- Email: [email protected]
Sequential determination of Pu and Am radioisotopes in environmental samples; a comparison of two separation procedures
ARTICLE IN PRESS
Applied Radiation and Isotopes 65 (2007) 504–511 www.elsevier.com/locate/apradiso
Sequential determination of Pu and Am radioisotopes in environmental samples; a comparison of two separation procedures Rozˇle Jakopicˇ, Polona Tavcˇar, Ljudmila Benedik Department of Environmental Sciences, Jozˇef Stefan Institute, Jamova 39, 1000 Ljubljana, Slovenia Received 29 May 2006; received in revised form 28 November 2006; accepted 13 December 2006
Abstract Two separation methods for the sequential determination of Am and Pu radionuclides are presented and the results obtained are compared. Analysis involves leaching the sample with concentrated nitric acid (HNO3), followed by radiochemical separation using extraction chromatographic resins (UTEVA, TRU) and anion exchange. Sources for alpha spectrometry were prepared by microprecipitation on neodymium fluoride (NdF3). The chemical recoveries were determined using 242Pu and 243Am tracers. The methods were tested on reference materials and on two sediments. All the results were in good agreement with the reference values. The evaluation of uncertainty is also included. r 2007 Published by Elsevier Ltd. Keywords: Actinides; Extraction chromatography; Anion exchange; Micro-precipitation technique; Alpha spectrometry; Uncertainty
1. Introduction Man-made radionuclides are released into the environment as a result of nuclear activities, atomic weapons testing and accidents at nuclear power plants. Of the many different man-made radionuclides, 241Am, 239Pu, 240Pu and 238 Pu are alpha emitting radionuclides that need special attention. These radionuclides are important long-term radioactive pollutants due to their long half-lives and long persistence in the environment. Due to increased energy production by nuclear reactors, reprocessing and waste disposal, the increased potential for contamination and public concern over the potential hazards, the accurate and rapid determination of artificial radionuclides in the environment is very important. Analysis of alpha emitting man-made radionuclides is complex and requires several steps. One of the important steps in such analysis is separation prior to measurement. The goal of separation is to eliminate all possible radioactive and non-radioactive interferences that could otherwise degrade the spectrum or overlap the peaks of interest. Corresponding author. Tel.: +386 1 5885 201; fax: +386 1 5885 346.
E-mail address: [email protected] (R Jakopicˇ). 0969-8043/$ - see front matter r 2007 Published by Elsevier Ltd. doi:10.1016/j.apradiso.2006.12.005
Several investigators have reported different procedures for separation and determination of man-made radionuclides. In the past, radiochemical separations were usually made by liquid–liquid extraction or ion exchange. The separations were tedious, time-consuming and generated a large amount of wastes. In the last 10 years, extraction chromatography has become a very attractive method for separation of the actinides in different environmental and biological samples (Horwitz et al., 1993; Pilvio¨ and Bickel, 1998; Mellado et al., 2001, 2002; Alvarez and Navarro, 1996; Kaye et al., 1995). Simple and fast separations with high chemical yield and good selectivity are achieved. The reason for the good selectivity lies in the organic extractant which is bound on an inert support and selectively retains certain radionuclides. In this paper, we compared separation techniques for plutonium and americium radionuclides using ion exchange (Dowex 1X8) and extraction chromatography using TRU and UTEVA resins (Eichroms Industries) in soil samples and sediments. For extraction chromatography we used a method developed by Toribio et al. (2001) which we slightly modified to fit our purposes. We used leaching in concentrated nitric acid (HNO3) for sample preparation instead of total digestion in a microwave oven and an
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
