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Characterization of radioactive contaminants and water treatment trials for the Taiwan Research Reactor's spent fuel pool
Journal of Hazardous Materials 233–234 (2012) 140–147
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Characterization of radioactive contaminants and water treatment trials for the Taiwan Research Reactor’s spent fuel pool Chun-Ping Huang ∗ , Tzung-Yi Lin, Ling-Huan Chiao, Hong-Bin Chen Institute of Nuclear Energy Research, 1000, Wenhua Road, Jiaan Village, Longtan Township, Taoyuan County 32546, Taiwan, ROC
h i g h l i g h t s
Deal with a practical radioactive contamination in Taiwan Research Reactor spent fuel pool water. Identify the properties of radioactive contaminants and performance test for water treatment materials. The radioactive solids were primary attributed by ruptured spent fuels, spent resins, and metal debris. The radioactive ions were major composed by uranium and fission products. Diatomite-based ceramic depth filter can simultaneously removal radioactive solids and ions.
a r t i c l e
i n f o
Article history: Received 26 September 2011 Received in revised form 8 June 2012 Accepted 2 July 2012 Available online 11 July 2012 Keywords: Radioactive contaminant Water treatment Spent fuel pool Taiwan Research Reactor
a b s t r a c t There were approximately 926 m3 of water contaminated by fission products and actinides in the Taiwan Research Reactor’s spent fuel pool (TRR SFP). The solid and ionic contaminants were thoroughly characterized using radiochemical analyses, scanning electron microscopy equipped with an energy dispersive spectrometer (SEM-EDS), and inductively coupled plasma optical emission spectrometry (ICP-OES) in this study. The sludge was made up of agglomerates contaminated by spent fuel particles. Suspended solids from spent ion-exchange resins interfered with the clarity of the water. In addition, the ionic radionuclides such as 137 Cs, 90 Sr, U, and ␣-emitters, present in the water were measured. Various filters and cation-exchange resins were employed for water treatment trials, and the results indicated that the solid and ionic contaminants could be effectively removed through the use of <0.9 m filters and cation exchange resins, respectively. Interestingly, the removal of U was obviously efficient by cation exchange resin, and the ceramic depth filter composed of diatomite exhibited the properties of both filtration and adsorption. It was found that the ceramic depth filter could adsorb -emitters, ␣-emitters, and uranium ions. The diatomite-based ceramic depth filter was able to simultaneously eliminate particles and adsorb ionic radionuclides from water. © 2012 Elsevier B.V. All rights reserved.
1. Introduction Decommissioning of the spent fuel pools (SFPs) is a necessary step in reducing radiological hazards in accordance with Taiwan’s national policy. Similarly, the health and safety of workers must be protected from radiological and non-radiological hazards associated with the radioactivity measurement and composition analyses of radioactive wastes. The results from the characterizations of SFPs can be used for further decommissioning planning to provide dose and risk assessments and to identify the types of safety and radiological protection required to protect workers, the general public, and the environment [1]. Treatment is a significant phase
∗ Corresponding author. Tel.: +886 3 4711400x3725; fax: +886 3 4713841. E-mail address: [email protected] (C.-P. Huang). 0304-3894/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2012.07.009
in the management of radioactive wastes, focusing on the volume reduction of generated wastes, safety enhancement of storage, and cost reduction for the further management phase. Taiwan Research Reactor (TRR), designed by Atomic Energy of Canada Limited (AECL), is a Canada Deuterium Uranium (CANDU) research reactor. The reactor uses heavy water as a moderator, light water as a coolant, graphite as a reflector, and natural uranium as fuel. In 1988, after the final shutdown of the TRR, the spent fuel was temporarily stored in a SFP, and it was batchstabilized using a hot cell for on-site dry storage. However, some of the spent fuel rods were disrupted during the wet storage period, and approximately 926 m3 of water was contaminated by spent, fine, fuel particles. It was expected that the water had been polluted by hazardous species, such as the fission products, cesium and strontium, and the actinides, such as uranium and the transuranic elements in ionic or solid form [2]. In 1990, filtration
C.-P. Huang et al. / Journal of Hazardous Materials 233–234 (2012) 140–147
and ion-exchange units were contaminated by spent fuel debris; afterward, the spent resins were temporarily stored in the TRR SFP. Since March 2010, based on the TRR SFP cleanup strategy, the spent resins were poured out and washed within the TRR SFP [3]. Thus, the turbidity and radiation dose rate increase during the handlings. Converting the radionuclides from water to solid media is essential to ensure long-term safe storage or disposal. Membrane processes (i.e., microfiltration, ultrafiltration, and reverse osmosis) and ion-exchange/adsorption technologies are widely applied for the treatment of liquid radioactive wastes (LRWs), and organic membrane filters (i.e., cellulose acetate, polypropylene, polysulphone, etc.) and ion-exchange resins (i.e., polystyrene cross-linked with divinyl benzene) are materials usually employed [4–8]. However, spent organic materials may be decomposed by high levels of radioactive energy generated by alpha decay and release of combustible gases; therefore, further stabilization is necessary prior to long-term storage or disposal [9]. Inorganic materials are much preferred for use in the treatment of LRWs, which contain alpha-emitters, due to their radioactive stability [10,11]. Inorganic filters manufactured from ceramics, metal, graphite, or any combination of these materials have been produced or are under development in an attempt to overcome the radiation stabilitylimiting features of polymeric membranes. The available ranges pore sizes of inorganic membranes are generally located on microfiltration and ultrafiltration applications [6,12]. Therein, ceramic membranes (e.g., Membralox® and CeRAM INSIDE® ) have been employed to treat non-active model solutions [12]. Regarding maintenance, the back-washable property of ceramic membranes is effective at reducing the frequency replacement and secondary wastes. The high cost of ceramic membranes due to their specific manufacturing process is the main drawback in their use. Diatomite, a relatively low cost and abundant material, is natural, amorphous silica formed by the deposition of a diatom skeleton [13]. Usually, the powder form of diatomite is used as a pre-coated filter aid and packed bed filler for filtration and adsorption purposes [14,15]. Natural and modified forms of diatomite have been applied to aqueous solutions to remove organic compounds, metal ions, uranium ions, and radionuclides [16–22]. In filtration systems, the use of powdered filter aids often results in clogging that reduces the lifetime of the filter, increases system maintenance costs, and hinders the recovery of the filters [23]. Following the development of filtration technology, the particulate diatomite was processed into a fine, porous ceramic filter candle. The main advantages of using the diatomite-based ceramic depth filters in water treatment systems are its simple operation, simple maintenance, and low cost [24]. Although diatomite-based ceramic depth filters have been widely applied in other filtration processes for more than a hundred years [24,25], they are rarely used for the treatment of LRWs. Identifying the properties of contaminants and testing the performance of water treatment materials are the major objectives of this work. The radionuclides and element compositions were analyzed using radiochemical analyses, scanning electron microscopy-energy dispersive spectrometer (SEM-EDS), and inductively coupled plasma-optical emission spectrometry (ICPOES). At the Institute of Nuclear Energy Research (INER), the primary acceptance criteria for LRWs treatment facilities using evaporation and/or reverse osmosis are a gross ˛ < 37 and ˇ < 370 Bq mL−1 . To meet the acceptance criteria, the contaminants removal efficiencies of purification materials, such as organic filter membranes, cation-exchange resins, and diatomite-based ceramic depth filters, were investigated. The capability of diatomite-based ceramic depth filters to remove radioactive species was of particular interest.
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2. Experimental procedure and instrument 2.1. Sample collection and treatment Most of the spent fuel particles were collected in a confinement precipitation bag, but small numbers were remained in the TRR SFP. On March 2010, the settled sludge in the TRR SFP was withdrawn and collected by suction filtration with a membrane filter (ϕ25 mm, 0.2 m, cellulose acetate, Advantec MFS, Inc., Pleasanton, CA, USA). The dried retentate was weighted and then placed into a digestion vessel (HP-500 Plus, Teflon-PFA) with 15 mL of nitric acid (65%, Merck) until fully liquefied using microwave digestion (CEM MARS-5, Matthews, NC, USA) at 200 ◦ C for 20 min. The liquefied sample was diluted to 100 mL by deionized water. Certain elements (i.e., U, Fe, Al, Si, and Ca) in the sludge were analyzed using ICP-OES (HORIBA Jobin Yvon, Japan). Non-filtrated original water in TRR SFP was taken once every two weeks to monitor the activity concentrations of 137 Cs and 90 Sr. During the cleanup of the spent resins, uncharacterized suspended solids interfered with the clarity of water. On June 2011, the original water was sampled from the surface of TRR SFP. After filtration, the collected retentate, or suspended solids, was dried, digested, and diluted in a process similar to that used for the treatment of sludge. The elements in the suspended solids were also analyzed using ICP-OES. The morphology and elemental composition percentage of the suspended solids were analyzed using a SEM (JEOL JSM-6510, Japan) equipped with an EDS (OXFORD INCAx-act, UK). The turbidity of the water was measured using a portable turbidimeter (WTW Turb 355 IR/T, Germany).
