Biosorption Capability of Chitosan for Removal of Cs-137 and/or Co-60 from Radioactive Waste Solution Simulates

Abstract

Biosorption is an impurity-free application developed from the use of nuclear technology for peaceful purposes in everyday life and can be used to treat wastewater streams contaminated with various radionuclides. In this study, a laboratory decontamination experimental approach was developed to apply commercial chitosan as a biosorbent applied for removing radiocesium (Cs-137) and/or radiocobalt (Co-60) from spiked aqueous media. The factors assumed to affect the biosorption of both radionuclides included contact time, pH, and initial radioactivity content. In addition, the biosorbent dose and temperature of the process were studied. Both the biosorption capacity and the biosorption efficiency of the treatment process were calculated. According to FT-IR analysis, it can be assumed that the chitosan amine group (-NH₂) is almost accountable for the biosorption of both radionuclides from waste solution simulates. Based on the data obtained, commercial chitosan can be considered an economical and efficient biosorbent for handling low- and medium-level radioactive wastewater streams containing cesium and/or cobalt radionuclides. The acquired data showed that 144 h is an adequate time to remove more than 94% of radiocobalt and about 93% of radiocesium, from a separate solution for each, at pH ~6.5 and using 0.5 g of commercial chitosan.

1. Introduction

Recent years have seen a rapid development of nuclear technologies in our daily lives, with the construction of several nuclear power plants across the globe [1]. Due to this development, nuclear accidents that pose a threat to the world’s ecosystems raise an increasing concern about the potential effects of radioactive pollutants released into the environment. In a nation with or without nuclear fuel cycle operations, nuclear research, radioisotope manufacturing, and application, in addition to decontaminating and dismantling of nuclear sites, are the main sources of radioactive waste. Radioactive waste comes in a variety of chemical and physical forms and contains a wide range of radionuclides classified into three main categories, namely low-, intermediate-, and high-level radioactive waste. Several techniques have been used for the elimination of radioactive nuclides from spiked waste streams generated due to nuclear applications in our daily life [2]. These techniques include ion exchange, reverse osmosis, membrane separation, chemical treatment, and nanofiltration, in addition to other processes [3,4]. These methods can proceed unaccompanied or in combination with other methods, e.g., chemical precipitation before nanofiltration [5]. New methods have been developed and characterized as cost-effective techniques to lower or totally eliminate radio-ions to their allowable limits, consequently preparing the treated waste stream for the next management steps. Sorption has been demonstrated to be of a distinguished pattern because of its simple device and operation, cost-effectiveness, efficiency, and exceptional chelating performance; moreover, it can be regenerated using desorption processes [6]. This sorption process could be followed by stabilization using cement mixed with various additives, such as bitumen [7], asphaltene [8], glass [9], polymer [10], cement based-materials [11], and natural clay [12].

Chitosan (CS) can be a low-cost, natural, effective, and sustainable material candidate for the treatment of radioactive waste streams. CS is an N-deacetylation derivative of naturally occurring chitin polysaccharide that can be derived from crustaceans and microbes [13]. The inclusion of both functional groups, amine (-NH₂) and hydroxyl (OH), in the backbone of CS provides a unique sorption characteristic to the polymer. Amine groups are principally responsible for the sorption of metal cations from waste solutions. Three different molecular weights of chitosan have been modified using p-coumaric acid (p-CA) to enhance its water solubility and antioxidant property [14].

Commercial CS can originate from the shells of shrimps and similar sea crustaceans, in addition from some insects and fungi. In general, the biosorption practice is supposed to occur through three successive steps: film diffusion, intraparticle diffusion, and biosorption [15]. Initially, film diffusion facilitates the external mass transfer of radioactive ions from the bulk solution via the border film to the outer surface of the CS materials. Second, intraparticle diffusion facilitates the diffusion of the radioactive ions from the outer surface to the inner surface of CS moieties by crossing the CS pores. Finally, the process is continued with the biosorption of radio-ions on the binding sites of CS until it reaches equilibrium [16].