additional precipitation step on calcium oxalate (CaC2O4) was introduced prior to loading the sample solution onto the column. According to previous studies in our laboratory (Benedik et al., 1999), man-made radionuclides are usually adsorbed superficially and therefore leaching instead of total dissolution may be used. For ion exchange chromatography we used a method developed by Rubio Montero et al. (2000) with a modification in sample preparation by leaching in concentrated HNO3 and using 9 M hydrochloric acid (HCl) to remove thorium from the column. For the reduction of plutonium from Pu(IV) to Pu(III), hydroiodic acid was replaced with ammonium iodide (NH4I). Americium determination was also included in the separation scheme. All the radionuclides were measured by alpha spectrometry and the sources were prepared by micro-precipitation on neodymium fluoride. The results obtained for both procedures were compared in terms of activities and recoveries. 242Pu and 243Am tracers were used for the determination of the recovery. In the final presentation of an analytical result, it is very important to evaluate measurement uncertainty. In the past sometimes the quality of the results was defined by the standard deviation of repeated measurements, or was not even reported (Drolc and Rosˇ , 2002). In a radioactivity measurement the standard deviation is given by the square root of the number of counts (L’Annunziata, 1998), but this is only the statistical uncertainty of counting. Measurements uncertainty must also include other sources of uncertainty such as sample collection, tracer activity concentration, detector efficiency, chemical recovery, etc. The evaluation of uncertainty requires the analyst to look closely at all stages of the method and at all possible sources of uncertainty (Ellison et al., 2000). 2. Experimental 2.1. Samples Reference materials IAEA-135 (Radionuclides in Irish Sea Sediment), IAEA-368 (Pacific Ocean Sediment), IAEA-300 (Radionuclides in Baltic Sea Sediment), IAEA375 (Radionuclides in Soil), Soil-6 (Radionuclides in Soil), Ribble Sediment (sediment from the tidal zone of the river Ribble, Lancashire, UK) and intertidal sediment from the Cumbrian coast, UK. Ribble sediment is influenced both by tidal-borne radionuclides originating from Sellafield and the manufacture of fuel elements at Springfields upstream. 2.2. Reagents and materials All reagents used in the analysis were of analytical grade. The following reagents were used for leaching and for preparation of the solutions used in the radiochemical separation: 65% HNO3, 36% HCl, 40% HF, 32% H2O2, 25% ammonia (NH3), CH3OH, C2H5OH, NH4SCN,
505
NH4I, H2C2O4 2H2O, ascorbic acid (powder form), NaNO2, NH2OH HCl, TRU Spec and UTEVA Spec Resins (Eichroms Industries, 100–150m particle size resin), anion exchange resin (Dowex 1X8, 100–200 mesh). The following reagent solutions were prepared:
0.6 M ferrous sulphamate—Fe(NH2SO3)2 About 28.5 g of sulphamic acid (NH2SO3H) was added to 75 mL of deionised water, heated and 3.5 g of powdered iron was added and left to dissolve. Then the solution was filtered through a 0.2 mm filter and diluted to 100 ml with deionised water. The solution was prepared fresh weekly. 0.1 M ammonium hydrogen oxalate—NH4HC2O4 About 6.31 g of oxalic acid dihydrate (H2C2O4 2H2O) and 7.11 g of ammonium oxalate hydrate ((NH4)2 C2O4 H2O) were dissolved in 900 mL of deionised water and diluted to 1 L with deionised water. 10 mg m L1 neodymium fluoride substrate solution— NdF3 About 4.2 mg of neodymium (III) oxide (Nd2O3) was dissolved in 454 mL of 1 M HNO3 in a polyethene bottle and then 48 mL of 40% hydrofluoric acid was added.
2.3. Tracers 242
Pu standardised solution with an activity concentration of 1.06 Bq g1 (the expanded uncertainty 0.8%, k ¼ 2) and 243Am standardised solution with an activity concentration of 0.746 Bq g1 (the expanded uncertainty 1.89%, k ¼ 2) were purchased from the National Physical Laboratory, Teddington, Middlesex, UK. 2.4. Instrumentation For the alpha spectrometric measurements, a silicon surface barrier detector (PIPS) with an active area of 450 mm2 and counting efficiency of 31%71% connected to an EG&G ORTEC Maestro Gammavision MCA Emulator multichannel analyser system was used. The counting efficiency of the alpha detector was determined by measuring the activity of a 226Ra alpha source prepared by micro-precipitation (SRM 4967, activity concentration 2729 Bq g1, expanded uncertainty 1.18%, k ¼ 3). 2.5. Sample preparation The samples were dried to constant weight at 105 1C. After drying, the samples were ashed to decompose organic matter in an electric muffle furnace at 550 1C for 20 h. Approximately 1–2 g of ashed sample together with 242Pu and 243Am tracers were leached with 100 mL hot concentrated HNO3 with magnetic stirring overnight. During the leaching the beaker was covered with a watch glass to prevent significant evaporation. After cooling the solution to room temperature, the leachant and the residue were separated by filtration.
