X-ray photoelectron, and Raman) of the adsorbed species and ev sorption mechanism at the molecular level is feasible, however, at ult cannot occur. Therefore, the effect of the ionic strength on the adsorp investigated to gain insights on the adsorption efficiency and evaluate complexes formed (e.g., outer-or inner-sphere complexes). Gene strength of the solution increases, a decline of the adsorption efficienc acteristic of non-specific, simple electrostatic interactions and the pred sphere complex formation. However, when high ionic strength does no in the adsorption efficiency, the existence of specific interactions and t ner-sphere complexes governs the adsorption process [20][21][22][23][24]. Figur adsorption efficiency (log10Kd values) was not reduced with increasing suming specific interactions and the formation of inner-sphere compl and active moieties (e.g., aromatic rings and carboxylate groups) on t This is in agreement with previous studies on the adsorption of triva oxidized biochar fibers, which had been performed at increased meta (at the mmol/L concentration range) and by means of spectroscopic inv FTIR [41] and XPS [42], which indicated the formation of inner-sphere the metal ion and the surface-active moieties.

Effect of Ionic Strength on 241 Am Adsorption
At increased metal ion concentrations, spectroscopic characterization (Fourier transform, X-ray photoelectron, and Raman) of the adsorbed species and evaluation of the adsorption mechanism at the molecular level is feasible, however, at ultra-trace levels, this cannot occur. Therefore, the effect of the ionic strength on the adsorption efficiency was investigated to gain insights on the adsorption efficiency and evaluate the type of surface complexes formed (e.g., outer-or inner-sphere complexes). Generally, as the ionic strength of the solution increases, a decline of the adsorption efficiency is observed, characteristic of non-specific, simple electrostatic interactions and the predominance of outer sphere complex formation. However, when high ionic strength does not result in a decline in the adsorption efficiency, the existence of specific interactions and the formation of inner-sphere complexes governs the adsorption process [20][21][22][23][24]. Figure 4 shows that the adsorption efficiency (log 10 K d values) was not reduced with increasing ionic strength, assuming specific interactions and the formation of inner-sphere complexes between Am 3+ and active moieties (e.g., aromatic rings and carboxylate groups) on the biochar surface. This is in agreement with previous studies on the adsorption of trivalent lanthanides by oxidized biochar fibers, which had been performed at increased metal ion concentration (at the mmol/L concentration range) and by means of spectroscopic investigations such as FTIR [41] and XPS [42], which indicated the formation of inner-sphere complexes between the metal ion and the surface-active moieties.

Effect of Temperature on 241 Am Adsorption
The thermodynamic parameters (ΔH° and ΔS°) of the Am 3+ adsorp and oxidized biochars were evaluated by determining the associated different temperatures and plotting lnKd versus 1/T, according to the v The plot of lnKd against 1/T is shown in Figure 5. The thermod were evaluated by the slope and intercept obtained from the linear re responding experimental data as ΔH° = −206 kJ mol -1 and ΔS°= −0.5 J K kJ mol -1 and ΔS°= −1.2 J K -1 mol -1 , for the adsorption of Am 3+ on the pr biochar, respectively. Furthermore, the ΔG° values at 25, 40, and 60 °C oxidized biochar were −56.9 kJ mol -1 , −49.4 kJ mol -1 , −39.4.9 kJ mol -1 a −136.2 kJ mol -1 , −112.4.2 kJ mol -1 , respectively. These values indicate sorption process, in contrast to the observations obtained from the ads the same adsorbents and the same concentration range. The corres U(VI) were ΔH° = 1.5 kJ mol -1 and ΔS° = 1.6 kJ K -1 mol -1 [23]. Moreov namic parameters differed from the corresponding parameters obtain ments performed using Sm 3+ as the Am 3+ analogue (ΔH° = 34 kJmol -1 mol -1 ) was performed at higher (mmol range) metal ion concentratio be attributed to the fact that at higher metal ion concentrations, the as