Nuclear energy has become increasingly important in the context of sustainable technological growth and increased energy demand, and it is a wise alternative due to its low greenhouse gas emissions. However, the radioactive waste generated is a problem and must be handled properly. The occurrence of radioactive cesium and cobalt in a waste stream is a main concern since they are highly radiotoxic elements. Hence, the enhancement of economical and effective materials for the removal of these radionuclides to be ready for the subsequent immobilization process is of great significance. In this study, the authors offer CS as a biosorbent that can effectively bioremove these two radionuclides.

The purpose of this study is to analyze and investigate the factors that can affect the biosportive efficiency of CS for both Cs-137 and Co-60, such as contact time, waste solution pH, process temperature, and initial radioactivity content.

2. Experimental Approach

2.1. Materials

CS, provided by Sigma Reagent Co., Ltd., Saint Louis, MO, USA, is a hydrophilic polymer that sorbs water at ambient temperature. It was used as a biosorbent for the elimination of cesium-137 and/or cobalt-60 from spiked wastewater. All other chemicals used in the study were of analytical purity and were used without further purification.

2.2. Characterization of the CS under Consideration

2.2.1. Elemental Analysis

The elemental analysis of CS (C, N and H) was carried out using an elemental analyzer (Flash Elemental Analyzer EA1112). The CS used in the study, based on elemental analysis, comprised nearly 42.59% of carbon (C) and 7.328% of hydrogen (H), in addition to 21.466% of nitrogen (N); see Figure 1. The C/N mass ratio can be applied to calculate the deacetylation degree (DD) percentage of the used CS [17].

It was reported by Ssekatawa et al. that low-to-moderate solubility figures for CS are attributed to a high percentage of amino acid groups, i.e., a high nitrogen percentage, and consequently a high DD% value [18]. However, in this study, the DD% value was ≈200%, which represented a substantially high value compared to the values published in similar work [18]. Therefore, the significantly low solubility of the used CS under consideration can be attributed to a high DD% value and foremost to low demineralization and sharing of the leftover acetyl groups (glucosamine and N-acetylglucosamine units) by the CS backbone, which is named as the configuration of deacetylation [19].

The physicochemical and functional properties of CS are influenced by its DD% and molecular weight. CS has numerous functional properties, but its large molecular weight is a primary feature that affects its solubility and limits its potential applications. CS’s compactness and thermal stability significantly increase as its molecular weight reduces [20]. Therefore, these mentioned characteristics can enable recommending the proposed CS as a preferable biosorbent for the treatment of some low- and intermediate-level radioactive waste streams.

2.2.2. Fourier Transform Infrared Spectroscopy (FT-IR)

Fourier transform infrared spectroscopy was used to find the most interesting functional groups of CS. The vacuum-dried samples were examined on an analar-grade potassium bromide (KBr) disk using FT-IR (Jasco FT-IR-6100E Fourier transform infrared spectrometer, Japan) in the range of 400–4000 cm⁻¹. The structural characterization of CS was categorized, and the data obtained are presented in Figure 2. The majority of the peaks were interrelated to the carbohydrate configuration. The broad and strong band near 3440 cm⁻¹ could be attributed to O-H and N-H stretching vibrations. The band at ≈2915 cm⁻¹ could be due to the C-H stretching vibration, while the stretching form of the acetylated amine was detected close to 1650 cm⁻¹. The absorbed band near 1323 cm⁻¹ was mostly attributed to the O=C-O groups’ stretching frequency, while the bands near 1150 cm⁻¹ and 1100 cm⁻¹ could be explained as an asymmetric C-O-C etheric bridge stretching vibration style and C-O group deformations, respectively. The peak near 1256 cm⁻¹ corresponded to the free primary amine (-NH₂) at the C₂ position of glucosamine. Comparable characterizations for extracting CS have been previously published [21].