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
506
2.6. Separation procedure Methods A and B are schematically presented in Fig. 1. 2.6.1. Method A (extraction chromatography) The filtrate from sample preparation in 2.5 was heated until boiling and approximately 10 g of oxalic acid (H2C2O4 2H2O) was added. The pH of the solution was adjusted to 5.5–6.0 with NH3 solution. The actinides were co-precipitated on CaC2O4, while the iron remained mainly in solution as the oxalate complex (Moreno et al., 1997). For samples that have a low calcium content, additional calcium in the form of calcium chloride (CaCl2) was added to carry down actinides in the CaC2O4 precipitation step. The solution was cooled and the oxalate precipitate was centrifuged for 10 min at 2500 rpm and washed several times with deionised water. The oxalate precipitate was fumed twice with concentrated HNO3 and 32% hydrogen peroxide (H2O2) and evaporated to dryness. The residue was dissolved in 3 M HNO3 and then Fe(NH2SO3)2 was added to reduce plutonium to the trivalent oxidation state. The content of Fe(III) was checked by adding one drop of
1 M ammonium thyocyanate (NH4SCN) solution to 1 drop of sample solution. If the reaction was positive, then up to 200 mg of ascorbic acid (C6H8O6) was added to keep iron in the +2 valency state. Sequential separation of plutonium and americium from the sample solution was performed using UTEVA and TRU resins (100–150m, Eichroms Industries). About 3 g of UTEVA or TRU resin material were mixed with deionised water and used to fill a chromatographic column (1 cm internal diameter, 20 cm long, obtained from Bio-Rad Laboratories). The UTEVA column was conditioned with 60 mL of 3 M HNO3. The sample solution was loaded onto the column, followed by washing with 60 mL 3 M HNO3. Plutonium and americium radioisotopes were eluted from the column, while uranium and thorium radioisotopes were adsorbed and therefore removed. The effluent containing plutonium and americium was evaporated to dryness and redissolved in 2 M HNO3. The solution was loaded onto the TRU column, which was conditioned with 2 M HNO3. The TRU column was washed with 60 mL of 2 M HNO3–0.1 M sodium nitrite. Sodium nitrite (NaNO2) was used to oxidise Pu(III) to Pu(IV) to allow Pu and Am separation. After converting
sample (up to 2 g) + 242Pu and 243Am
Method A
co-precipitation on CaC2O4
leaching with conc. HNO3 and filtration
Method B
evaporation to dryness, dissolution in 1M HNO3,
oxalate destruction,
oxidation state adjustment,
dissolution in 3M HNO3,
load solution in 8M HNO3
oxidation state adjustment Dowex 1X8 column (1) UTEVA column (U, Th)
3M HNO3 (Pu, Am) evaporation to dryness,
8M HNO3 (Am, Fe)
9M HCl (Th, discard)
dissolution in 2M HNO3 9M HCl/0.1M NH4I (Pu) TRU column evaporation to dryness, 2M HNO3/0.1M NaNO2 (discard)
9M HCl, 4M HCl (Am)
0.1M NH4HC2O4 (Pu)
dissolution in 9M HCl
Dowex 1X8 column (2)
9M HCl Am → purification
Fig. 1. Analytical procedures for the sequential determination of plutonium and americium radioisotopes.
