The Transformation and Leachability of Fly Ash/Cement Waste Forms Subjected to the Simultaneous Effect of Heat and Chemistry

Zheng, Zhao; Wen, Hua; Li, Yuxiang; Qin, Min; Wang, Yao

doi:https://doi.org/10.1155/2022/7983231

Advances in Materials Science and Engineering

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Abstract Introduction Materials Results and Discussion Conclusion Data Availability Conflicts of Interest Authors’ Contributions Acknowledgments References Copyright Related Articles

Research Article | Open Access

Volume 2022 | Article ID 7983231 | https://doi.org/10.1155/2022/7983231

The Transformation and Leachability of Fly Ash/Cement Waste Forms Subjected to the Simultaneous Effect of Heat and Chemistry

Zhao Zheng,^1,2Hua Wen ,^1,2Yuxiang Li,³Min Qin,^1,2and Yao Wang^1,2

Academic Editor: Shengwen Tang

Received03 Sept 2022

Revised22 Sept 2022

Accepted28 Sept 2022

Published12 Oct 2022

Abstract

The utilization of cement waste forms with high content of fly ash is a potential method in large-volume immobilization of low- and intermedium-level radioactive waste at near-surface. However, the migration behavior of radionuclides in fly ash/cement waste forms under geological environment is still unclear. This study researched the transformation of products, pore structure as well as the leachability of Ca²⁺, Na⁺, and simulated radioactive nuclides (Sr²⁺ and Cs⁺) in fly ash/cement waste forms under simultaneous effect through X-ray diffractometer (XRD), mercury intrusion porosimetry (MIP), and inductively coupled plasma optical emission spectroscopy (ICP-OES). Furthermore, the relationship between the decalcification and the leaching of Sr²⁺ and Cs⁺ was established by the fitting curves of apparent diffusion coefficient. The results indicated that the products of fly ash/cement waste forms were kept stable under simultaneous effect and the pore structure has excellent resistance to temperature. The increasing salt concentrations of leachant and temperature promoted the leaching of radionuclides while the protection layer formed at 60–80°C inhibited this phenomenon. The relationship of apparent diffusion coefficient between Sr²⁺ and Ca²⁺ was quadratic nonlinear, while the relationship between Cs⁺ and Ca²⁺ showed a linear relationship. Compared to cement waste forms, fly ash/cement waste forms were more stable. In addition, the long-term retention ability of fly ash/cement waste forms to Cs⁺ was more outstanding than Sr²⁺.

1. Introduction

According to the national 14th five-year plan of China, the country’s installed nuclear power capacity in operation will reach 70 million kilowatts by 2025. A large amount of radioactive waste liquid is produced during the operation of nuclear reactors, reprocessing of spent fuel, and decommissioning of nuclear facilities. Indeed, the volume of low- and intermedium-level radioactive waste (LILRW) liquid accounts for more than 95%. The large-volume immobilization at near-surface is a preferred method as it can improve the immobilization efficiency of LILRW and simplify the process. In order to reduce the central temperature and improve the thermal stability of the large-volume waste forms, it is necessary to add mineral admixtures, such as fly ash, metakaolin, and slag with a high dosage (up to 85%). After reaction, such mineral admixtures will transform into geopolymers [1, 2]. Besides, the fly ash-based geopolymer is regarded as a new potential solidified material for LILRW due to the characteristics of environmental protection, strong acid resistance, low permeability, and drying shrinkage [3–6], especially the excellent selective absorption and retention ability on simulated nuclides [7, 8].

In despite of the multiple protective barriers placed around the large-volume waste forms, the mass and heat transfer still occurred between the waste forms and the environment after near-surface disposal under real service conditions [9]. Due to the complexity of the underground environment (thermal-hydrological-mechanical-chemical), the combined action of these factors will lead to the degradation of the waste forms, including the dissolution and decalcification which would result in the loose pore structure and weaken the microhardness of waste forms [10–12].