2.2.3. Thermal Analysis

Thermogravimetric analysis (TGA) and derivative thermogravimetric (DTG) analysis for CS, applied for the biosorption process, were carried out under a nitrogen ambience at a pure nitrogen gas flow rate of 50 mL/min, at a heating rate of 10 °C/min, and from room temperature up to 800 °C using Perkin Elmer Pyris 6.

Figure 3 shows the TGA and DTG analysis curves of CS. The thermal degradation of CS was achieved mainly through two steps. The first one was perceived at a temperature lower than 190 °C and was ascribed to the vaporization of any H₂O molecules that were trapped in between the CS polymer chains with mass loss 2.789%. The second thermal CS processing occurred in the temperature range of 149.56–866 °C, corresponding to about 48% mass loss, and could be attributed to the amine unit decomposition and the thermal degradation of the CS backbone, including dehydration, deacetylation, and depolymerization of saccharide units, in addition to the rapid breakdown and carbonization of the CS chain. The total mass loss at the end of degradation was about 51.417%; consequently, the remained char weighed about 48% at 800 °C, and similar explanations have been previously published [22].

The distribution of the decomposition stages of CS and temperature ranges are given in

Table 1. TGA and DTG data for the CS biosorbent.

Biosorbent	First Stage				Second Stage				Remain Chart after 800 °C
	Onset, °C	Offset, °C	Peak Max., °C	Mass Loss, %	Onset, °C	Offset, °C	Peak Max., °C	Mass Loss, %
CS	43.65	129	88.8	−4.5	276	341	307	−24	≈48%

Onset temperature in °C; offset temperature in °C; peak maximum temperature in °C.

.

According to the thermal analysis of data obtained, the CS polymer under consideration had acceptable thermal stability, indicating it as a suitable biosorbent at temperatures up to 150 °C. The remarkable heat stability of CS may be owing to intramolecular and intermolecular hydrogen bonding [23].

2.2.4. Scanning Electron Microscope (SEM)

The morphological architecture of CS was examined by screening the surface application of a tiny amount under a scanning electron microscope (JEOL-JXA-840A, Tokyo, Japan) operated at 20 Kv after covering with a thin layer of gold. The dry pieces of the solid sample were subjected to gold plating in S150A Sputter Coater Edwards (UK).

The micrograph of the purchased CS is revealed in Figure 4. It is apparent from the figure that the product was characterized by the following:

-

Homogeneous spreading of CS particles that reflects the distinguishing nature of CS.

-

Great porous construction with extra flat pores (dark spots).

-

The average pore size varied from 243.2 to 708.2 nm. A near-round structure for pores was observed, which provides an extra surface area for effective biosorption.

-

Repeated compact arrangements were found.

Based on the results of the SEM inspection, it is possible to conclude that CS has appropriate particle homogeneity, making it a viable biosorbent for the majority of radiocontaminants in spiked wastewater.

2.3. Biosorption Experiments

Spiked carrier-free stock solutions of Cs-137 or Co-60 and a mixture of both radionuclides were prepared by mixing them with known activities in distilled water. Cs-137 was purchased from Amersham, while Co-60 was obtained from CoCl₂ through activation in the second Egyptian research reactor. These solutions were kept as stock radioactive solutions for all experiments in the study, unless otherwise indicated. Alternatives of the biosorption parameters were used to evaluate the optimal sorption conditions, i.e., pH (acidic, neutral, and basic), contact time (30–10,080 min), initial radioactivity contents (50–100 Bq/mL), and sorbent dose (10–1000 mg). At each time interval, 1 mL of the solution was withdrawn, counted with a NaI scintillation detector Canberra Model 802 (3″ × 3″ crystal size), and returned to the treatment container again.

In the study, all biosorption activities were carried out in the laboratory using a batch static method by equilibrating a predetermined mass of CS in a defined volume of spiking radioactive waste solution mimic. This solution was labeled by adding a specific concentration of carrier-free radiocesium (Cs-137, T_½ = 30.5 years) or radiocobalt (Co-60, T_½ = 5.25 years) or a mixture of both.