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
the column to the chloride system by means of 9 M HCl, americium was stripped with 60 mL of 4 M HCl and plutonium with 60 ml of 0.1 M NH4HC2O4. The approximate rates of loading and stripping solutions were 1 mL min1. 2.6.2. Method B (anion exchange) The filtrate from sample preparation in 2.5 was evaporated to dryness and the residue was dissolved in 1 M HNO3. About 1 g of hydroxylamine hydrochloride (NH2OH HCl) was added and the solution was heated to reduce plutonium to Pu(III). The completeness of the reaction was checked with a drop of 1 M ammonium thyocyanate solution. The solution was cooled and made 8 M in HNO3 by adding concentrated HNO3. Then 0.5 g sodium nitrite was added to ensure conversion of Pu(III) to Pu(IV). The solution was boiled to destroy excess hydroxylamine hydrochloride, cooled and transferred to the top of an anion exchange column of height 10 cm, diameter 1 cm (Dowex 1X8, 100–200 mesh), which was conditioned with 60 ml of 8 M HNO3. After washing Am and Fe from the column with 60 mL of 8 M HNO3 and Th with 100 ml of 9 M HCl, Pu radioisotopes were stripped with 60 mL freshly prepared 9 M HCl–0.1 M NH4I solution. The iodide reduced Pu(IV) to Pu(III) which was not retained by the column. The Am in the 8 M HNO3 effluent solution was evaporated to dryness and dissolved in 9 M HCl. The solution was loaded onto a second anion exchange resin column, which was prepared in 9 M HCl. Iron was retained, while Am passed through the column. The Am fraction was further purified from lanthanides. The solution was evaporated to dryness and dissolved in 1 M HNO3–methanol solution and loaded onto an anion exchange resin column. The column was washed with 60 mL of 1 M HNO3–methanol and 60 mL of 0.1 M HCl–0.5 M ammonium thyocyanate–80% methanol. Prior to stripping Am radioisotopes from the column with 100 mL of 1.5 M HCl–86% methanol, a wash with 30 mL 1 M HNO3–methanol was performed (Bains and Warwick, 1993). 2.7. Source preparation Alpha-counting sources for this work were prepared by neodymium fluoride micro-precipitation (Hindman, 1983). Solutions eluted from the columns were evaporated to dryness and fumed twice with 1 mL concentrated HNO3 and 1 mL 32% H2O2 to destroy NH4I and any organic compounds (resin) that might have been present in the eluate. Then 2 mL of 1 M HNO3 was added and the solution was transferred to a polypropylene centrifuge tube. About 50 mg Nd3+ solution and 0.5 mL 40% hydrofluoric acid were added, swirled and allowed to stand in an ice bath. After 30 min the neodymium fluoride suspension with plutonium radioisotopes was filtered through a 25 mm, 0.1 mm membrane filter (Supors Membrane Disc Filters, Gelman Laboratory) pre-washed
507
with 80% ethanol (C2H5OH) solution, deionised water and conditioned with 5 mL of neodymium substrate solution. The centrifuge tubes were rinsed with 5 ml of 0.58 M hydrofluoric acid solution, twice with deionised water and 5 mL 80% ethanol solution and the rinse solutions filtered. The filters were dried, mounted on an aluminium planchet and measured by alpha spectrometry. 3. Results and discussion The results obtained by sequential determination of plutonium and americium radionuclides are presented in Table 1 and in Figs. 2–8. The individual and the average chemical recoveries for each sample are presented in Table 1 and Figs. 7 and 8. Chemical recoveries are higher for method A than for method B, as shown in the bar chart in Figs. 7 and 8. Furthermore, the recoveries for plutonium radioisotopes are generally higher than for americium. For Pu the average recoveries in all samples used for method A are around 60% and for method B 42%. For americium the average recoveries are around 57% for method A and 43% for method B, respectively. For some samples (IAEA 300, IAEA 368, Ribble sediment), higher dispersion in recovery values was observed between replicates. The lowest chemical recoveries were obtained for americium radioisotopes in Cumbrian and Ribble sediment. Both these sediments have a higher content of iron than other samples, which was probably not removed completely in the preconcentration step and caused pre-elution of americium or interfered with adsorption of americium onto the TRU column. However, further work should be done to investigate this problem so as to allow higher recoveries. From Table 2 it can be seen that most of the results obtained by the two methods are in good agreement with the reference or literature values and previous work in our laboratory (Tavcˇar et al., 2005). For plutonium and americium radioisotopes in Cumbrian sediment lower results were observed; on the other hand, higher values were observed in Ribble sediment. In the case of 238Pu in Soil-6, the result was under the detection limit due to the low available mass of the sample. From the results in Table 2, it can also be concluded that leaching in this case for the determination of plutonium and americium radioisotopes was complete and therefore there was no need for total dissolution of the sample, suggesting negligible refractory artificial actinides in the samples. Leaching in concentrated HNO3 is also faster and simpler compared to total dissolution, thus saving time and chemicals. In Figs. 9 and 10, alpha spectra of Am radioisotopes isolated from IAEA-135 reference sample and Pu radioisotopes isolated from Ribble sediment are presented. In the case of the plutonium spectrum from IAEA-135 in Fig. 9, three peaks are observed, one belonging to tracer 242 Pu which was added at the beginning of the procedure to monitor the losses of plutonium during separation (leaching, radiochemical separation, source preparation) and
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
508
Table 1 Individual and average chemical recoveries for Pu and Am radioisotopes Sample: Pu
Method A
Method B
Individual Pu recovery (%)
Average Pu recovery (%)
Individual Pu recovery (%)
Soil-6 IAEA 135 IAEA 300 IAEA 368 IAEA 375 Cumbrian s. Ribble s.