Iron compounds (amorphous iron, hematite, magnetite, etc.) in geopolymer do not participate in geopolymer reaction basically [13], but these substances will dissolve in acidic solution and produce pores. Indeed, Al³⁺ has a higher solubility under acidic conditions while the solubility of Si⁴⁺ is higher under alkaline conditions [13]. Temuujin [14] studied and found that geopolymer has great solubility in strong acid solution and strong alkali solution. Nguyen and Škvára [15] found that the decomposition of Si–O–Si bond would stop due to the rapid leaching of Na⁺ and the occurrence of neutralization reaction in HCl solution with pH = 1, while Al–O–Si bond was more stable than Si–O–Si bond in NaOH solution with high pH. Bakharev [3] studied the durability of fly ash-based geopolymer in 5% acetic acid and sulfuric acid, and found that the degradation of geopolymer was related to depolymerization, dealuminization, replacement of Na⁺ and K⁺, including the formation of zeolite.

Due to the harsh external environment, the degradation of the waste forms occurred and the simulated nuclides eventually accelerated to migration outside [16]. Ofer-Rozovsky et al. [17] discovered the selectivity of geopolymers to Cs⁺ by leaching experiments, and the degree of selectivity was obtained by calculating the ratio of leaching rates of Na⁺ and Cs⁺ at different leaching periods. For the leaching characteristics of simulated nuclides, the International Atomic Energy Agency compared the apparent diffusion coefficient (ADC) with the leaching data and established the relationship between them [18]. Abdel Rahman et al. [19] conducts the mathematical analysis of leaching data in long period and found that Sr leaching in cement waste forms was the result of the combined action of first-order reaction model and diffusion mechanism. Current studies on leaching kinetics mainly focus on first-order reaction model (FRM), first-order reaction-diffusion model (FRDM), and first-order reaction-diffusion-dissolution model (FRDDIM) [20]. Sami et al. [21] used the above kinetic models to fit the leaching data and obtained a nonlinear regression relationship, which indicated that the leaching mechanism of simulated nuclide Sr in synthetic zeolite-rich cement followed the order of first-order kinetic reaction and diffusion reaction. Xu et al. [22] found that Sr and Cs followed different kinetic models in the leaching process by adding first-order reaction-dissolution model (FRDIM) and dissolution-diffusion kinetic model (DIDM) on the basis of Sami et al. [21]. Therefore, the migration of simulated nuclides from the interior of the waste forms to the external environment conforms to the dynamic characteristics and is dominated by the corresponding migration mechanism.

However, the migration behavior of radionuclides in fly ash/cement waste forms in geological environment is not clear, and the relationship between the degradation of fly ash/cement waste forms and migration of radionuclides has not been established. Especially, it is lack of the scientific evaluation method of long-term stability and safety of radionuclides in complex geological environment. Thus, this study simulated the different geological environment by changing the leaching temperatures and the concentrations of acid leaching solution. The transformation of products and pore structure as well as the leachability of main cations (Ca²⁺ and Na⁺) and simulated radioactive nuclides (Sr²⁺ and Cs⁺) were researched. Then, the time required for complete leaching of Sr²⁺ and Cs⁺ was calculated and deduced after the curves fitting of ADC. Also, the relationship between the decalcification of fly ash/cement waste forms and the leaching of Sr²⁺ and Cs⁺ was unveiled, which is contributed to the safety evaluation of radioactive nuclides under geological environment.

2. Materials and Experiments

2.1. Materials and Preparation

The low-Ca fly ash and Portland cement (P·O 42.5, China standard: GB175-2007) were used in this study. Table 1 shows the corresponding composition of oxides in the two kinds of raw powder materials, while the particle size and morphology can be seen in reference [23]. The water used in this study was ultrapure water in order to eliminate the influence of impurities on the leaching data. Meanwhile, the inorganic salts, such NaNO₃, NaOH, Na₂CO₃, and NaCl, are all analytically pure.

The LILRW of 300 g/L was prepared through solving the inorganic salts one by one. Table 2 shows the concentrations of the corresponding salts. The mass ratio of total salts occupied in LILRW was 25.2%.