The capacity of the biosorbent to uptake the radiopollutants was computed according to the following relation:

Biosorption efficiency percentage = ((a_o − a_t)/ a_o) × 100

where a_o = the initial activity in the spiked solution (Bq/mL),

a_t = the activity content in the solution at equilibrium time (t) (Bq/mL).

However, the capacity of the biosorbent (Bq/g) for removal of radionuclides can be determined using the following relation:

Biosorption capacity in Bq/mg = (a_o − a_t) × V/m

where V = the volume of the spiked solution under treatment (mL),

m = the mass of the biosorbent used (mg).

The factors that can affect the biosorption of Cs-137 and/or Co-60 from the spiked waste stream simulate, as stated before, were studied systematically.

3. Results and Discussion

3.1. Biosorption of Cs-137 and/or Co-60 from Spiked Wastewater with CS

The course of the biosorption of radiocontaminants from their waste streams can be explained through three mechanisms that can happen in parallel or consecutively. These processes are as follows: first, the physical biosorption process, which is due to molecular interaction forces between the sorbent and the sorbate; second, chemical sorption reactions, which can be explained on the basis of chemical bonds; and third, exchange sorption reactions, which proceed through electrolytic attraction between ions of opposite charges.

3.2. Consequence Chemical Mechanism of CS Biosorption Capability

The chemically active groups on the CS molecules were assumed to be associated with C₃-OH, C₆-OH, and C₂-NH₂, in addition to acetyl amino and glycoside bonds.

The latter two categories are quite stable, which means that they are difficult to fracture. Furthermore, because C₃-OH is differentiated as a secondary hydroxyl, it cannot spin freely, and its steric difficulty is so large that it is difficult to react. As a result, the chemically reactive capabilities of C₆-OH and C₂-NH₂ in CS molecules provide an advantage for those groups to launch additional groups via various forms of molecular derivations [24].

However, the -NH₂ group is more active than the C₆-OH group in CS molecules, which is attributed to the fact that the lone pair of electrons can be provided first by nitrogen of the amine group and then by oxygen of the hydroxyl group [25]. Thus, and according to Wang et al., in CS, the C₆-OH group does not react until the all C₂-NH₂ groups totally react [24]. According to several published studies, the amine group is the principal reactive site for metal ion sorption. Furthermore, hydroxyl groups can aid in the binding of metal ions [26]. However, due to the fact that some of these amine sites interact with hydrogen ions at lower pH, only a fraction of the free amino groups is available for metal binding.

The results shown in

Table 2. The biosorption efficiency percentages of Co-60 and Cs-137 from the spiked * streams using the CS biosorbent as a function of contact time.

Contact Time, h	Biosorption Efficiency, %
			Mixture *
	Co-60	Cs-137	Co-60	Cs-137
1/2	17.62	93.63	93.93	54.17
1	32.95	94.04	95.83	57.74
2	59.77	94.24	97.53	62.50
3	88.51	94.35	98.40	66.67
4	90.04	92.28	98.70	69.05
24	90.42	92.57	98.77	70.24
48	92.91	92.72	98.87	70.60
72	93.30	92.83	98.97	71.13
96	93.87	92.96	99.07	71.54
120	93.06	93.07	99.17	72.62
144	94.25	93.17	99.23	72.92
168	99.52	93.26	99.3	73.21

Mass of the biosorbent = 0.5 g; volume of the waste stream = 50 mL; pH value of the biosorption process ≈ 6.5; * a mixture of both Co-60 and Cs-137 was used to spike the solution.