63 46 78 74 60 62 78
7178 4473 62713 66711 5277 5778 64719
43 36 40 41 53 32 50
Sample: Am
Method A
Soil-6 IAEA 135 IAEA 300 IAEA 368 IAEA 375 Cumbrian s. Ribble s.
78 42 49 58 49 51 51
73 66
56
46
44 43 47 52 53 31 48
40
38
50 41 51
51
40
4173 4075 4675 4576 5271 3271 4675
Method B
Individual Am recovery (%)
Average Am recovery (%)
Individual Am recovery (%)
62 61 68 81 50 50 37
5874 6879 6077 71715 4973 5273 4176
29 45 50 47 46 28 30
57 74 63 60 51 54 45
Average Pu recovery (%)
55 54
53
46
40 55 74 58 45 24 27
Average Am recovery (%)
55
42
41 41 42
43
36
42711 5077 55717 4979 4472 2673 3175
The uncertainties correspond to the standard deviation of the average at the 71 s level.
239/240Pu
241Am
reference value method A method B previous work
Bq/kg
1.5 1
reference value method A merhod B previous work
4 Bq/kg
2
5
3 2 1
0.5
0
0
IAEA-300
IAEA-300
IAEA-368
IAEA-375
Soil-6
sample Fig. 2. A comparison of activity concentration in Bq kg IAEA-300, IAEA-368, IAEA-375 and Soil-6.
1
of
241
Am in
IAEA-375 sample
Soil-6
Fig. 4. A comparison of activity concentration in Bq kg1 of IAEA-300, IAEA-375 and Soil-6.
239/240
Pu in
239/240Pu 241Am
1200 800
reference value method A method B previous work
1000 800 Bq/kg
1600
Bq/kg
1200 reference value method A method B previous work
600 400
400
200
0
0 IAEA-135
Cumbrian Sediment
Ribble Sediment
sample
Fig. 3. A comparison of activity concentration in Bq kg1 of Cumbrian, Ribble sediment and IAEA-135.
241
Am in
Cumbrian Sediment
Ribble IAEA-368 Sediment sample
IAEA-135
Fig. 5. A comparison of activity concentration in Bq kg1 of IAEA-135, IAEA-368, Ribble and Cumbrian sediment.
239/240
Pu in
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
of radioactive wastes resulting from fuel fabrication and reprocessing activities.
238Pu
250
reference value method A method B previous work
Bq/kg
200 150 100 50
3.1. Evaluation of measurement uncertainty The procedure for evaluation of the uncertainty of the result of 239/40Pu activity concentration is presented below. The activity concentration of 239/40Pu is calculated from Eq. (1)
0 Cumbrian Sediment
IAEA-135 Ribble Sediment sample
IAEA-368
Fig. 6. A comparison of activity concentration in Bq kg1 of Cumbrian, Ribble sediment, IAEA-135 and IAEA-368.
90 method A
238
Pu in
method B
80 Pu recovery (%)
70 60 50 40
20 10 0 Soil-6
Cumbrian Ribble Sediment Sediment
Sample
Fig. 7. Pu recoveries in various reference materials obtained by two methods.
90
method A
method B
80 Am recovery (%)
A239=40 Pu ¼
P239=40 Pu ; Zchem det ms t
(1)
where A239=40 Pu is the activity concentration of 239/40Pu in Bq kg1, P239=40 Pu is the peak area of 239/40Pu, Zchem is the chemical recovery, edet is the detector efficiency, t is the time of measurement in s and ms is the sample mass in kg. The uncertainty associated with the mass u(ms) of the sample was estimated using data from the calibration certificate and the analytical balance manufacturer’s recommendations on uncertainty estimation as 0.1 mg. We assumed a rectangular distribution and converted to a normal distribution by dividing by 31/2 Eq. (2) 0:1 mg uðms Þ ¼ pffiffiffi ¼ 0:058 mg: 3
30
IAEA-135 IAEA-300 IAEA-368 IAEA-375
509
(2)
The uncertainty in chemical recovery u(Zchem) for 239/40Pu consists of the uncertainty of the tracer activity contribution uðA242 Pu Þ, the peak area of plutonium tracer uðP242 Pu Þ, the mass of plutonium tracer added to the sample uðm242 Pu Þ and the detector efficiency u(edet) and it is calculated from Eq. (3) uðZchem Þ ¼ Zchem sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi uðm242 Pu Þ 2 uðZdet Þ 2 uðA242 Pu Þ 2 uðP242 Pu Þ 2 . þ þ þ m242 Pu Zdet A242 Pu P242 Pu ð3Þ
70 60
The uncertainty of the tracer activity uðA242 Pu Þ was taken from the certificate. The uncertainty of the peak area uðP242 Pu Þ was calculated as the square root of the peak area pffiffiffiffiffiffiffiffiffiffiffiffi (4) uðP242 Pu Þ ¼ P242 Pu .