The fly ash/cement waste forms were prepared by mixing the raw powder materials with the LILRW under the water/binder ratio of 0.4. After stirring for 5–7 min, the fresh slurry was subjected to steam curing under 90°C for 4 h. Then, the sample was demolded and continued to be cured under 90°C for 28 d.

2.2. Leaching Experiment

Leaching experiment was conducted following with the Chinese standard GBT7023-2011 [24]. The samples were fastened with nylon thread and hung in a polyethylene bottle (volume = 250 ml) with 240 ml of NH₄Cl solution. The concentration of NH₄Cl solution ranged from 0 to 5 M while the leaching temperature was between 20 and 80°C. The Cumulative Fraction Leached (CFL) of cations in fly ash/cement waste forms were calculated through the equation shown below:where a_n is the leached mass of cations in leachate at different cumulative leaching times (), A₀ is the initial mass of cations in samples () before leaching, S is the surface area of the samples (cm²), and V is the volume of the samples (cm³). t is the cumulative leaching times and (Δt)_n is the leaching internal (d), (Δt)_n = (Δt)_n − (Δt)_n − 1.

The leaching period was from 1 d to 180 d. At each leaching period, the leaching solution was refreshed. The leaching solution was ultrapure water and NH₄Cl solution with different concentrations while the leaching temperatures ranged from 20 to 80°C. After the leaching experiment, the samples were crushed. Some of the fragment samples were kept in alcohol directly, and the others were ground into powders after being mixed with alcohol. Before the tests, all the samples were dried at 45°C for 24 h.

2.3. Characterization

The oxides composition of fly ash and Portland cement were determined by X-ray fluorescence (Axios, PANalytical, Netherlands) after drying to constant weight. The crystalline phases of fly ash/cement waste forms were tested by X-ray diffractometer (DMAX1400, Rigaku, Japan) from 3 to 80° of 2θ at a scanning rate of 8°/min. The pore structure of fly ash/cement waste forms with a diameter between 4 and 8 mm was determined through mercury intrusion porosimetry (Micromeritic AutoPore IV9500, USA) while the intrusion pressure ranged from 0 to 33000 PSI. The concentrations of leached cations in the leaching solution were detected through the inductively coupled plasma optical emission spectroscopy (iCP-OES 7000, Thermo Fisher Scientific, USA). Before the testing, the concentrations of Cs⁺ in leaching solution were diluted into 0–10 ppm and KCl was added in the leaching solution as a shielding agent for ionization. The concentrations of Sr²⁺ and other cations were diluted into 0–30 ppm. Indeed, the pH of all the leaching solution was adjusted to about 7.

3. Results and Discussion

3.1. Transformation of Products and Pore Structures

It can be seen from Figure 1(a) that the main crystal phase in the cement waste forms of simulated LILRW is not different from the waste forms with pure NaNO₃ solution in previous study [24], and the crystalline phases are Ca(OH)₂, CaCO₃, and NaNO₃ as well as the residual C₂S. As in the simulated LILRW, the main component is still NaNO₃. Indeed, the content of other sodium salts is rather low, which does not appear in the XRD pattern and does not affect the types of crystal products phase in the cement waste forms. As seen from Figure 1(b), the main crystal phase obtained from the fly ash/cement waste forms mixed with simulated LILRW is not different from the waste forms with pure NaNO₃ solution under low content of NaOH in previous study [23]. The main crystal phase is mainly silica, mullite, and NaNO₃. However, under the condition of high NaOH, the phase was slightly different from the waste forms with NaNO₃ solution, and only the chabazite phase was generated but there was no Na–P1 zeolite which existed in the condition of NaNO₃ solution. Combined with previous studies about the effect of NaOH contents on the product and structure of fly ash/cement waste forms and the leaching rate of simulated radionuclides [23, 25], the NaOH content in the fly ash/cement waste forms of simulated LILRW in the study below was determined to be 10%.