Effect of Temperature on 241 Am Adsorption
The thermodynamic parameters (∆H • and ∆S • ) of the Am 3+ adsorption by the pristine and oxidized biochars were evaluated by determining the associated K d values at three different temperatures and plotting lnK d versus 1/T, according to the van't Hoff equation: The plot of lnK d against 1/T is shown in Figure 5. The thermodynamic parameters were evaluated by the slope and intercept obtained from the linear regression of the corresponding experimental data as ∆H  [23]. Moreover, these thermodynamic parameters differed from the corresponding parameters obtained from the experiments performed using Sm 3+ as the Am 3+ analogue (∆H • = 34 kJ mol −1 and ∆S • = 193 JK −1 mol −1 ) was performed at higher (mmol range) metal ion concentrations [42]. This could be attributed to the fact that at higher metal ion concentrations, the assumption of a large excess of binding site species is not relevant, and different adsorption mechanisms prevail on the biochar surface.

Removal of 241 Am from Seawater, Groundwater, and Wastewater
The removal of 241 Am from the seawater, groundwater, and wastewater from the local wastewater treatment plant was investigated after spiking the respective solutions with 241 Am and in contact with two different biochar doses of 0.01 and 0.1 g. After a 24 h contact time, the radionuclide concentration in the solution was analyzed by alpha spectroscopy. Figure 5 shows the characteristic alpha-spectra of 241 Am corresponding to the contaminated seawater samples prior and after treatment with 0.01 and 0.1 g of the pristine and oxidized biochars. The spectra in Figure 6 clearly indicate that the 241 Am levels declined in the presence of biochar fibers, particularly in the presence of the oxidized counterparts. The latter was attributed to the strong affinity of the carboxylic moieties, which were present on the BC_ox surface [20][21][22][23] for Am 3+ , resulting in the formation of inner-sphere complexes. Figure 6 shows that increasing the amount of biochar resulted in higher removal efficiencies. Based on the alpha spectra obtained for the three different aqueous matrices and the two different doses of BC and BC_ox, the relative quantity of Am 3+ removed from the solution was calculated and the corresponding data are graphically summarized in Figure   Figure 5. lnK d as a function of 1/T for the adsorption of Am 3+ by oxidized biochar fibers at an initial 241 Am concentration of 1 × 10 −12 mol/L, biochar mass = 0.01 g, pH = 4, and 3 days of contact time.

Removal of 241 Am from Seawater, Groundwater, and Wastewater
The removal of 241 Am from the seawater, groundwater, and wastewater from the local wastewater treatment plant was investigated after spiking the respective solutions with 241 Am and in contact with two different biochar doses of 0.01 and 0.1 g. After a 24 h contact time, the radionuclide concentration in the solution was analyzed by alpha spectroscopy. Figure 5 shows the characteristic alpha-spectra of 241 Am corresponding to the contaminated seawater samples prior and after treatment with 0.01 and 0.1 g of the pristine and oxidized biochars. The spectra in Figure 6 clearly indicate that the 241 Am levels declined in the presence of biochar fibers, particularly in the presence of the oxidized counterparts. The latter was attributed to the strong affinity of the carboxylic moieties, which were present on the BC_ox surface [20][21][22][23] for Am 3+ , resulting in the formation of inner-sphere complexes. Figure 6 shows that increasing the amount of biochar resulted in higher removal efficiencies.