50 40 30 20 10 0 IAEA-135
IAEA-300
IAEA-368
IAEA-375
Soil-6
Cumbrian Sediment
Ribble Sediment
Sample
Fig. 8. Am recoveries in various reference materials obtained by two methods.
238
239/240
other two to Pu and Pu present in the sample. The doublet peak of 239/240Pu cannot be resolved by alpha spectrometry due to closeness of the alpha energies of 239 Pu and 240Pu. In the Am spectrum isolated from Ribble sediment in Fig. 9, two peaks are observed, one of 243 Am tracer and other of 241Am, which is present in the sample. The highest activities were observed in Ribble and Cumbrian sediments. This was expected since both sediments originated from areas contaminated by disposal
The uncertainty in detector efficiency consists of the uncertainty of 226Ra activity uðA226 Ra Þ, uncertainty of peak area uðP226 Ra Þ, mass uðm226 Ra Þof 226Ra added and the recovery of micro-precipitation u(Zmicro) and is calculated from Eq. (5) det ¼
P226 Ra . A226 Ra Zmicro m226 Ra t
(5)
In the determination of 239/40Pu, the combined uncertainty is calculated from Eq. (6) uðA239=40 Pu Þ ¼ A239=40 Pu sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi uðP239=40 Pu Þ 2 uðZchem Þ 2 uðdet Þ 2 uðms Þ 2 . þ þ þ P239=40 Pu Zchem det ms ð6Þ
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
510
Table 2 Activity concentrations in Bq kg1 dry weight for selected Pu and Am radioisotopes Sample
Activity concentration (Bq kg1)
IAEA 135
238
Isotope Pu
239/240
Pu Am
241 238
Soil-6
Pu
239/240
Pu 241 Am IAEA 300
238
Pu
239/240
Pu 241 Am
IAEA 368
238
Pu
239/240
Pu Am
241
IAEA 375
238
Pu
239/240
Pu 241 Am Cumbrian sediment
238
Pu
239/240
Pu Am
241
Ribble sediment
238
Pu
239/240
Pu 241 Am
This work Method A
This work Method B
4274(2) 209713 321720
3974(2) 215715 291715
o 0.2(3) 1.0070.11 0.4070.09
o0.2(4) 1.0770.10 0.3670.07
Previous work (Tavcˇar et al., 2005)
Reference or literature value 43 (42–45)a 213 (205–226)a 318 (310–325)
4674 217712 313714 o 0.2 1.0370.16 0.3970.08
o0.036 1.0470.07 0.39–0.45b 0.15 3.55 (3.44–3.65)a 1.38 (1.2–1.5)a
0.1270.04(4) 3.6270.30 1.3570.16
0.1370.03(3) 3.4770.36 1.2170.12
o 0.2 3.4970.23 1.2170.18
9.370.7(2) 3072 1.370.3
8.770.8(3) 3272 1.270.2
8.271.0 3175 1.470.2
8.570.9a 3172a 1.3 (1.21.5)
0.10470.047(3) 0.4070.04 0.1270.02
0.09070.020(4) 0.3770.05 0.1070.02
0.08070.030 0.3470.06 0.1570.03
0.07170.014 0.3070.04 0.1370.07
196710(2) 917751 1474791
194714(2) 878753 1430782
193719 930774 1508799
21174c 995719c 1590713c
3372(2) 186712 300732
3272(3) 195713 311721
3273 187710 288723
3375 16178 283–351d
Pu is decay corrected to the reference date. Uncertainties are given at 71 s. Number of determinations are given in parentheses. Certified value. b Moreno et al. (1997). c Adsley et al. (1998). d Benedik et al. (1999).