2.2. Water treatment trial system The water treatment trial system was cascaded within three housings, and each one was suitable for the placement of one filter measuring 10 in. in length or a packed bed cartridge. A centrifugal pump (1.0 L min−1 ) was used to feed water, and the sampling ports were assembled on the end of each housing unit. The original water was poured into the storage container in 10-L increments for each treatment trial. Two clusters of purification media were applied to treat the water. The first cluster of purification media (Trial I) had a sediment filter with 5 m pores, a ceramic depth filter with 0.9 m pores, a cation-exchange resin packed cartridge (wet weight ∼450 g and bed volume ∼500 mL), and a disposable, hollow fiber filter cartridge with 0.1–0.01 m pores. The sediment filter was processed using melted and blown polypropylene (PP) and the filter was constructed with many layers for true depth filtration (Purerite PS-05, 10 × 2.5 ). The cation-exchange resin (polystyrene cross-linked with divinyl benzene, SO3 Na) is produced by Tai-Young Chemical Co., Taiwan (DIAON SK1B), and the minimum exchange capacity is 2.0 meq mL−1 . The matrix of the hollow fiber membrane is PP, and the outer and inside diameters of fibers are 0.4–0.6 mm and 0.3–0.5 mm, respectively. The second cluster of purification media (Trial II) was cascaded using ceramic depth filters with 5-, 0.9-, and 0.22-m pore sizes in a series. The matrices of the ceramic depth filters were diatomite, provided by World Minerals® in the US. Bulk filters with 5-, 0.9-, and 0.22-m pore sizes were manufactured by sieving the diatomite particulates with wire sieves. The sieved diatomite particulates (250 g) were wetted and the resultant agglomerates were shaped using a cylindrical module and were sintered at 1100 ◦ C to provide sufficient mechanical strength. The ceramic depth filter used was a cartridge type that was open at both ends and had a plastic mounts and rubber washers at both ends with a diameter of 49 mm, a length of 248 mm, and an external surface area of 380 cm2 .
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First and second batch of original water for the treatment of Trial I and Trial II were respectively withdrawn from TRR SFP in September and December 2010. 2.3. Radiochemical analysis and instrument The same geometric vials were each filled with 100 mL of water and fully liquefied solid sample. The ␥-emitters of each sample were counted for 1000s with a gamma spectrometry system and a highpurity germanium detector (HPGe, GC4020) coupled to a personal computer analyzer using Genie 2000 software. The HPGe detector was enclosed inside CANBERRA 747 shielding with a 10-cm-thick lead, coated with 1 mm of tin and 1.6 mm of copper, to reduce the amount of background radiation from the environment. The activities of samples’ 90 Sr, gross ␣, and gross  were measured by a Tri-Carb® 2910TR low-background liquid scintillation analyzer (LSA, Packard, USA) with an alpha/beta discriminator. The activity of 90 Sr was determined using the Cerenkov counting method [26]. 2 mL of the sample was precipitated as carbonate, dissolved in 8 M nitric acid, and extracted with Sr resin (Eichrom); after 7 days of ingrowth of 90 Y, the daughter nuclides, the activities of 90 Sr were measured [27,28]. All of the analyses of gross ␣ and  levels were performed on 0.2 mL of the liquid samples in a Ultima Gold AB cocktail (PerkinElmer) filled to 20 mL in a plastic scintillation vial. The activities of samples’ gross ˛ and ˇ were simultaneous measured using LSA pulse decay analysis (PDA). The radionuclides of 36 Cl (Emax = 0.71 MeV) and 241 Am (Emax = 5.5 MeV) were used for beta and alpha sources, respectively. The optimum discriminator setting was 191, which was automatically generated by the instrument. 3. Results and discussion 3.1. Composition analysis of contaminants 3.1.1. Composition of sludge Because there was spent fuel-contaminated sludge present on the bottom of the TRR SFP, any operation involving underwater equipment movement and/or extraction would resuspend the sludge. Before the flush treatment of the spent resins, most of the solid contaminants were able to settle down during static periods. The membrane filter was used to collect 8.6 mg of the deep brown sludge. The analysis of gamma spectrometry for sludge indicated that the 57 Co, 60 Co, 137 Cs, and 241 Am radionuclides were measured, and the major ␥-emitter was 137 Cs (Table 1). The detection of 241 Am implies that other transuranic elements (i.e., Np, Pu, etc.) were present in the sludge, and most of the ␣-emitters should be transuranic elements. One of the pure -emitters, 90 Sr, was measured. − particles can also arise from radionuclides other than the most abundant fission products (i.e., 137 Cs and 90 Sr). In Table 1, Table 1 Radionuclides and elements in sludge. Radionuclide/element 137
−1
Cs (Bq g ) Co (Bq g−1 ) 60 Co (Bq g−1 ) 241 Am (Bq g−1 ) 90 Sr (Bq g−1 ) Gross  (Bq g−1 ) Gross ␣ (Bq g−1 ) U (mg g−1 ) Fe (mg g−1 ) Al (mg g−1 ) Si (mg g−1 ) Ca (mg g−1 ) 57
Quantity 4.13 × 106 1.63 × 104 5.87 × 104 1.50 × 105 3.66 × 105 1.75 × 107 2.16 × 106 2.42 1.41 0.40 0.14 0.12
Fig. 1. Monitoring of activity concentrations of 137 Cs and 90 Sr in TRR SFP.
the gross ␣ and gross  were much greater than the individual activity concentrations of radionuclides because of the lack of a proper high-resolution alpha analysis and because certain other pure -emitters were not measured. The element analysis results are also listed in Table 1. The composition of the sludge was complex because various sources contributed to the contamination. For example U, Al, and Fe come from spent fuel, aluminum shell tube cutting debris, and corrosion of steel devices, respectively. The Si and Ca may have been derived from dusts. 3.1.2. Composition of contaminants in water The trend of 137 Cs and 90 Sr activity concentrations in the TRR SFP water is shown in Fig. 1. It was found that the activity concentration of 137 Cs dramatically increased since May 2010 due to the extraction and flush treatment of spent resins. Because each of the original water samples was taken at a different time, the concentration of contaminants was not the same in subsequent experiments. The suspended solids interfered with the clarity of the water and with the execution of underwater tasks; therefore, the removal of turbidity was necessary. The only detected ␥-emitter in the water was 137 Cs. Because of their high water solubility, over 80% of -emitters were contributed by both 137 Cs and 90 Sr (Table 2). Using a 0.2-m membrane filter, the removal efficiencies of -emitters and ␣-emitters were 12.2% and 20.7%, respectively (Table 2). The removal of residual radionuclides should depend on the adsorption/ion-exchange method. Furthermore, the specific activities of transuranic elements were found to be approximately 104 –105 times greater than that of 238 U, and ␣ activity contributed by U (4.8 mg L−1 ) was extremely low (approximately 5 × 10−3 Bq mL−1 ). The major ␣-emitters presented in the water should be transuranic elements. In the measurement of 241 Am, the water sample was increased from 100 mL to 1 L for the analysis of ␥ spectrometry, but there was no 241 Am signal detected. The major ␣-emitters present in the water may be contributed by other transuranic elements, such as Pu. The leaching of radionuclides may have been caused by the degradation of ion-exchange resins under high radioactive irradiation and long-term storage. After filtering, the turbidity of the water was decreased from 5 NTU to 0.2 NTU. In addition, the low conductivity of the water (37 S cm−1 ) implied that there was only a small concentration of other competitive ions. Since the cleanup of spent resins, uncharacterized suspended solids have dispersed in the TRR SFP and few have settled down. Similar to sludge collection, the suspended solids were obtained from the filtration of the original water. However, the membrane
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Table 2 Radionuclides and elements in water and suspended solids. Radionuclide/element 137
−1
Cs (Bq mL ) 90 Sr (Bq mL−1 ) 60 Co (Bq mL−1 ) Gross  (Bq mL−1 ) Gross ␣ (Bq mL−1 ) U (mg L−1 ) Fe (mg L−1 ) Al (mg L−1 ) Si (mg L−1 ) Ca (mg L−1 ) Turbidity (NTU) a
Original 2.25 × 10 3.99 × 102 – 3.03 × 103 6.15 × 101 4.80 1.35 0.33 1.84 0.92 5.00 3
Filtratea
Suspended solids
Removal efficiency (%)
1.85 × 103 3.49 × 102 – 2.66 × 103 4.88 × 101 4.80 0.64 0.16 1.75 0.85 0.20
2.38 × 102 3.36 × 101 6.75 × 100 2.95 × 102 5.00 × 100 1.28 1.10 0.28 0.49 0.20 –
17.8 12.5 – 12.2 20.7 – – – – – –
Filtrate of the original water obtained with a 0.2-m membrane filter.