Removal of 241 Am from Seawater, Groundwater, and Wastewater
The removal of 241 Am from the seawater, groundwater, and wastewater from the local wastewater treatment plant was investigated after spiking the respective solutions with 241 Am and in contact with two different biochar doses of 0.01 and 0.1 g. After a 24 h contact time, the radionuclide concentration in the solution was analyzed by alpha spectroscopy. Figure 5 shows the characteristic alpha-spectra of 241 Am corresponding to the contaminated seawater samples prior and after treatment with 0.01 and 0.1 g of the pristine and oxidized biochars. The spectra in Figure 6 clearly indicate that the 241 Am levels declined in the presence of biochar fibers, particularly in the presence of the oxidized counterparts. The latter was attributed to the strong affinity of the carboxylic moieties, which were present on the BC_ox surface [20][21][22][23] for Am 3+ , resulting in the formation of inner-sphere complexes. Figure 6 shows that increasing the amount of biochar resulted in higher removal efficiencies. Based on the alpha spectra obtained for the three different aqueous matrices and the two different doses of BC and BC_ox, the relative quantity of Am 3+ removed from the solution was calculated and the corresponding data are graphically summarized in Figure   Figure 6. Alpha spectra of 241 Am contaminated seawater samples prior and after contact with the 0.01 g and 0.1 g biochar fibers prior (BC) and after (BC_ox) oxidation. [ 241 Am] = 1 × 10 −12 mol/L, ambient conditions. Based on the alpha spectra obtained for the three different aqueous matrices and the two different doses of BC and BC_ox, the relative quantity of Am 3+ removed from the solution was calculated and the corresponding data are graphically summarized in Figure 7. The oxidized biochar fibers presented the highest removal efficiency (~80%) in all three natural waters. Increasing the biochar dose from 0.01 to 0.1 g resulted in a 10-40% increase in the removal efficiency, depending on the aqueous system. At the highest adsorbent dose of 0.1 g, the removal of 241 Am by the oxidized biochar was twice that of the pristine sample (~80 compared to~40%) for the seawater and wastewater. In the case of the groundwater, the efficiency of the oxidized biochar was slightly reduced, but was nevertheless considerably higher compared to the pristine biochar, with 70% removal compared to~50%, respectively. The slightly decreased performance of the oxidized biochar on the groundwater may be attributed to the elevated concentrations of Ca 2+ and Fe 3+ ions, which may compete with Am 3+ for sites on the adsorbent surface [23,42]. These observations confirm the high adsorption affinity of the oxidized biochar fibers toward Am 3+ and establish the applicability of the material for the treatment of americium contaminated natural waters and wastewaters.
7. The oxidized biochar fibers presented the highest removal efficienc natural waters. Increasing the biochar dose from 0.01 to 0.1 g resulted i in the removal efficiency, depending on the aqueous system. At the dose of 0.1 g, the removal of 241 Am by the oxidized biochar was twice sample (~80 compared to ~40%) for the seawater and wastewater. groundwater, the efficiency of the oxidized biochar was slightly redu theless considerably higher compared to the pristine biochar, with 70% to ~50%, respectively. The slightly decreased performance of the oxid groundwater may be attributed to the elevated concentrations of Ca 2+ a may compete with Am 3+ for sites on the adsorbent surface [23,42]. Thes firm the high adsorption affinity of the oxidized biochar fibers toward the applicability of the material for the treatment of americium contam ters and wastewaters. Furthermore, the alpha spectroscopic data were used to evaluate ciated with each biochar sample for the adsorption of Am 3+ from the corresponding to three different water systems. The calculated Kd valu in Figure 8 and were similar (log10Kd ~ 3.5 at pH = 7 and pH = 9) to values determined in the laboratory, de-ionized water solutions for th (BC_ox), and about half a logarithmic unit lower (log10Kd ~ 2.5 at pH the non-oxidized counterpart. The lower Kd values associated with th can be attributed to the presence of competing cations in the natural Ca 2+ , Fe 3+ ), which may be because hard Lewis acids interact and occup sites on the biochar surface, resulting in adsorption affinity for the respective systems [23,41]. Moreover, in the case of the pristine biochar cations (e.g., Na + , K + ) may interact with the biochar surface via cation tions, significantly affecting the Am 3+ adsorption by the pristine biocha Furthermore, the alpha spectroscopic data were used to evaluate the K d values associated with each biochar sample for the adsorption of Am 3+ from the aqueous solutions corresponding to three different water systems. The calculated K d values are summarized in Figure 8 and were similar (log 10 K d~3 .5 at pH = 7 and pH = 9) to the corresponding values determined in the laboratory, de-ionized water solutions for the oxidized biochar (BC_ox), and about half a logarithmic unit lower (log 10 K d~2 .5 at pH = 7 and pH = 9) for the non-oxidized counterpart. The lower K d values associated with the seawater samples can be attributed to the presence of competing cations in the natural water system (e.g., Ca 2+ , Fe 3+ ), which may be because hard Lewis acids interact and occupy the surface active sites on the biochar surface, resulting in adsorption affinity for the Am 3+ cations in the respective systems [23,41]. Moreover, in the case of the pristine biochar, even conservative cations (e.g., Na + , K + ) may interact with the biochar surface via cation-π system interactions, significantly affecting the Am 3+ adsorption by the pristine biochar materials. Nevertheless, the evaluated K d values were similar to the mean linear distribution value (log 10 K d = 3.7) reported for Am 3+ in soils [43], indicating the supreme affinity of the oxidized biochar fibers for Am 3+ .