238
a
1360
220 241Am 243Am
tracer
110
Pu
680 242
Pu tracer
238
340
55
3880.00
239/240
1020 Counts
Counts
165
4497.00
5028.00 Energy (keV)
5473.00
5831.00
2951.00
4309.00
Pu
5667.00
7024.00
8382.00
Energy (keV)
Fig. 9. Alpha spectrum of americium radioisotopes isolated from IAEA135 reference material.
Fig. 10. Alpha spectrum of plutonim radioisotopes isolated from Ribble sediment.
Standard measurement uncertainty and relative standard uncertainty in the determination of 239/40Pu in IAEA 368 are summarised in Table 3 and schematically presented in Fig. 11.
two different radiochemical procedures and measured by alpha spectrometry. Thin sources for alpha measurements were prepared by the micro-precipitation technique. The results demonstrated the usefulness of a simultaneous procedure and therefore a shorter time for the analysis. The major source of uncertainty in the result of the measurement was identified as the contribution from detector efficiency uncertainty, chemical recovery uncertainty and uncertainty in peak area.
4. Conclusion Plutonium and americium radioisotopes in various samples were determined in the same sample aliquot using
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511 Table 3 Standard measurement uncertainty and relative standard uncertainty in the determination of 239/40Pu activity concentration in IAEA 368 Value x
Standard Relative standard measurement measurement uncertainty u(x) uncertainty u(x)/x
environment, mass balances and modelling of the environmental processes and risk analysis). The authors wish to express their thanks to Dr. P.J. Day, University of Manchester, who kindly supplied the Ribble Sediment sample. References
Detector efficiency 0.1149 g m(226Ra) P(226Ra) 17189 A(226Ra) 2729 Bq g1 t 300 s Zcopr 0.60 0.31 edet
5.8 10 g 131 11 Bq g1 — 0.02 0.01
5 10 0.008 0.004 — 0.03 0.03
Chemical recovery 0.03468 g m(242Pu) P(242Pu) 1728 A(242Pu) 1.06 Bq g1 t 208000 s edet 0.31 Zchem 0.74
5.8 105 g 42 0.00424 Bq g1 — 0.01 0.03
0.002 0.02 0.004 — 0.03 0.04
Activity concentration 0.6065 g ms P(239/40Pu) 843 t 208000 s 0.31 edet Zchem 0.74
5.8 105 g 29 — 0.01 0.03
1 104 0.03 — 0.03 0.04
A(239/40Pu)
2 Bq kg1
0.06
30 Bq kg1
511
5
4
Fig. 11. Relative distribution of uncertainties in measurement of activity concentration in IAEA 368.
239/40
Pu
Acknowledgement This work was financially supported by Ministry of Education, Science and Sport of Slovenia (Project group P1-0143, Cycling of nutrients and contaminants in the
Adsley, I., Andrew, D., Arnold, D., Bojanowski, R., Bourlat, Y., Byrne, A.R., Crespo, M-T., Desmond, J., De Felice, P., Fazio, A., Gasco´n, J.L., Grieve, R.S., Holmes, A.S., Jerome, S.M., Korun, M., Magnoni, M., Odell, K.J., Popplewell, D.S., Poupaki, I., Sutton, G., Toole, J., Wakerley, M.W., Wershofen, H., Woods, M.J., Youngman, M.J., 1998. The characterisation of an intertidal sediment from the Cumbrian coastline. Appl. Radiat. Isot. 49, 1295–1300. Alvarez, A., Navarro, N., 1996. Methods for actinides and Sr-90 determination in urine samples. Appl. Radiat. Isot. 47, 869–873. Bains, M.E.D., Warwick, P.E., 1993. The separation of actinides from lanthanides by anion exchange in methanol/hydrogen chloride medium, and its application to routine separation. Sci. Tot. Environ. 130–131, 437–445. Benedik, L., Pintar, H., Byrne, A.R., 1999. The leachability of some natural and man-made radionuclides from soil and sediments on acid attack. J. Radioanal. Nucl. Chem. 240, 859–865. Drolc, A., Rosˇ , M., 2002. Evaluation of measurement uncertainty in the determination of total phosphorous using standardized spectroscopic method ISO 6878. Acta Chim. Slov. 49, 409–423. Ellison, S.L., Rosslein, M., Williams, A., 2000. Quantifying uncertainty in analytical measurement. EURACHEM/CITAC. Hindman, F.D., 1983. Neodymium fluoride mounting for alpha spectrometric determination