filter was almost clogged after filtrating 100 mL of the original water, and the dried retentate (Fig. 2(a)) was less than 0.2 mg, which was below the sensitivity of 4-digit balance. Hence, the quantity or activity of the suspended solids was expressed as the number per mL of water (Table 2). 137 Cs and trace amounts of 60 Co were measured by ␥ spectrometry, and the activity of 241 Am was below the minimum detectable activity (MDA, <10−1 Bq mL−1 ). The activity concentrations of 90 Sr and the ␣-emitters that remained in the suspended solids were 33.6 Bq mL−1 and 5 Bq mL−1 , respectively. The solids in the original water was not uniformly distributed, some higher specific gravity solids present in the bottom of the vial may not be efficiently measured. These solids which contained uranium and other metal debris were further collected by the filter membrane. Thus, the uranium and other metal debris present in original water were underestimated. The radioactivity attributed
to suspended solids was not significant, but the turbidity was significant. In Fig. 2(a), the gold suspended solids are a part of the ion-exchange resins. The morphology and elements on the surface of the suspended solids were observed using SEM-EDS. The suspended solids were composed of fine, spherical particle agglomerates with the particles ranging from 0.5 to >1 m in size (Fig. 2(b)). The compositions of the suspended solids, as determined by EDS, are listed in Table 3. The elements, such as C, O and S, are derived from ion-exchange resins. Due to over 20 years of radiolysis, part of the spent resins may be degraded and crushed under highly radioactive conditions. The specific gravity of the spent resins (1.09–1.20) was notably close to that of water, so it was difficult to naturally precipitate the crushed, spent resins colloids. 3.2. Water treatment 3.2.1. Efficiency of contaminants removal in the Trial I system In the Trial I system, filters with different particle rejection pore sizes (5, 0.9, and 0.1–0.01 m) and cation-exchange resin were employed to remove suspended solids and radioactive ions from the TRR SFP water. The turbidity of the first batch original water was 3.63 NTU, but it decreased to 1.76, 0.9, and 0.48 NTU after passing through the PP, ceramic, and hollow fiber filters, respectively (data not shown). The removal capabilities of -emitters, ␣-emitters, and U for Trial I system are shown in Fig. 3 and the removal efficiencies of each unit are listed in Table 4. The removal efficiencies in Table 4 are calculated with respect to the previous residual activities, not the initial activities. The turbidity was slightly decreased, but only a few radioactive species were blocked by the PP filter. Interestingly, after passing through the ceramic depth filter, most of the 137 Cs, 90 Sr, -emitters, and ␣-emitters were trapped (Fig. 3(a) and (b)). More than 95% of -emitters and ␣-emitters were removed by the filter, but only 55% of U was removed (Fig. 3(c) and Table 4). Therefore, it was determined that the ceramic filter had dual purification
Table 3 SEM-EDS analysis of suspended solids.
Fig. 2. (a) Appearance and (b) SEM-image of suspended solids in the TRR SFP.
Element
Weight%
C O Mg Al Si S Cl K Ca Fe U Totals
23.92 34.46 0.86 6.24 12.59 1.17 0.75 2.32 0.88 10.28 6.54 100.00
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Table 4 The removal efficiencies of radionuclides by a specific purification unit. Unit
137 Cs efficiency (%) (residual, Bq mL−1 )
90 Sr efficiency (%) (residual, Bq mL−1 )
Gross ˇ efficiency (%) (residual, Bq mL−1 )
Gross ˛ efficiency (%) (residual, Bq mL−1 )
U efficiency (%) (residual, mg L−1 )
PP filter (5 m)
2.6 (1.12 × 103 ) 96.6 (3.76 × 101 ) 90.9 (3.42 × 100 ) 84.4 (5.30 × 10−1 ) 99.7 99.1 (1.74 × 101 ) 71.2 (5.01 × 100 ) 78.6 (1.07 × 100 ) 99.9
2.3 (6.35 × 101 ) 93.6 (4.10 × 100 ) 77.0 (9.30 × 10−1 ) 18.7 (7.50 × 10−1 ) 98.8 28.9 (2.29 × 102 ) 46.3 (1.23 × 102 ) 99.3 (8.90 × 10−1 ) 99.7
0.7 (1.51 × 103 ) 96.9 (4.74 × 101 ) 82.2 (8.42 × 100 ) 46.6 (4.50 × 100 ) 99.7 77.0 (5.84 × 102 ) 46.7 (3.11 × 102 ) 96.7 (1.02 × 101 ) 99.6
1.9 (2.65 × 101 ) 97.8 (5.80 × 10−1 ) 71.4 (1.70 × 10−1 ) 40.1 (1.00 × 10−1 ) 99.6 54.0 (2.16 × 101 ) 29.2 (1.53 × 101 ) 95.1 (7.50 × 10−1 ) 98.4
3.5 (1.93) 54.9 (0.87) 88.5 (0.10) – (0.10) 95.0 29.3 (3.11) 20.6 (2.47) 57.1 (1.06) 75.9
Ceramic filter (0.9 m) Cation exchange resin Hollow fiber filter Overall Trial I 1st filter (5.0 m) 2nd filter (0.9 m) 3rd filter (0.22 m) Overall Trial II
capabilities:particles rejection and ion adsorption. Ionic radionuclides and fine particles were further removed by the treatment of ion-exchange resin and hollow fiber filter, respectively (Fig. 3). More than 99% of -emitters and ␣-emitters were removed by a Trial I system (Table 4). 3.2.2. Efficiency of contaminants removal in the Trial II system In the Trial II system, three cascaded ceramic filters with different pore sizes were used to investigate the major particle size distributions of the suspended solids. Additionally, the capability of ceramic filters to remove particles and ionic radionuclides was investigated. The second batch of original water contained much higher concentrations of radioactive ions and suspended solids than the first batch. The turbidity of the original water was 4.37 NTU, which was decreased to 2.3, 1.9, and 0.8 NTU by each of the ceramic filters, respectively (data not shown). The removal efficiencies of -emitters, ␣-emitters, and U by the Trial II system are shown in Fig. 4 and Table 4. Based on the results of Trial I, the radionuclides, 137 Cs and 90 Sr, cannot be removed by the filter with a 5.0-m rejection pore. However, approximately 99% and 29% of 137 Cs and 90 Sr were respectively removed by the first filter (5.0 m) via adsorption mechanism (Table 4). The diatomitebased ceramic filter tended to have higher adsorption efficiency of 137 Cs than 90 Sr (Fig. 4(a)). Hence, the diatomite-based ceramic filter exhibits adsorption selectivity on different ions. The second filter (0.9 m) in Trial II was found to have lower removal efficiencies of 90 Sr, -emitters, ␣-emitters, and U than observed in Trial I system due to the higher initial concentrations of radioactive ions and suspended solids. Most of the 90 Sr present in the water are ions, and the removal efficiency of 90 Sr will decrease as the initial concentration increase. On the other hand, the suspended solids with different sizes and compositions were separated by different pore sizes of the filters. Only few amounts of suspended solids can be removed by the first filter. However, some of the suspended solids with sizes ranging from 5 to 0.9 m may contain lower concentrations of -emitters, ␣-emitters, and U, but these species will occupy the adsorption sites of the second filter and lead to the decrease of adsorption capacity. The residual fine suspended solids (<0.9 m) contain higher concentrations of -emitters, ␣-emitters, and U can be further removed by the third filter (Fig. 4(b) and (c)). 3.2.3. Contaminant removal capability of the ceramic filter Five batches of original water (totaling 50 L) were treated using the Trial II system without replacing the filters. The accumulations of activities on the filters were plotted with the total feed-in volume (Fig. 5). After treating 30 L of the water, the
accumulation of -emitters (a major was 137 Cs) on the first filter almost reached a plateau, which implied adsorption saturation (Fig. 5(a)). The adsorption tendency of the 137 Cs was similar to that of the -emitters (data not shown). As the first filter reached saturation, the removal of residual contaminants depended on the second and third filters. Moreover, due to the higher particle rejection capability of the second filter, the accumulation of -emitters on the second filter was found to be higher than the first filter. Because the first and second filters removed most of the -emitters, the loading of the third filter was relatively low. After the treatment of 50 L of the water, the overall removal efficiencies of the -emitters and ␣-emitters were greater than 95% and 91%, respectively (Fig. 5(a) and (b)). It is well-known that the materials with larger specific surface areas will have more adsorption sites. In a comparison of the three filters, which were composed of different particle sizes of raw diatomite, the third filter should have the largest specific surface area and the greatest adsorption capacity. Furthermore, the fine suspended solids which have higher concentration of U can be effectively filtrated by the third filter, so the third filter captured more U than the first filter, as observed in Fig. 5(c). On the other hand, the adsorption efficiency of ions was proportional to the numbers of adsorption sites of the filter. The molar concentration of U (∼20 M) in the original water was much greater than the other radionuclides (e.g., 137 Cs ∼6 × 10−3 M), so the numbers of active adsorption sites will have obviously effect on adsorption efficiency. During the filtration, more and more suspended solids deposited on the surface of filters and occupied the adsorption sites, so the removal efficiency of U gradually decreased. Because the water in the TRR SFP had a complicated composition with various radionuclides in ionic and solid form, distinguishing the filtration or adsorption effects was difficult. After treating 50 L of the water, the volumetric flowrate of the system decreased to 0.33 L min−1 (only one third of the initial flow, data not shown), and the primary fouling occurred to the third filter. Furthermore, the specific ˛ activity for the first filter was 5120 Bq g−1 , which is classified as TRU waste. However, the further conditioning and storage of spent ceramic filters was relatively simple and safer than conditioning and storing spent organic purification media. During the treatment of water, the surface dose rate of each filter was gradually increased, and the major radiation dose rate was contributed by the 137 Cs (/␥-emitter). The maximum surface dose rate of the first filter was 863 Sv/h; therefore the replacement or disposal of the spent filter could be contact-handled (<2 mSv/h).