Materials and Methods
All experiments were carried out in 30 mL polyethylene (PE) sc under ambient conditions (23 ± 2 °C). The americium isotope, 241 Am, w experiments ( 241 Am standard tracer solution, 7.367 kBq/g, North Ame Los Angeles, CA, USA). This standard solution was diluted to prepar solutions with an initial concentration of 1 mBq/mL. The biochar sam from the pyrolysis of the vegetable sponge gourd. The preparation performed by carbonization at 650 °C under inert (N2) atmosphere fo idation was carried out by treating the biochar with 8 M HNO3 for 1 h ization of the materials by surface and spectroscopic methods has b scribed elsewhere [20][21][22]. It is worth mentioning that, in contrast t biochar fibers, which are characterized by the presence of graphite she ylic moieties govern the surface charge and chemistry of oxidized bio periments were performed in laboratory solutions using de-ionized w regions (pH 2, 4, 7, and 9) and in naturally-occurring aqueous matric water (GW), wastewater (WW), and seawater (SW) samples. The grou pled from a local well; the wastewater, which corresponds to the ef treatment, was obtained from a municipal wastewater treatment plan sample was collected from a coastal area of the island. The pH and m

Materials and Methods
All experiments were carried out in 30 mL polyethylene (PE) screw capped bottles under ambient conditions (23 ± 2 • C). The americium isotope, 241 Am, was employed in the experiments ( 241 Am standard tracer solution, 7.367 kBq/g, North American Scientific Inc., Los Angeles, CA, USA). This standard solution was diluted to prepare reference and test solutions with an initial concentration of 1 mBq/mL. The biochar samples were prepared from the pyrolysis of the vegetable sponge gourd. The preparation of the biochar was performed by carbonization at 650 • C under inert (N 2 ) atmosphere for two hours, the oxidation was carried out by treating the biochar with 8 M HNO 3 for 1 h, and the characterization of the materials by surface and spectroscopic methods has been extensively described elsewhere [20][21][22]. It is worth mentioning that, in contrast to the non-modified biochar fibers, which are characterized by the presence of graphite sheets, and the carboxylic moieties govern the surface charge and chemistry of oxidized biochar fibers. The experiments were performed in laboratory solutions using de-ionized water at different pH regions (pH 2, 4, 7, and 9) and in naturally-occurring aqueous matrices such as groundwater (GW), wastewater (WW), and seawater (SW) samples. The groundwater was sampled from a local well; the wastewater, which corresponds to the effluent of secondary treatment, was obtained from a municipal wastewater treatment plant; and the seawater sample was collected from a coastal area of the island. The pH and main components of the environmental waters, which have been analyzed as described elsewhere [10,11], are summarized in Table 1. The pH in the laboratory solutions was adjusted using either 0.1 M HCl or 0.1 M NaOH solutions and the pH measurements were carried out using a conventional laboratory pH meter (pH 211 microprocessor pH meter, Hanna Instruments, Woonsocket, RI, USA). The radiometric analysis of 241 Am was carried out using an alpha-spectrometer (Alpha Analyst Integrated Alpha Spectrometer, Canberra Industries, Fussy, France), as described elsewhere [7,38]. In addition, reference and control samples of the radionuclide were also analyzed by liquid scintillation counting (LSC, Triathler, Hidex, Turku, Finland). Alphaspectrometric analysis was performed in triplicate and the LSC measurements were carried out in parallel to compare and validate the data obtained from both radiometric methods.
The mean values and the associated standard deviations were used for the graphical presentations. The detection limits were determined at 0.05 mBq and 0.03 mBq for the LSC and alpha-spectrometric measurements, respectively.