indicate that the biosorption of Co-60 and/or Cs-137 increased with the escalating contact time until equilibrium states were assumed to be acquired. The biosorption process of radiocobalt seemed to take place in two distinct phases: it was rapid at first and then became nearly steady as the contact time rose, up to the achievement of equilibrium. The quick sorption phase of Co-60 out of the spiked radioactive waste solution simulate can be explained by the biosorption exchange process [24]. The initial fast biouptake of the radionuclide by the sorbent surface can be ascribed to the fact that all the active sites on the surface of the biosorbent were free. Slowing down of the sorption process, by increasing the contact time, could be due to the reduction in the existing active sites. For radiocobalt, equilibrium was achieved after 48 h (Table 1). However, the bioremoval of Cs-137 acquired a maximum value (after 3 h), after which the desorption process could occur. This may be due to the extreme solubility of cesium, and consequently, a later back-release could dominate (Table 1 and Table 2). A similar trend for lower sorption of Cs-137 compared to Co-60 using chitin and CS has been reported [13]. However, the biosorption efficiency percentage of Co-60 was nearly unchanged, even after 168 h, demonstrating that virtually, no back-release from the complex formed between the radionuclide and CS occurred.

In general, 144 h is an adequate time for the removal of more than 98% of radiocobalt from its waste solution spiked with the two radionuclides. The biouptake effectiveness of the biosorbent for Co-60 and the competitor Cs-137 was studied for a 50 mL solution of the waste simulate spiked with each radio contaminate using 0.5 g of CS at pH = 6.5 ± 0.1 and at room temperature (25 ±5 °C). Radiocobalt presented an adverse impact relative to the biosorption of radiocesium under the settled stated experimental conditions (Table 1).

Alternatively, it has been published that a distinct biosorbent has an affinity for some metal ions in preference to others [27].

The CS under investigation had a greater selective bioremoval affinity for Co-60 than for Cs-137. This is often referred to as the difference in the ionic radii of cobalt (192 pc) as contrasted with the higher ionic radii of cesium (265 pc), which helps its biosorption on accessible active sites. Furthermore, because cesium is a highly soluble cation, it is expected that biosorption and desorption processes would occur concurrently. However, more research is needed for grafting new functional groups with a CS backbone to enhance the biosorption of radiocesium, as well, from its radioactive liquid waste streams.

Bench static experiments were carried out to assess the parameters that are known to affect the biosorption of Co-60 and Cs-137 from their binary radioactive waste solution simulate, and these included the radionuclide species, contact time, the pH value of the medium, the biosorbent dose, the activity content, and the temperature of the treated medium.

3.3. Effect of Radio-ion Species

The biosorption process of metal ions can be affected by the existence of other species in the treated solution [28].

The sorption capacity of Cs-137 and/or Co-60 species by electrostatic interaction of functional groups on CS was greatly impacted by other competing factors and anion species present in the solution. Based on their physicochemical features, the following mechanisms for ion affinity with CS were proposed: (1) ions with higher valence bind with higher affinity to CS polymers, (2) ions with higher polarizability bind better with sorbents, and (3) ions with higher ionic charge/ionic radius ratios have higher binding affinity to sorbents.

Figure 5 presents the behavior of the biosorption from the 50 mL waste solution simulate that had pH 5–6 and was spiked with Co-60 based on 0.5 g of CS as a biosorbent.

The biosorption progression of Co-60 appeared to happen in two distinctive phases: it was rapid at first up to 3 h and then was recorded to be nearly steady by increasing the contact time, presenting the attainment of equilibrium. The rapid biosorption stage of Co-60 out of the spiked waste stream simulate can be explained by the biosorption exchange process [15]. The bioremoval of radiocobalt through the sorbent surface was fast at first; this can be attributed to the fact that all the active pores on the surface of the CS were free. Slowing down of the sorption process, by escalating the contact time, could be due to the reduction in the vacant active sites, as previously explained. For radiocobalt, equilibrium was achieved at more than 92% after 48 h.

Figure 6 shows that the behavior of the biosorption from the 50 mL waste solution simulate had pH 5–6 and was spiked with Cs-137 based on 0.5 g of CS as a biosorbent.

Alternatively, the biosorption of Cs-137 attained a ≈94% maximum value (after 3 h), after which the desorption process could occur, which can be due to the high solubility of cesium, as previously stated, and consequently, a later back-release could predominate. A similar trend for the final lower biosorption of Cs-137 relative to Co-60 using chitin and CS has been published [16]. Moreover, the biosorption efficiency percentage of Co-60 was nearly unaffected, even after 168 h, demonstrating that almost no back-release from the complex formed between the radionuclide and the CS occurred. However, 168 h was sufficient time for the removal of more than 93% of the radiocesium from the spiked waste streams (Figure 6).