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Fig. 3. The removal of (a) -emitters, 137 Cs, and 90 Sr (b) ␣-emitters, and (c) U from the water in treatment Trial I. The original activity concentrations of -emitters, 137 Cs, 90 Sr, and ␣-emitters are shown as 1520, 1150, 65, and 27 Bq mL−1 , respectively; initial concentration of U is 2 mg L−1 ; pH = 5.48; conductivity = 30.3 S cm−1 .
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Fig. 4. The removal of (a) -emitters, 137 Cs, and 90 Sr (b) ␣-emitters, and (c) U from the water in treatment Trial I. The original activity concentrations of -emitters, 137 Cs, 90 Sr, and ␣-emitters are shown as 2540, 2010, 322, and 47 Bq mL−1 , respectively; initial concentration of U is 4.4 mg L−1 ; pH = 4.98; conductivity = 42.0 S cm−1 .
80 70
40
100
80 60 50
60
40
30 20
20
Overallremoval efficiency (%)
(a) -emitters (MBq)
C.-P. Huang et al. / Journal of Hazardous Materials 233–234 (2012) 140–147
Accumulation of
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Table 5 Studies of radionuclides removal by diatomite materials. Radionuclide
Efficiency
Removal
Form of diatomite
Reference
137
Cs, 134 Cs, 60 Co 137 Cs 90 Sr Gross  Gross ␣
85% 94% 94% 95% 91%
2.21 MBq Powder (60 mesh, 50 kg) 100 MBq Filter candles (750 g) 19.5 MBq 149 MBq 2.61 MBq
[21] This study
U
43% 90%
132 mg 6.10 mg 159 mg
This study [22]
10
Filter candles (750 g) Powder (0.25 mm, 1 g) Powder (modified, 1 g)
0
0 0
20
40
60
100
80
120
140
160
4. Conclusions
Feed-in volume (L) 100
(b) 1.4
80 1.0 60
0.8 0.6
40
0.4 20
Overallremoval efficiency (%)
Accumulation -emitters (MBq)
1.2
0.2 0
0.0 0.0
0.5
1.0
1.5
2.0
2.5
3.0
Acknowledgements
Feed-in volume (L)
(c) 70 60
Accumulation of U (mg)
80 50 60
40 30
40
20 20
Overall removal efficiency (%)
100
10 0
0 0
50
100
15 0
200
To meet the acceptance criteria of INER’s LRWs treatment facilities, the removal of radioactive solids and ions in TRR SFP water is achieved by both trial systems. Based on the analyses, the radioactive solids were quite complex, they were primary attributed by the ruptured spent fuel, spent resins, and metal debris. Most of the radioactive solids could be removed by the filters with less than 0.9 m of rejection pores. Most of the radioactive ions, 137 Cs, 90 Sr, and U, in TRR SFP water could be removed by cation exchange resin and diatomite-based ceramic filter. The removal of U by cation exchange resin was found to be more efficient than diatomitebased ceramic filter. The diatomite-based ceramic filters exhibited excellent removal efficiency on the particles filtration and ions adsorption. Using the candle type diatomite filters for the treatment of TRR SFP water is feasible.
250
300
350
Feed-in volume (L)
Fig. 5. The activity accumulation and removal efficiency () of (a) -emitters, (b) ␣-emitters and (c) U with total feed-in mass of contaminants for ceramic filters with different pore sizes 5.0 m (䊉), 0.9 m () and 0.22 m ().
The studies on the radionuclides removal using diatomite materials have been summarized in Table 5. Three candle type diatomite filters (750 g) were employed to remove the radioactive ions and suspended solids in this study. The activity concentrations of radionuclides present in the TRR SFP water were found to be much higher than the other work [21], and more than 90% of 137 Cs, 90 Sr, emitters, and ␣-emitters were removed in this study. The removal of U by the filters was probably retarded by the suspended solids, so the removal efficiency is relative lower than the other work [22].
The authors would like to thank the National Science Council of Taiwan for financially supporting this research under Contract No. ANS02. Mr. Te-Yen Su is appreciated for the SEM-EDS analysis. References [1] International Atomic Energy Agency, Radiological characterization of shut down nuclear reactors for decommissioning purposes, IAEA Technical Reports Series No. 389, IAEA, Vienna, Austria, 1998. [2] Z. Hozer, E. Szab, T. Pinter, B. Varju Ilona, T. Bujtas, G. Farkas, N. Vajda, Activity release from damaged fuel during the Paks-2 cleaning tank incident in the spent fuel storage pool, J. Nucl. Mater. 392 (2009) 90–94. [3] C.P. Huang, K.Y. Chang, H.B. Chen, Flush treatment of high activity and transuranic contaminated mixed-bed ion-exchange resin in TRR spent fuel pool, in: 2010 The 3rd East Asia Forum on Radwaste Management, Gyeongiu, Korea, 2010. [4] International Atomic Energy Agency, Handling and processing of radioactive waste from nuclear application, IAEA Technical Reports Series No. 402, IAEA, Vienna, Austria, 2001. [5] R.O.A. Rahman, H.A. Ibrahium, Y.T. Hung, Liquid radioactive wastes treatment: a review, Water 3 (2011) 551–565. [6] International Atomic Energy Agency, Application of membrane technologies for liquid radioactive waste processing, IAEA Technical Reports Series No. 431, IAEA, Vienna, Austria, 2004. [7] D.V. Marinin, G.N. Brown, Studies of sorbent/ion-exchange materials for the removal of radioactive strontium from liquid radioactive waste and high hardness groundwaters, Waste Manage. 20 (2000) 545–553. [8] K. Raj, K.K. Prasad, N.K. Bansal, Radioactive waste management practices in India, Nucl. Eng. Des. 236 (2006) 914–930. [9] Electric Power Research Institute, Evaluation of hydrogen diffusion from lowlevel radioactive waste containers, EPRI NP-6244, EPRI, CA, US, 1989. [10] B.T. Kim, H.K. Lee, H. Moon, K.J. Lee, Adsorption of radionuclides from aqueous solutions by inorganic adsorbents, Sep. Sci. Technol. 30 (1995) 3165– 3182. [11] E. Galunin, M.D. Alba, M.J. Santos, T. Abrao, M. Vidal, Examination of competitive lanthanide sorption onto smectites and its significance in the management of radioactive waste, J. Hazard. Mater. 186 (2011) 1930–1941. [12] G. Zakrzewska-Trznadel, Radioactive solutions treatment by hybrid complexation-UF/NF process, J. Membr. Sci. 225 (2003) 25–39. [13] H.E.G.M.M. Bakr, Diatomite: its characterization, modifications and applications, Asian J. Mater. Sci. 2 (2010) 121–136.