(a)

(b)

As seen in Figure 2, the crystalline products in the fly ash/cement waste forms of simulated LILRW did not change after leaching in ultrapure water (0 M NH₄Cl). After leaching in 0.6 M NH₄Cl solution, the diffraction peak of NaNO₃ disappeared in the matrices compared to Figure 1(b). After leaching in 5 M NH₄Cl solution, the diffraction peak of NaNO₃ also disappeared, but the diffraction peak of NH₄Cl appeared. In all the leached samples, the crystalline phase, including zeolite phases in the fly ash/cement waste forms remain unchanged under the simultaneous effect, while the main change is in the C–S–H gel which suffered the decalcification. According to previous studies, the structure and composition of cement hydration products, especially C–S–H gel, were significantly degraded under the accelerated decalcification [26–28], which resulted in the loose pore structure of waste forms and the significant increase of leaching rate of simulated radionuclides.

As seen in Figure 3 and Table 3, the median pore diameter of cement waste forms increases obviously with the increasing concentrations of leaching solution except 5 M NH₄Cl. The porosity increases firstly and then decreases with the concentration of leaching solution, and the highest value of 20.99% is obtained at the NH₄Cl concentration of 0.6 M. The median pore diameter increases with the leaching temperature ranging from 20 to 80°C while the highest porosity of 10.53% is obtained at 60°C.

(a)

(b)

As seen in Figure 4 and Table 4, the median pore diameter of fly ash/cement waste forms also increases obviously when the leaching solution ranges from ultrapure water to 3 M NH₄Cl solution, which is similar with the change of cement waste forms. The highest porosity of 20.79% is also obtained at the NH₄Cl concentration of 0.6 M and then decreases with the increasing NH₄Cl concentration. However, the median pore diameter and porosity show little change when the leaching temperature ranges from 20 to 60°C. Though the porosity of 8.35% at 80°C is the highest at different leaching temperatures, it only increased by 1% compared with that at other leaching temperatures.

(a)

(b)

Based on the above results, it can be concluded that the products of fly ash/cement waste forms are more stable than cement waste forms under simultaneous effect. Also, the degradation of pore structure of fly ash/cement waste forms is mainly caused by the decalcification of hydration products of cement [29]. Moreover, the pore structure of fly ash/cement waste forms has excellent resistance to temperature than cement waste forms as the crystalline zeolite phases does not dissolute in such moderate condition.

3.2. Leachability

Figures 5–8 show the CFL of cations in fly ash/cement waste forms under simultaneous effect. As seen in Figure 5, the CFL of Na⁺ in ultrapure water is always much lower than that in NH₄Cl solution. However, with the increase of leaching temperature, the CFL of Na⁺ in ultrapure water is closer to that in NH₄Cl solution. At the leaching period of 180 d, the CFL of Na⁺ in ultrapure water is equivalent to that in 5 M solution at 80°C. In the early stage of leaching, the CFL of Na⁺ increases while the NH₄Cl concentration ranges from 0.6 to 5 M, but decreases in the later stage. Indeed, this transition becomes earlier at the higher leaching temperature, which occurred under 14 d at 20°C and 40°C, 7 d at 60°C, and 3 d at 80°C, respectively.

(a)

(b)

(c)

(d)

(a)

(b)

(c)

(d)

(a)

(b)

(c)

(d)

(a)

(b)

(c)

(d)

As seen in Figure 6, the CFL of Ca²⁺ in ultrapure water is less than 1.5 × 10⁻³ cm at various leaching temperatures. At the early stage of leaching, the CFL of Ca²⁺ increases following with the rising NH₄Cl concentration. However, the CFL of Ca²⁺ in 3 M and 5 M NH₄Cl solutions is less than 0.6 M and 1 M adversely at a specific leaching period. This specific leaching period is shortened when the leaching temperature ascends, which is 154 d at 20°C, 28 d at 40°C, 63 d at 60°C, and 21 d at 80°C, respectively. With the increase of leaching temperature, the disparity of the CFL of Ca²⁺ between 3 and 5 M NH₄Cl solution and 0.6–1 M becomes more and more obvious.