3.4. Influence of Interaction Time

It is critical to determine the time dependence of the process in equilibrium in any investigated sorption process. Figure 7 depicts the effect of contact time on the efficiency of biosorption of both radionuclides from a binary system of the 50 mL water spiked with Cs-137 and Co-60 using 0.5 g of CS for a time array of 30–1080 min. It is clear from the figure that 1440 min was enough to reach equilibrium for a biosorption of more than 96% from the total activity added. The biosorption was quick at the initial stage for both radio-ions tested. After that, the sorption process showed a reduction in speed until equilibrium was acquired as the active pores bound with radio-ions. However, longer contact times, after 168 h, can lead to back-release of the sorbed ions [29]. This is most likely due to the high solubility of radiocesium as well as the protonation of the amine groups, which create an electrical repulsion between the radioelements in the spiked wastewater and the amine groups of CS.

3.5. Effect of pH of the Treated Spiked Solution

CS is a weak base, and its dissociation equilibrium is described by:

$$CS-{NH}_3^{+}+H_{2}O\leftrightarrow CS-{NH}_2+H_3{O}^{+}$$

At pH values around neutral, radiocation biosorption might be caused by the free electron doublet of nitrogen on amine groups. In addition, the speciation and dispersal of radiocation species in the waste stream are principal factors that determine the impact of pH on the biosorption capacity. Speciation at a certain pH value can impact the biosorption operation, that is, the existence of species that cannot combine with the biosorbent. Near neutral pH, the radiocations Cs (I) and Co (II) exist as free ions in the aqueous stream and they can be biosorbed onto CS by chelation according to a previously published equation [30]:

$$CS-{NH}_2+M^{2+}\leftrightarrow CS-{NH}_2-M^{2+}$$

When the pH increases, the surface of the beads becomes negatively charged due to the deprotonation process [26]. The consequence is a decrease in the repulsive force between the amine groups of CS and the metal ions in the solution, which increases the removal of metal ions from the solution until an ideal value is obtained. The removal of metal ions from the solution decreases when the pH rises over the optimal range, because insoluble metal hydroxide starts to precipitate out of the solution.

The free electron pairs of nitrogen of the amine groups and the vacant orbitals of the ions produce coordination bonds that allow the amine groups to establish bonds with radio-ions, radiocobalt, and/or radiocesium. Moreover, the biosorption of both cations at pH values close to neutrality and at lower pH values, where the amine group is protonated, may be caused by the free electron doublet of nitrogen of the amine groups. At these pH values, the polymer acquires cationic groups that can bind anions through electrostatic interactions [31].

As previously stated, the radionuclide sorption capability of active amino groups can be notably affected by the pH of the treated stream simulate, through which it can ameliorate the surface charge of CS and its ionization extent [31]. It also influences the degree and manner of reaction assumed between sorbate and sorbents. The biosorption of both Cs-137 (I) and Co-60 (II) out of the 50 mL waste stream simulate binary system on 0.5 g of CS at different three pH ranges was performed to reveal the impact of the solution pH on the biosorption efficiency, and the data obtained are illustrated in Figure 8. It is worthy of notice that the nominated pH values were adjusted using dilute hydrochloric acid or dilute sodium hydroxide solution. It is apparent from the figure that the protonation of the amine groups caused electrical repulsion between the radio-ions and the amine groups of CS, which is mostly responsible for the drop in the biosorption efficiency percentages at acidic pH levels. Because of the deprotonation outcome, the CS surface showed a negative charge when the pH value increased [32]. Accordingly, the repulsive force that occurs between the radionuclides and the amine groups reduced, enhancing the removal of the radio-ions up to an optimal value. At alkaline pH, more than the optimal one, insoluble cobalt hydroxide started to precipitate, i.e., past pH 7, the escalation in the biosorption percentage can be attributed to the precipitation of radiocobalt as cobalt hydroxide (Co(OH)₂), more easily than the sorption of Co-60 by CS. The demarcation between the value of radiocobalt sorbed and the precipitated amount is hard. The same result has been reported by other investigators [33,34]. However, and according to Chadha et al., below pH 3, i.e., an acidic medium, CS can be dissolved [35].