C.-P. Huang et al. / Journal of Hazardous Materials 233–234 (2012) 140–147 [14] A.S. Murathan, Removal of strontium, aluminium, manganese and iron ions from aqueous solutions in packed beds, Fresenius Environ. Bull. 13 (2004) 481–484. [15] R.W. Mcindoe, Diatomaceous earth filtration for water supplies I, J. Am. Water Work Assoc. (1969) 50–53. [16] M. Aivalioti, I. Vamvasakis, E. Gidarakos, BTEX, MTBE adsorption onto raw and thermally modified diatomite, J. Hazard. Mater. 178 (2010) 136–143. [17] M.A. Al-Ghouti, M.A.M. Khraisheh, M. Tutuji, Flow injection potentiometric stripping analysis for study of adsorption of heavy metal ions onto modified diatomite, Chem. Eng. J. 104 (2004) 83–91. [18] G.G. Martirosyan, A.G. Manukyan, E.B. Ovsepyan, K.A. Kostanyan, A study of adsorption-structural properties of natural and processed diatomites, Russ. J. Appl. Chem. 76 (2003) 531–535. [19] J.F. Lemonas, Diatomite, Am. Ceram. Soc. Bull. 76 (1997) 92–95. [20] N. Caliskan, A.R. Kul, S. Alkan, E.G. Sogut, I. Alacabey, Adsorption of zinc(II) on diatomite and manganese-oxide-modified diatomite: a kinetic and equilibrium study, J. Hazard. Mater. 193 (2011) 27–36. [21] A.E. Osmanlioglu, Natural diatomite process for removal of radioactivity from liquid waste, Appl. Radiat. Isot. 65 (2007) 17–20.
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[22] M. Sprynskyy, I. Kovalchuk, B. Buszewski, The separation of uranium ions by natural and modified diatomite from aqueous solution, J. Hazard. Mater. 181 (2010) 700–707. [23] V. Krapukhin, V. Zhurin, V. Lavrikov, V. Kulemin, Removal of suspended solids from radioactive solutions using helical filtering elements, Atom. Energy 86 (1999) 447–449. [24] B. Michen, A. Diatta, J. Fritsch, C. Aneziris, T. Graule, Removal of colloidal particles in ceramic depth filters based on diatomaceous earth, Sep. Purif. Technol. 81 (2011) 77–87. [25] O. San, A. Imaretli, Preparation and filtration testing of diatomite filtering layer by acid leaching, Ceram. Int. 37 (2011) 73–78. [26] T.C. Chu, J.J. Wang, Y.M. Lin, Radiostrontium analytical method using crown-ether compound and Cerenkov counting and its applications in environmental monitoring, Appl. Radiat. Isot. 49 (1998) 1671–1675. [27] N. Vajda, C.K. Kim, Determination of radiostrontium isotopes: a review of analytical methodology, Appl. Radiat. Isot. 68 (2010) 2306–2326. [28] ASTM, Standard Test Method for Sr-90 in Water, American Society for Testing and Materials, New York, 1995.
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Characterization of radioactive contaminants and water treatment trials for the Taiwan Research Reactor’s spent fuel pool Chun-Ping Huang ∗ , Tzung-Yi Lin, Ling-Huan Chiao, Hong-Bin Chen Institute of Nuclear Energy Research, 1000, Wenhua Road, Jiaan Village, Longtan Township, Taoyuan County 32546, Taiwan, ROC
h i g h l i g h t s
Deal with a practical radioactive contamination in Taiwan Research Reactor spent fuel pool water. Identify the properties of radioactive contaminants and performance test for water treatment materials. The radioactive solids were primary attributed by ruptured spent fuels, spent resins, and metal debris. The radioactive ions were major composed by uranium and fission products. Diatomite-based ceramic depth filter can simultaneously removal radioactive solids and ions.
a r t i c l e
i n f o
Article history: Received 26 September 2011 Received in revised form 8 June 2012 Accepted 2 July 2012 Available online 11 July 2012 Keywords: Radioactive contaminant Water treatment Spent fuel pool Taiwan Research Reactor
a b s t r a c t There were approximately 926 m3 of water contaminated by fission products and actinides in the Taiwan Research Reactor’s spent fuel pool (TRR SFP). The solid and ionic contaminants were thoroughly characterized using radiochemical analyses, scanning electron microscopy equipped with an energy dispersive spectrometer (SEM-EDS), and inductively coupled plasma optical emission spectrometry (ICP-OES) in this study. The sludge was made up of agglomerates contaminated by spent fuel particles. Suspended solids from spent ion-exchange resins interfered with the clarity of the water. In addition, the ionic radionuclides such as 137 Cs, 90 Sr, U, and ␣-emitters, present in the water were measured. Various filters and cation-exchange resins were employed for water treatment trials, and the results indicated that the solid and ionic contaminants could be effectively removed through the use of <0.9 m filters and cation exchange resins, respectively. Interestingly, the removal of U was obviously efficient by cation exchange resin, and the ceramic depth filter composed of diatomite exhibited the properties of both filtration and adsorption. It was found that the ceramic depth filter could adsorb -emitters, ␣-emitters, and uranium ions. The diatomite-based ceramic depth filter was able to simultaneously eliminate particles and adsorb ionic radionuclides from water. © 2012 Elsevier B.V. All rights reserved.
1. Introduction Decommissioning of the spent fuel pools (SFPs) is a necessary step in reducing radiological hazards in accordance with Taiwan’s national policy. Similarly, the health and safety of workers must be protected from radiological and non-radiological hazards associated with the radioactivity measurement and composition analyses of radioactive wastes. The results from the characterizations of SFPs can be used for further decommissioning planning to provide dose and risk assessments and to identify the types of safety and radiological protection required to protect workers, the general public, and the environment [1]. Treatment is a significant phase
∗ Corresponding author. Tel.: +886 3 4711400x3725; fax: +886 3 4713841. E-mail address: [email protected] (C.-P. Huang). 0304-3894/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2012.07.009
in the management of radioactive wastes, focusing on the volume reduction of generated wastes, safety enhancement of storage, and cost reduction for the further management phase. Taiwan Research Reactor (TRR), designed by Atomic Energy of Canada Limited (AECL), is a Canada Deuterium Uranium (CANDU) research reactor. The reactor uses heavy water as a moderator, light water as a coolant, graphite as a reflector, and natural uranium as fuel. In 1988, after the final shutdown of the TRR, the spent fuel was temporarily stored in a SFP, and it was batchstabilized using a hot cell for on-site dry storage. However, some of the spent fuel rods were disrupted during the wet storage period, and approximately 926 m3 of water was contaminated by spent, fine, fuel particles. It was expected that the water had been polluted by hazardous species, such as the fission products, cesium and strontium, and the actinides, such as uranium and the transuranic elements in ionic or solid form [2]. In 1990, filtration
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and ion-exchange units were contaminated by spent fuel debris; afterward, the spent resins were temporarily stored in the TRR SFP. Since March 2010, based on the TRR SFP cleanup strategy, the spent resins were poured out and washed within the TRR SFP [3]. Thus, the turbidity and radiation dose rate increase during the handlings. Converting the radionuclides from water to solid media is essential to ensure long-term safe storage or disposal. Membrane processes (i.e., microfiltration, ultrafiltration, and reverse osmosis) and ion-exchange/adsorption technologies are widely applied for the treatment of liquid radioactive wastes (LRWs), and organic membrane filters (i.e., cellulose acetate, polypropylene, polysulphone, etc.) and ion-exchange resins (i.e., polystyrene cross-linked with divinyl benzene) are materials usually employed [4–8]. However, spent organic materials may be decomposed by high levels of radioactive energy generated by alpha decay and release of combustible gases; therefore, further stabilization is necessary prior to long-term storage or disposal [9]. Inorganic materials are much preferred for use in the treatment of LRWs, which contain alpha-emitters, due to their radioactive stability [10,11]. Inorganic filters manufactured from ceramics, metal, graphite, or any combination of these materials have been produced or are under development in an attempt to overcome the radiation stabilitylimiting features of polymeric membranes. The available ranges pore sizes of inorganic membranes are generally located on microfiltration and ultrafiltration applications [6,12]. Therein, ceramic membranes (e.g., Membralox® and CeRAM INSIDE® ) have been employed to treat non-active model solutions [12]. Regarding maintenance, the back-washable property of ceramic membranes is effective at reducing the frequency replacement and secondary wastes. The high cost of ceramic membranes due to their specific manufacturing process is the main drawback in their use. Diatomite, a relatively low cost and abundant material, is natural, amorphous silica formed by the deposition of a diatom skeleton [13]. Usually, the powder form of diatomite is used as a pre-coated filter aid and packed bed filler for filtration and adsorption purposes [14,15]. Natural and modified forms of diatomite have been applied to aqueous solutions to remove organic compounds, metal ions, uranium ions, and radionuclides [16–22]. In filtration systems, the use of powdered filter aids often results in clogging that reduces the lifetime of the filter, increases system maintenance costs, and hinders the recovery of the filters [23]. Following the development of filtration technology, the particulate diatomite was processed into a fine, porous ceramic filter candle. The main advantages of using the diatomite-based ceramic depth filters in water treatment systems are its simple operation, simple maintenance, and low cost [24]. Although diatomite-based ceramic depth filters have been widely applied in other filtration processes for more than a hundred years [24,25], they are rarely used for the treatment of LRWs. Identifying the properties of contaminants and testing the performance of water treatment materials are the major objectives of this work. The radionuclides and element compositions were analyzed using radiochemical analyses, scanning electron microscopy-energy dispersive spectrometer (SEM-EDS), and inductively coupled plasma-optical emission spectrometry (ICPOES). At the Institute of Nuclear Energy Research (INER), the primary acceptance criteria for LRWs treatment facilities using evaporation and/or reverse osmosis are a gross ˛ < 37 and ˇ < 370 Bq mL−1 . To meet the acceptance criteria, the contaminants removal efficiencies of purification materials, such as organic filter membranes, cation-exchange resins, and diatomite-based ceramic depth filters, were investigated. The capability of diatomite-based ceramic depth filters to remove radioactive species was of particular interest.