The adsorption studies were carried out by adding 0.01 or 0.1 g of the biochar to 20 mL of the radionuclide solution at an activity concentration of 25 Bq/L ([ 241 Am] = 1 × 10 −12 mol/L) in 30 mL screw capped PE vials. The effect of contact time was investigated at pH = 4 and under the aforementioned experimental conditions. The effect of pH was studied at pH = 2, 4, 7, and 9, and the effect of ionic strength (I) was investigated using aqueous NaClO 4 solutions at various concentrations (no electrolyte added, 0.05, 0.1, 0.5, and 1 M). The effect of temperature was investigated at 25, 40, and 60 • C, at pH = 4. The biochar-radionuclide solution suspensions were agitated on a shaker (SK-R1807, DLAB, Beijing, China) at an agitation rate of 65 rpm for 24 h, after which the suspension was allowed to rest. For the analysis of 241 Am, aliquots of 200 µL were obtained from the supernatant to evaluate the radionuclide concentration in solution by alpha-spectrometry or liquid scintillation counting. The radiometric methods were calibrated using standard reference solutions and sources. The sorption efficiency was evaluated using the partition coefficient (K d ). The K d applies because of the picomolar 241 Am concentrations used and the relatively large excess of the available surface binding sites. The following equation describes the partition coefficient: where [Am 3+ ] ads is the activity of 241 Am adsorbed by the biochar fibers (Bq/g) and [Am 3+ ] aq (Bq/L) is the 241 Am activity concentration in solution at equilibrium. The quantity of 241 Am adsorbed by biochar fibers (dry mass) was calculated from the total 241 Am activity adsorbed by subtracting the 241 Am activity adsorbed by the vial walls. The latter was not negligible and had to be taken into account. Therefore, for each separate test solution (also including the effect of pH, temperature, and other parameters), a reference solution under exactly the same conditions (without the biochar material) was prepared, and the activity concentration of americium was determined in parallel. In addition, under the experimental conditions, Am 3+ is expected to be the predominant oxidation state in solution [1][2][3][4][5][6][7]. Furthermore, the sorption efficiency was expressed as % relative removal and was calculated as follows: where [Am 3+ ] R is the 241 Am concentration in the reference solutions. The experiments were performed in triplicate and the mean values and uncertainties were used for the data evaluation and graphical presentations.

Conclusions
Given the scarcity of data on treatment methods for 241 Am-contaminated waters and the established hazard from the emission of alpha particles from its decay, there is a clear need to develop and optimize the adsorption processes. In this work, it was demonstrated that a low-cost, carbonaceous adsorbent is suitable for the treatment of 241 Am radioactivity in a range of aqueous environments, at very low concentrations. Compared to conventional adsorbents based on silica or other inorganic oxides, biochar has the added advantage of coming from a renewable biomass, thus rendering it more environmentally friendly. Of the two biochar samples tested, it appeared that the oxidized biochar was more efficient due to the higher number of carboxylic moieties. The performance of this sample was maintained at more complex, naturally-occurring waters, a promising observation toward scaling-up of the process. Furthermore, biochar materials, in general, particularly after oxidation, exhibit high physical and chemical stability. In addition, numerous previous investigations on the recoverability of biochar after adsorption have shown that the material retains its increased adsorption efficiency even after five adsorption-desorption cycles.
Since the published results on the treatment of radionuclide-contaminated water and wastewater by biomass-based materials is limited, future work will focus on preparing engineered biochars to further improve the adsorption capacity. Additionally, given the complexity of real wastewater samples from nuclear power stations, the selectivity of such adsorbents for specific residual radionuclides should be investigated.