3.6. Effect of Treatment Temperature

A key component of biosorption treatment is temperature, which has a positive correlation with the kinetic energy of the sorbent in spiked wastewater [36]. The impact of temperature on biosorption efficiency was performed for the 50 mL binary spiked solution, which had 0.5 g of CS biosorbent at both ambient temperatures of 30 ± 5 °C and 9 ± 1 °C, and the data obtained are presented in Figure 9.

It is obvious that at room temperature, the biosorption efficiency percentage of the total activity was greater than that at ~9 °C. This can be due to the fact that as the temperature increases, the kinetic energy also increases, consequently enhancing the movement of radionuclides onto the sorbent pores and shortening the time of attaining sorption equilibrium [37,38]. Furthermore, it is assumed that the process of ion diffusion into the CS pores is an endothermic reaction and that the interaction of amino groups with radiocations is an exothermic reaction. It is also expected that the positive enthalpy alteration will be larger than the negative enthalpy alteration caused by the complexes that form in solution between the spiking ions and the amino group, resulting in a total positive enthalpy alteration [36]. The square of the regression coefficient, R², was calculated to be near 0.7, indicating that the practicability of the biosorption mechanism is a factor of the treated solution temperature, so an escalation in temperature, up to certain limit, is essential for efficient sorption treatment. Even the applied CS is, to a great extent, thermally stable, yet the uncontrolled rise in temperature can deteriorate CS stability; therefore, the maximum temperature for the treatment process should be signified.

3.7. Effect of Biosorbent Dosage

One important consideration for figuring out the ideal amount or saturation peak is the dose of the biosorbent. At this peak, any additional mass of the biosorbent will not add to any detectable enhancement in the biosorption progress [39].

The biosorption processes were carried out near neutral pH, i.e., at pH ≈ 6.5, and at ambient temperature (i.e., 30 ± 5 °C) for the waste solution simulate binary system spiked with both Cs-137 and Co-60. After 168 h, it was obvious that the biosorption efficiency of radioactivity was enhanced by escalating the CS dosage. The total maximum peak was recorded at 0.5 g CS dosage. After this peak, the biosorption efficiency percentages declined with further escalation of the biosorbent doses, keeping the volume of the treated spiked solutions constant at 50 mL and the exposed reaction area nearly fixed (Figure 10). After the peak point, as the CS masses increased, the numbers of available binding sites for the radiocontaminants decreased. This reduction can be due to the overlapping of the biosorbent layers over each other or the agglomeration of too many biosorbent particles, and consequently, the active sites available for the sorption of radioactive ions were reduced or blocked [40]; similar results have been obtained in various published studies [41,42,43,44,45]. Therefore, it can be concluded that a 0.5 g CS dose for the 50 mL solution under treatment is considered the optimal biosorbent dosage under the stated conditions.

3.8. Impact of Initial Radioactivity Content in the Treated Solution

The impact of the initial total radioactivity content (Cs-137 and Co-60) of the 50 mL treated spiked solution, pH ~ 6.5, keeping the sorbent dose constant at 0.5 g was studied, and the data obtained are presented in Figure 11.

The feasibility and efficiency of any biosorption process are not only related to a function of the properties of the biosorbent, but they also depend on the radicontaminants’ concentrations in the waste stream. The effect of the total initial radiocobalt and radiocesium contents was studied at room temperature. The total biouptake percentages showed slight increases with increasing total radioactivity up to a total activity of 19,800 Bq. Increasing the total activity up to 29,460 Bq/50 mL was accompanied with a reduction in the biosorption efficiency percentages to 75%. An increase in the initial radioactivity content implies that more radioactivity is present in the solution, which causes more ions to be linked to the same amount of CS and, ultimately, saturation of the sorbent pores. This can be explained by the fact that at lower initial radioactivity concentrations, the ratio of radio-ions to the CS mass is low. It is worthy of notice that even the biosorption efficiency of CS decreased at the excessive total radioactivity added, yet it seemed to be constant after a long contact time. This could be attributed, as previously stated, to the saturation of the CS surface, where radio-ions bind to nearly all the binding sites. Accordingly, too much of the added radicontaminants was not sorbed and remained in the bulk of the treated solution.