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2. Experimental procedure and instrument 2.1. Sample collection and treatment Most of the spent fuel particles were collected in a confinement precipitation bag, but small numbers were remained in the TRR SFP. On March 2010, the settled sludge in the TRR SFP was withdrawn and collected by suction filtration with a membrane filter (ϕ25 mm, 0.2 m, cellulose acetate, Advantec MFS, Inc., Pleasanton, CA, USA). The dried retentate was weighted and then placed into a digestion vessel (HP-500 Plus, Teflon-PFA) with 15 mL of nitric acid (65%, Merck) until fully liquefied using microwave digestion (CEM MARS-5, Matthews, NC, USA) at 200 ◦ C for 20 min. The liquefied sample was diluted to 100 mL by deionized water. Certain elements (i.e., U, Fe, Al, Si, and Ca) in the sludge were analyzed using ICP-OES (HORIBA Jobin Yvon, Japan). Non-filtrated original water in TRR SFP was taken once every two weeks to monitor the activity concentrations of 137 Cs and 90 Sr. During the cleanup of the spent resins, uncharacterized suspended solids interfered with the clarity of water. On June 2011, the original water was sampled from the surface of TRR SFP. After filtration, the collected retentate, or suspended solids, was dried, digested, and diluted in a process similar to that used for the treatment of sludge. The elements in the suspended solids were also analyzed using ICP-OES. The morphology and elemental composition percentage of the suspended solids were analyzed using a SEM (JEOL JSM-6510, Japan) equipped with an EDS (OXFORD INCAx-act, UK). The turbidity of the water was measured using a portable turbidimeter (WTW Turb 355 IR/T, Germany).
2.2. Water treatment trial system The water treatment trial system was cascaded within three housings, and each one was suitable for the placement of one filter measuring 10 in. in length or a packed bed cartridge. A centrifugal pump (1.0 L min−1 ) was used to feed water, and the sampling ports were assembled on the end of each housing unit. The original water was poured into the storage container in 10-L increments for each treatment trial. Two clusters of purification media were applied to treat the water. The first cluster of purification media (Trial I) had a sediment filter with 5 m pores, a ceramic depth filter with 0.9 m pores, a cation-exchange resin packed cartridge (wet weight ∼450 g and bed volume ∼500 mL), and a disposable, hollow fiber filter cartridge with 0.1–0.01 m pores. The sediment filter was processed using melted and blown polypropylene (PP) and the filter was constructed with many layers for true depth filtration (Purerite PS-05, 10 × 2.5 ). The cation-exchange resin (polystyrene cross-linked with divinyl benzene, SO3 Na) is produced by Tai-Young Chemical Co., Taiwan (DIAON SK1B), and the minimum exchange capacity is 2.0 meq mL−1 . The matrix of the hollow fiber membrane is PP, and the outer and inside diameters of fibers are 0.4–0.6 mm and 0.3–0.5 mm, respectively. The second cluster of purification media (Trial II) was cascaded using ceramic depth filters with 5-, 0.9-, and 0.22-m pore sizes in a series. The matrices of the ceramic depth filters were diatomite, provided by World Minerals® in the US. Bulk filters with 5-, 0.9-, and 0.22-m pore sizes were manufactured by sieving the diatomite particulates with wire sieves. The sieved diatomite particulates (250 g) were wetted and the resultant agglomerates were shaped using a cylindrical module and were sintered at 1100 ◦ C to provide sufficient mechanical strength. The ceramic depth filter used was a cartridge type that was open at both ends and had a plastic mounts and rubber washers at both ends with a diameter of 49 mm, a length of 248 mm, and an external surface area of 380 cm2 .
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First and second batch of original water for the treatment of Trial I and Trial II were respectively withdrawn from TRR SFP in September and December 2010. 2.3. Radiochemical analysis and instrument The same geometric vials were each filled with 100 mL of water and fully liquefied solid sample. The ␥-emitters of each sample were counted for 1000s with a gamma spectrometry system and a highpurity germanium detector (HPGe, GC4020) coupled to a personal computer analyzer using Genie 2000 software. The HPGe detector was enclosed inside CANBERRA 747 shielding with a 10-cm-thick lead, coated with 1 mm of tin and 1.6 mm of copper, to reduce the amount of background radiation from the environment. The activities of samples’ 90 Sr, gross ␣, and gross  were measured by a Tri-Carb® 2910TR low-background liquid scintillation analyzer (LSA, Packard, USA) with an alpha/beta discriminator. The activity of 90 Sr was determined using the Cerenkov counting method [26]. 2 mL of the sample was precipitated as carbonate, dissolved in 8 M nitric acid, and extracted with Sr resin (Eichrom); after 7 days of ingrowth of 90 Y, the daughter nuclides, the activities of 90 Sr were measured [27,28]. All of the analyses of gross ␣ and  levels were performed on 0.2 mL of the liquid samples in a Ultima Gold AB cocktail (PerkinElmer) filled to 20 mL in a plastic scintillation vial. The activities of samples’ gross ˛ and ˇ were simultaneous measured using LSA pulse decay analysis (PDA). The radionuclides of 36 Cl (Emax = 0.71 MeV) and 241 Am (Emax = 5.5 MeV) were used for beta and alpha sources, respectively. The optimum discriminator setting was 191, which was automatically generated by the instrument. 3. Results and discussion 3.1. Composition analysis of contaminants 3.1.1. Composition of sludge Because there was spent fuel-contaminated sludge present on the bottom of the TRR SFP, any operation involving underwater equipment movement and/or extraction would resuspend the sludge. Before the flush treatment of the spent resins, most of the solid contaminants were able to settle down during static periods. The membrane filter was used to collect 8.6 mg of the deep brown sludge. The analysis of gamma spectrometry for sludge indicated that the 57 Co, 60 Co, 137 Cs, and 241 Am radionuclides were measured, and the major ␥-emitter was 137 Cs (Table 1). The detection of 241 Am implies that other transuranic elements (i.e., Np, Pu, etc.) were present in the sludge, and most of the ␣-emitters should be transuranic elements. One of the pure -emitters, 90 Sr, was measured. − particles can also arise from radionuclides other than the most abundant fission products (i.e., 137 Cs and 90 Sr). In Table 1, Table 1 Radionuclides and elements in sludge. Radionuclide/element 137
−1
Cs (Bq g ) Co (Bq g−1 ) 60 Co (Bq g−1 ) 241 Am (Bq g−1 ) 90 Sr (Bq g−1 ) Gross  (Bq g−1 ) Gross ␣ (Bq g−1 ) U (mg g−1 ) Fe (mg g−1 ) Al (mg g−1 ) Si (mg g−1 ) Ca (mg g−1 ) 57
Quantity 4.13 × 106 1.63 × 104 5.87 × 104 1.50 × 105 3.66 × 105 1.75 × 107 2.16 × 106 2.42 1.41 0.40 0.14 0.12
Fig. 1. Monitoring of activity concentrations of 137 Cs and 90 Sr in TRR SFP.