The CFL of Sr²⁺ in ultrapure water is always less than 6 × 10⁻⁴ cm under various leaching temperatures, as seen in Figure 7. But the CFL of Sr²⁺ in 0.6–3 M NH₄Cl solution increases significantly following with the rising NH₄Cl concentrations, and the gap expands with the extension of leaching period. The CFL of Sr²⁺ in 5 M NH₄Cl solution was lower than that in 3 M at 20°C and higher than that in 3 M at 40 and 60°C. Nevertheless, it is significantly lower than that in 3 M at 80°C after the leaching period of 154 d.

Compared with the leaching condition of NH₄Cl solution, the CFL of Cs⁺ under the condition of ultrapure water is always very low, just close to 1 × 10⁻⁵ cm, as seen in Figure 8. At each leaching temperature, the CFL of Cs⁺ increases when the concentration of leaching solution rises. However, at 40°C and 80°C, the CFL of Cs⁺ in 5 M and 3 M began to be lower than that in 0.6 M at 154 d and 28 d of leaching period, respectively.

From Figures 5–8, it can be concluded that the CFL of all the cations increase with the rising temperatures and concentrations which accelerated the degradation [30]. But the leaching of cations decrease at a specific leaching period, especially at high leaching temperature of 60 and 80°C. This is mainly attributed to the crystallization of NH₄Cl at the surface and inside of fly ash/cement waste forms, which led to a dense surface to restrain the leaching of cations [31]. Also, the high NH₄Cl concentration has decreased the porosity of fly ash/cement waste forms (Figure 4 and Table 4), which contributed to the retention of radioactive nuclides. Moreover, a transition layer was formed between the inner nondegraded layer and the outer fully degraded layer [32]. The existence of transition layer was also the main factor that led to the reduction of ion leaching rate at the later stage. So, it can be concluded that the dense structure formed at the rather high temperature (60–80°C) and high NH₄Cl (3 M and 5 M) is contributed to the stability of the fly ash/cement waste forms.

3.3. Fitting of Apparent Diffusion Coefficient

The fitting process was based on the equation of Fick’s second law in a semi-infinite medium, and apparent diffusion coefficient (ADC) as well as the time for complete leaching of radionuclides were calculated [21, 33, 34]. In this study, the ADC of cations in the fly ash/cement waste forms was assumed to be constant. In order to avoid the influence of the crystalline NH₄Cl on the calculation of ADC, the leaching data at the leaching period of 1–14 d were selected. As the fitting curves of other leaching temperatures of Sr²⁺ and the fitting curves of Ca²⁺ and Cs⁺ are similar to Figure 9, the rest of the fitting curves are not listed one by one. Based on the phase transformation and leachability of Na⁺, it can be deduced the leached Na⁺ was mainly free Na⁺ which has little influence on the leaching of radioactive nuclides. Thus, the fitting curves of Na⁺ have not been discussed below.

After fitting, the ADC and the time for complete leaching of Sr²⁺ in the fly ash/cement waste forms are listed in Table 5. The time for complete leaching of Sr²⁺ in fly ash waste forms at ultrapure water and 20°C was the longest (2.15 × 10⁸ d) while the shortest time was almost one month at 5 M of NH₄Cl solution and 80°C.

It can be seen from Figure 10 that both the ADC of Ca²⁺ and Sr²⁺ in fly ash/cement waste forms increase with the rising leaching temperatures and concentrations of leaching solution. As seen in Figure 11, the relationship of the ADC between Ca²⁺ and Sr²⁺ in fly ash/cement waste forms is quadratic nonlinear (Y = 4.44 × 10⁻⁵ − 0.21X + 1403.33X²), which is different from the linear relationship in cement waste forms [31]. When the ADC of Ca²⁺ is lower than 4 × 10⁻⁴ cm²/d, the ADC of Sr²⁺ and Ca²⁺ can be approximately regarded as a linear relationship. However, the ADC of Sr²⁺ increases exponentially with the rising ADC of Ca²⁺ when the ADC of Ca²⁺ is above 4 × 10⁻⁴ cm²/d.