3.9. Biosorption Kinetics

Film diffusion, intraparticle diffusion, and biosorption are the three stages generally expected to occur in the biosorption process [46]. Initially, film diffusion facilitates the external mass transfer of radioactive ions from the bulk solution via the border film to the outer surface of the CS materials. Second, intraparticle diffusion facilitates the migration of radioactive ions from the outer surface to the interior surface of CS moieties via CS pores. Finally, the process is maintained by the biosorption of radio-ions on the binding sites of CS until equilibrium is reached [47].

The slope and intercept of the graph illustrating the biosorption capacity ($q_t$) versus the square root of contact time (t^0.5) may be used to calculate the values of the rate constant of intraparticle diffusion and the y-intercept [48]. Figure 12a describes the biosorption of radiocobalt, which included three parts of the linear line.

The first part is concerned with the influence of the boundary layer. The radiocobalt ions were transferred to CS’s outer surface at a high diffusion rate via the film. When the outside surface was saturated, the radioactive elements diffused through the pores to CS’s interior surface. This process occurred as a result of the intraparticle diffusion impact and was designated as the second component with a lesser inclination. Finally, the last half of the curve reflects the equilibrium sorption of radio-ions on the binding sites of CS. A similar trend can be described for the biosorption of radiocesium, and the data obtained are presented in Figure 12b. Comparable behaviors for the biosorption of different materials applying CS and its composites have been reported in the literature [49,50]. The model states that the straight line should meet the origin when intraparticle diffusion is the rate-describing phase in the biosorption route for radioactive components. Nonetheless, none of the model lines for the investigated CS passed through the origin. When the y-intercept value is higher, it suggests that the biosorption route is driven not just by intraparticle diffusion, but also by a constant degree of the boundary layer [51]. Since the curves do not pass the origin and have some sort of intercepts (Figure 12a,b), this can suggest the presence of a significant boundary layer. Consequently, this ascertains that the kinetics of the overall sorption mechanism is dependent on the transfer of radio-ions from the bulk solution to the solid-phase boundary. Briefly, the larger the magnitudes of the y-intercept, the greater the impact of the boundary layer.

According to Figure 12a,b, the square of the regression coefficients, R², were 0.99, 0.96, and 0.997 for radiocobalt sorption and 0.959, 0.995, and 0.999 for radiocesium biosorption. These values indicate that the biosorption of both radionuclides happened through the three mentioned steps and more likely through the last step i.e., the binding sites of CS.

4. Conclusions

A new potential biosorbent was developed for sustainable accessibility of pairs of radio-ions to the binding site of the CS polymer. A comprehensive study was carried out in order to retrieve the impact of adsorption parameters on the binding of cobalt-60 and/or cesium-137 onto a CS biosorbent. The biosorption efficiency of both radioelements was a function of the pH of the spiked waste solution, contact time, adsorbent dosage, initial radioactivity added, treatment temperature, and others. The result obtained indicate that the sorbent under consideration can be recommended, i.e., CS is efficient for the bioremoval of Co-60 and/or Cs-137 (more than 94% of radiocobalt and about 93% of radiocesium). This process can be primarily regulated by chemisorption, electrostatic attraction, ion exchange, or other chemical or physical processes. Furthermore, this new sorbent with a high amino content may be a good low-cost option for the treatment of several low- and medium-level radioactive waste streams created by peaceful application of nuclear technology in our daily lives. However, additional research is required to enhance the adsorption characteristics of CS by physical and/or chemical changes.