the gross ␣ and gross  were much greater than the individual activity concentrations of radionuclides because of the lack of a proper high-resolution alpha analysis and because certain other pure -emitters were not measured. The element analysis results are also listed in Table 1. The composition of the sludge was complex because various sources contributed to the contamination. For example U, Al, and Fe come from spent fuel, aluminum shell tube cutting debris, and corrosion of steel devices, respectively. The Si and Ca may have been derived from dusts. 3.1.2. Composition of contaminants in water The trend of 137 Cs and 90 Sr activity concentrations in the TRR SFP water is shown in Fig. 1. It was found that the activity concentration of 137 Cs dramatically increased since May 2010 due to the extraction and flush treatment of spent resins. Because each of the original water samples was taken at a different time, the concentration of contaminants was not the same in subsequent experiments. The suspended solids interfered with the clarity of the water and with the execution of underwater tasks; therefore, the removal of turbidity was necessary. The only detected ␥-emitter in the water was 137 Cs. Because of their high water solubility, over 80% of -emitters were contributed by both 137 Cs and 90 Sr (Table 2). Using a 0.2-m membrane filter, the removal efficiencies of -emitters and ␣-emitters were 12.2% and 20.7%, respectively (Table 2). The removal of residual radionuclides should depend on the adsorption/ion-exchange method. Furthermore, the specific activities of transuranic elements were found to be approximately 104 –105 times greater than that of 238 U, and ␣ activity contributed by U (4.8 mg L−1 ) was extremely low (approximately 5 × 10−3 Bq mL−1 ). The major ␣-emitters presented in the water should be transuranic elements. In the measurement of 241 Am, the water sample was increased from 100 mL to 1 L for the analysis of ␥ spectrometry, but there was no 241 Am signal detected. The major ␣-emitters present in the water may be contributed by other transuranic elements, such as Pu. The leaching of radionuclides may have been caused by the degradation of ion-exchange resins under high radioactive irradiation and long-term storage. After filtering, the turbidity of the water was decreased from 5 NTU to 0.2 NTU. In addition, the low conductivity of the water (37 S cm−1 ) implied that there was only a small concentration of other competitive ions. Since the cleanup of spent resins, uncharacterized suspended solids have dispersed in the TRR SFP and few have settled down. Similar to sludge collection, the suspended solids were obtained from the filtration of the original water. However, the membrane
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Table 2 Radionuclides and elements in water and suspended solids. Radionuclide/element 137
−1
Cs (Bq mL ) 90 Sr (Bq mL−1 ) 60 Co (Bq mL−1 ) Gross  (Bq mL−1 ) Gross ␣ (Bq mL−1 ) U (mg L−1 ) Fe (mg L−1 ) Al (mg L−1 ) Si (mg L−1 ) Ca (mg L−1 ) Turbidity (NTU) a
Original 2.25 × 10 3.99 × 102 – 3.03 × 103 6.15 × 101 4.80 1.35 0.33 1.84 0.92 5.00 3
Filtratea
Suspended solids
Removal efficiency (%)
1.85 × 103 3.49 × 102 – 2.66 × 103 4.88 × 101 4.80 0.64 0.16 1.75 0.85 0.20
2.38 × 102 3.36 × 101 6.75 × 100 2.95 × 102 5.00 × 100 1.28 1.10 0.28 0.49 0.20 –
17.8 12.5 – 12.2 20.7 – – – – – –
Filtrate of the original water obtained with a 0.2-m membrane filter.
filter was almost clogged after filtrating 100 mL of the original water, and the dried retentate (Fig. 2(a)) was less than 0.2 mg, which was below the sensitivity of 4-digit balance. Hence, the quantity or activity of the suspended solids was expressed as the number per mL of water (Table 2). 137 Cs and trace amounts of 60 Co were measured by ␥ spectrometry, and the activity of 241 Am was below the minimum detectable activity (MDA, <10−1 Bq mL−1 ). The activity concentrations of 90 Sr and the ␣-emitters that remained in the suspended solids were 33.6 Bq mL−1 and 5 Bq mL−1 , respectively. The solids in the original water was not uniformly distributed, some higher specific gravity solids present in the bottom of the vial may not be efficiently measured. These solids which contained uranium and other metal debris were further collected by the filter membrane. Thus, the uranium and other metal debris present in original water were underestimated. The radioactivity attributed
to suspended solids was not significant, but the turbidity was significant. In Fig. 2(a), the gold suspended solids are a part of the ion-exchange resins. The morphology and elements on the surface of the suspended solids were observed using SEM-EDS. The suspended solids were composed of fine, spherical particle agglomerates with the particles ranging from 0.5 to >1 m in size (Fig. 2(b)). The compositions of the suspended solids, as determined by EDS, are listed in Table 3. The elements, such as C, O and S, are derived from ion-exchange resins. Due to over 20 years of radiolysis, part of the spent resins may be degraded and crushed under highly radioactive conditions. The specific gravity of the spent resins (1.09–1.20) was notably close to that of water, so it was difficult to naturally precipitate the crushed, spent resins colloids. 3.2. Water treatment 3.2.1. Efficiency of contaminants removal in the Trial I system In the Trial I system, filters with different particle rejection pore sizes (5, 0.9, and 0.1–0.01 m) and cation-exchange resin were employed to remove suspended solids and radioactive ions from the TRR SFP water. The turbidity of the first batch original water was 3.63 NTU, but it decreased to 1.76, 0.9, and 0.48 NTU after passing through the PP, ceramic, and hollow fiber filters, respectively (data not shown). The removal capabilities of -emitters, ␣-emitters, and U for Trial I system are shown in Fig. 3 and the removal efficiencies of each unit are listed in Table 4. The removal efficiencies in Table 4 are calculated with respect to the previous residual activities, not the initial activities. The turbidity was slightly decreased, but only a few radioactive species were blocked by the PP filter. Interestingly, after passing through the ceramic depth filter, most of the 137 Cs, 90 Sr, -emitters, and ␣-emitters were trapped (Fig. 3(a) and (b)). More than 95% of -emitters and ␣-emitters were removed by the filter, but only 55% of U was removed (Fig. 3(c) and Table 4). Therefore, it was determined that the ceramic filter had dual purification
Table 3 SEM-EDS analysis of suspended solids.
Fig. 2. (a) Appearance and (b) SEM-image of suspended solids in the TRR SFP.