(a)

(b)

The ADC of Cs⁺ (D_cs) and time for complete leaching (T_end) are listed at Table 6. The time for complete leaching of Cs⁺ in fly ash waste forms at ultrapure water and 20°C was the longest (5.15 × 10⁷ d) while the shortest time was almost two days at 5 M of NH₄Cl solution and 60°C. It can be seen from Figure 12 that the highest ADC of Cs⁺ is obtained at the leaching temperature of 60°C but not the highest temperature of 80°C. This trend further indicated that the high temperature (80°C) would indeed enhance the retention ability of fly ash/cement waste forms to radioactive nuclides. Indeed, the ADC of Cs⁺ increases when the concentrations of leaching solution ascend. As seen in Figure 13, the ADC of Cs⁺ increases obviously with the rising ADC of Ca²⁺, which shows a linear relationship as: Y = −1.84 × 10⁻¹⁰ + 6.86 × 10⁻⁶X.

Comparing the fitting curves in Figures 11 and 13, it can be found that the leaching rate of Cs⁺ was higher than Sr²⁺ at the early period of leaching experiment while it was opposite at the latter period. Due to the little ionic radius of Cs⁺ and the unstable physical adsorption on zeolites [35–37], the free and adsorbed Cs⁺ preferred to leach out in the solution environment all the time, which was not affected by the decalcification obviously. However, Sr²⁺ existed in fly ash/cement waste forms by chemical adsorption on C–S–H gels and partial Ca²⁺ in C–S–H gels was substituted by Sr²⁺ [38–41]. Thus, the decalcification occurred in the latter period can accelerate the leaching of Sr²⁺ sharply. Meanwhile, the decalcification led to the polymerization of C–S–H gels and the decrease of bridging oxygen, which further weakened the adsorption ability of C–S–H gels [24, 37].

4. Conclusion

Above all, it can be concluded that the degradation of fly ash/cement waste forms was still attributed to the decalcification of the hydration products of cement. Compared to cement waste forms, fly ash/cement waste forms were more stable under simultaneous effect. Furthermore, the long-term stability of Cs⁺ in fly ash/cement waste forms was better than Sr²⁺.(1)The products of fly ash/cement waste forms are kept stable under a simultaneous effect. The pore structure transformation of fly ash/cement waste forms under simultaneous effect were mainly attributed to the degradation of cement waste forms. Indeed, the pore structure has excellent resistance to temperature.(2)Following with the increase of the concentrations of leaching solution and leaching temperature, all the cations in fly ash/cement waste forms accelerated to migrate outside. But the migration of cations was restrained at a specific leaching period under high leaching temperatures (60–80°C) and NH₄Cl (3–5 M) because of the protection layer formed under these conditions.(3)The ADC of both Cs⁺ and Sr²⁺ increased with the increase of Ca²⁺, and the relationship of ADC between Sr²⁺ and Ca²⁺ in fly ash/cement waste forms was quadratic nonlinear: Y = 4.44 × 10⁻⁵ − 0.21X + 1403.33X². However, the relationship of ADC between Cs⁺ and Ca²⁺ showed a linear relationship: Y = −1.84 × 10⁻¹⁰ + 6.86 × 10⁻⁶X.

Data Availability

The data that support the findings of this study are available from the corresponding author upon reasonable request.

Conflicts of Interest

The authors declare that they have no conflicts of interest..

Authors’ Contributions

Methodology, formal analysis, investigation, writing-original draft were performed by Zhao Zheng. Resources and supervision were performed by Hua Wen. Yuxiang Li: conceptualization and resources. Resources, formal analysis, and writing-review and editing were performed by Min Qin. Validation and resources were performed by Yao Wang.

Acknowledgments

This work was supported by the Scientific Research Fund of Southwest University of Science and Technology (No. 21zx7124). The authors gratefully acknowledge School of Civil Engineering and Architecture, School of Materials Science and Engineering (Southwest University of Science and Technology) for providing facilities to carry out the work.

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Copyright © 2022 Zhao Zheng et al. This is an open access article distributed under the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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