Element
Weight%
C O Mg Al Si S Cl K Ca Fe U Totals
23.92 34.46 0.86 6.24 12.59 1.17 0.75 2.32 0.88 10.28 6.54 100.00
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Table 4 The removal efficiencies of radionuclides by a specific purification unit. Unit
137 Cs efficiency (%) (residual, Bq mL−1 )
90 Sr efficiency (%) (residual, Bq mL−1 )
Gross ˇ efficiency (%) (residual, Bq mL−1 )
Gross ˛ efficiency (%) (residual, Bq mL−1 )
U efficiency (%) (residual, mg L−1 )
PP filter (5 m)
2.6 (1.12 × 103 ) 96.6 (3.76 × 101 ) 90.9 (3.42 × 100 ) 84.4 (5.30 × 10−1 ) 99.7 99.1 (1.74 × 101 ) 71.2 (5.01 × 100 ) 78.6 (1.07 × 100 ) 99.9
2.3 (6.35 × 101 ) 93.6 (4.10 × 100 ) 77.0 (9.30 × 10−1 ) 18.7 (7.50 × 10−1 ) 98.8 28.9 (2.29 × 102 ) 46.3 (1.23 × 102 ) 99.3 (8.90 × 10−1 ) 99.7
0.7 (1.51 × 103 ) 96.9 (4.74 × 101 ) 82.2 (8.42 × 100 ) 46.6 (4.50 × 100 ) 99.7 77.0 (5.84 × 102 ) 46.7 (3.11 × 102 ) 96.7 (1.02 × 101 ) 99.6
1.9 (2.65 × 101 ) 97.8 (5.80 × 10−1 ) 71.4 (1.70 × 10−1 ) 40.1 (1.00 × 10−1 ) 99.6 54.0 (2.16 × 101 ) 29.2 (1.53 × 101 ) 95.1 (7.50 × 10−1 ) 98.4
3.5 (1.93) 54.9 (0.87) 88.5 (0.10) – (0.10) 95.0 29.3 (3.11) 20.6 (2.47) 57.1 (1.06) 75.9
Ceramic filter (0.9 m) Cation exchange resin Hollow fiber filter Overall Trial I 1st filter (5.0 m) 2nd filter (0.9 m) 3rd filter (0.22 m) Overall Trial II
capabilities:particles rejection and ion adsorption. Ionic radionuclides and fine particles were further removed by the treatment of ion-exchange resin and hollow fiber filter, respectively (Fig. 3). More than 99% of -emitters and ␣-emitters were removed by a Trial I system (Table 4). 3.2.2. Efficiency of contaminants removal in the Trial II system In the Trial II system, three cascaded ceramic filters with different pore sizes were used to investigate the major particle size distributions of the suspended solids. Additionally, the capability of ceramic filters to remove particles and ionic radionuclides was investigated. The second batch of original water contained much higher concentrations of radioactive ions and suspended solids than the first batch. The turbidity of the original water was 4.37 NTU, which was decreased to 2.3, 1.9, and 0.8 NTU by each of the ceramic filters, respectively (data not shown). The removal efficiencies of -emitters, ␣-emitters, and U by the Trial II system are shown in Fig. 4 and Table 4. Based on the results of Trial I, the radionuclides, 137 Cs and 90 Sr, cannot be removed by the filter with a 5.0-m rejection pore. However, approximately 99% and 29% of 137 Cs and 90 Sr were respectively removed by the first filter (5.0 m) via adsorption mechanism (Table 4). The diatomitebased ceramic filter tended to have higher adsorption efficiency of 137 Cs than 90 Sr (Fig. 4(a)). Hence, the diatomite-based ceramic filter exhibits adsorption selectivity on different ions. The second filter (0.9 m) in Trial II was found to have lower removal efficiencies of 90 Sr, -emitters, ␣-emitters, and U than observed in Trial I system due to the higher initial concentrations of radioactive ions and suspended solids. Most of the 90 Sr present in the water are ions, and the removal efficiency of 90 Sr will decrease as the initial concentration increase. On the other hand, the suspended solids with different sizes and compositions were separated by different pore sizes of the filters. Only few amounts of suspended solids can be removed by the first filter. However, some of the suspended solids with sizes ranging from 5 to 0.9 m may contain lower concentrations of -emitters, ␣-emitters, and U, but these species will occupy the adsorption sites of the second filter and lead to the decrease of adsorption capacity. The residual fine suspended solids (<0.9 m) contain higher concentrations of -emitters, ␣-emitters, and U can be further removed by the third filter (Fig. 4(b) and (c)). 3.2.3. Contaminant removal capability of the ceramic filter Five batches of original water (totaling 50 L) were treated using the Trial II system without replacing the filters. The accumulations of activities on the filters were plotted with the total feed-in volume (Fig. 5). After treating 30 L of the water, the
accumulation of -emitters (a major was 137 Cs) on the first filter almost reached a plateau, which implied adsorption saturation (Fig. 5(a)). The adsorption tendency of the 137 Cs was similar to that of the -emitters (data not shown). As the first filter reached saturation, the removal of residual contaminants depended on the second and third filters. Moreover, due to the higher particle rejection capability of the second filter, the accumulation of -emitters on the second filter was found to be higher than the first filter. Because the first and second filters removed most of the -emitters, the loading of the third filter was relatively low. After the treatment of 50 L of the water, the overall removal efficiencies of the -emitters and ␣-emitters were greater than 95% and 91%, respectively (Fig. 5(a) and (b)). It is well-known that the materials with larger specific surface areas will have more adsorption sites. In a comparison of the three filters, which were composed of different particle sizes of raw diatomite, the third filter should have the largest specific surface area and the greatest adsorption capacity. Furthermore, the fine suspended solids which have higher concentration of U can be effectively filtrated by the third filter, so the third filter captured more U than the first filter, as observed in Fig. 5(c). On the other hand, the adsorption efficiency of ions was proportional to the numbers of adsorption sites of the filter. The molar concentration of U (∼20 M) in the original water was much greater than the other radionuclides (e.g., 137 Cs ∼6 × 10−3 M), so the numbers of active adsorption sites will have obviously effect on adsorption efficiency. During the filtration, more and more suspended solids deposited on the surface of filters and occupied the adsorption sites, so the removal efficiency of U gradually decreased. Because the water in the TRR SFP had a complicated composition with various radionuclides in ionic and solid form, distinguishing the filtration or adsorption effects was difficult. After treating 50 L of the water, the volumetric flowrate of the system decreased to 0.33 L min−1 (only one third of the initial flow, data not shown), and the primary fouling occurred to the third filter. Furthermore, the specific ˛ activity for the first filter was 5120 Bq g−1 , which is classified as TRU waste. However, the further conditioning and storage of spent ceramic filters was relatively simple and safer than conditioning and storing spent organic purification media. During the treatment of water, the surface dose rate of each filter was gradually increased, and the major radiation dose rate was contributed by the 137 Cs (/␥-emitter). The maximum surface dose rate of the first filter was 863 Sv/h; therefore the replacement or disposal of the spent filter could be contact-handled (<2 mSv/h).
C.-P. Huang et al. / Journal of Hazardous Materials 233–234 (2012) 140–147
Fig. 3. The removal of (a) -emitters, 137 Cs, and 90 Sr (b) ␣-emitters, and (c) U from the water in treatment Trial I. The original activity concentrations of -emitters, 137 Cs, 90 Sr, and ␣-emitters are shown as 1520, 1150, 65, and 27 Bq mL−1 , respectively; initial concentration of U is 2 mg L−1 ; pH = 5.48; conductivity = 30.3 S cm−1 .
145
Fig. 4. The removal of (a) -emitters, 137 Cs, and 90 Sr (b) ␣-emitters, and (c) U from the water in treatment Trial I. The original activity concentrations of -emitters, 137 Cs, 90 Sr, and ␣-emitters are shown as 2540, 2010, 322, and 47 Bq mL−1 , respectively; initial concentration of U is 4.4 mg L−1 ; pH = 4.98; conductivity = 42.0 S cm−1 .
80 70
40
100
80 60 50
60
40
30 20
20
Overallremoval efficiency (%)
(a) -emitters (MBq)
C.-P. Huang et al. / Journal of Hazardous Materials 233–234 (2012) 140–147
Accumulation of
146
Table 5 Studies of radionuclides removal by diatomite materials. Radionuclide
Efficiency
Removal
Form of diatomite
Reference
137
Cs, 134 Cs, 60 Co 137 Cs 90 Sr Gross  Gross ␣
85% 94% 94% 95% 91%
2.21 MBq Powder (60 mesh, 50 kg) 100 MBq Filter candles (750 g) 19.5 MBq 149 MBq 2.61 MBq
[21] This study
U
43% 90%
132 mg 6.10 mg 159 mg
This study [22]
10
Filter candles (750 g) Powder (0.25 mm, 1 g) Powder (modified, 1 g)
0
0 0
20
40
60
100
80
120
140
160
4. Conclusions
Feed-in volume (L) 100
(b) 1.4
80 1.0 60
0.8 0.6
40
0.4 20
Overallremoval efficiency (%)
Accumulation -emitters (MBq)
1.2
0.2 0
0.0 0.0
0.5
1.0
1.5
2.0
2.5
3.0
Acknowledgements
Feed-in volume (L)
(c) 70 60
Accumulation of U (mg)
80 50 60
40 30
40
20 20
Overall removal efficiency (%)
100
10 0
0 0
50
100
15 0
200
To meet the acceptance criteria of INER’s LRWs treatment facilities, the removal of radioactive solids and ions in TRR SFP water is achieved by both trial systems. Based on the analyses, the radioactive solids were quite complex, they were primary attributed by the ruptured spent fuel, spent resins, and metal debris. Most of the radioactive solids could be removed by the filters with less than 0.9 m of rejection pores. Most of the radioactive ions, 137 Cs, 90 Sr, and U, in TRR SFP water could be removed by cation exchange resin and diatomite-based ceramic filter. The removal of U by cation exchange resin was found to be more efficient than diatomitebased ceramic filter. The diatomite-based ceramic filters exhibited excellent removal efficiency on the particles filtration and ions adsorption. Using the candle type diatomite filters for the treatment of TRR SFP water is feasible.
250
300
350
Feed-in volume (L)
Fig. 5. The activity accumulation and removal efficiency () of (a) -emitters, (b) ␣-emitters and (c) U with total feed-in mass of contaminants for ceramic filters with different pore sizes 5.0 m (䊉), 0.9 m () and 0.22 m ().
The studies on the radionuclides removal using diatomite materials have been summarized in Table 5. Three candle type diatomite filters (750 g) were employed to remove the radioactive ions and suspended solids in this study. The activity concentrations of radionuclides present in the TRR SFP water were found to be much higher than the other work [21], and more than 90% of 137 Cs, 90 Sr, emitters, and ␣-emitters were removed in this study. The removal of U by the filters was probably retarded by the suspended solids, so the removal efficiency is relative lower than the other work [22].
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