of uranium, plutonium and americium. Anal. Chem. 55, 2460–2461. Horwitz, E.P., Chiarizia, R., Dietz, M.L., Diamond, H., Nelson, D.M., 1993. Separation and preconcentration of actinides from acidic media by extraction chromatography. Anal. Chim. Acta 281, 361–372. Kaye, J.H., Strebin, R.S., Orr, R.D., 1995. Rapid, quantitative analysis of americium, curium and plutonium isotopes in Hanford samples using extraction chromatography and precipitation plating. J. Radioanal. Nucl. Chem. Lett. 194, 191–196. L’Annunziata, M.F., 1998. Handbook of Radioactivity Analysis. Academic press, London. Mellado, J., Llaurado´, M., Rauret, G., 2001. Determination of Pu, Am, U, Th and Sr in marine sediment by extraction chromatography. Anal. Chim. Acta 443, 81–90. Mellado, J., Llaurado´, M., Rauret, G., 2002. Determination of actinides and strontium in fish samples by extraction. Anal. Chim. Acta 458, 367–374. Moreno, J., Vajda, N., Danesi, P.R., Larosa, J.J., Zeiller, E., Sinojmeri, M., 1997. Combined procedure for the determination of 90Sr, 241Am and Pu radionuclides in soil samples. J. Radioanal. Nucl. Chem. 226, 279–284. Pilvio¨, R., Bickel, M., 1998. Separation of actinides from bone ash matrix with extraction chromatography. J. Alloys Comp. 271–273, 49–53. Rubio Montero, M.P., Martı´ n Sa´nchez, A., Crespo Va´zquez, M.T., Gasco´n Murillo, J.L., 2000. Analysis of plutonium in soil samples. Appl. Radiat. Isot. 53, 259–264. Tavcˇar, P., Jakopicˇ, R., Benedik, L., 2005. Sequential determination of 241 Am, 237Np, Pu radioisotopes and 90Sr in soil and sediment samples. Acta Chim. Slov. 52, 60–66. Toribio, M., Garcı´ a, J.F., Rauret, G., Pilvio¨, R., Bickel, M., 2001. Plutonium determination in mineral soils and sediments by a procedure involving microwave digestion and extraction chromatography. Anal. Chim. Acta 447, 179–189.
Applied Radiation and Isotopes 65 (2007) 504–511 www.elsevier.com/locate/apradiso
Sequential determination of Pu and Am radioisotopes in environmental samples; a comparison of two separation procedures Rozˇle Jakopicˇ, Polona Tavcˇar, Ljudmila Benedik Department of Environmental Sciences, Jozˇef Stefan Institute, Jamova 39, 1000 Ljubljana, Slovenia Received 29 May 2006; received in revised form 28 November 2006; accepted 13 December 2006
Abstract Two separation methods for the sequential determination of Am and Pu radionuclides are presented and the results obtained are compared. Analysis involves leaching the sample with concentrated nitric acid (HNO3), followed by radiochemical separation using extraction chromatographic resins (UTEVA, TRU) and anion exchange. Sources for alpha spectrometry were prepared by microprecipitation on neodymium fluoride (NdF3). The chemical recoveries were determined using 242Pu and 243Am tracers. The methods were tested on reference materials and on two sediments. All the results were in good agreement with the reference values. The evaluation of uncertainty is also included. r 2007 Published by Elsevier Ltd. Keywords: Actinides; Extraction chromatography; Anion exchange; Micro-precipitation technique; Alpha spectrometry; Uncertainty
1. Introduction Man-made radionuclides are released into the environment as a result of nuclear activities, atomic weapons testing and accidents at nuclear power plants. Of the many different man-made radionuclides, 241Am, 239Pu, 240Pu and 238 Pu are alpha emitting radionuclides that need special attention. These radionuclides are important long-term radioactive pollutants due to their long half-lives and long persistence in the environment. Due to increased energy production by nuclear reactors, reprocessing and waste disposal, the increased potential for contamination and public concern over the potential hazards, the accurate and rapid determination of artificial radionuclides in the environment is very important. Analysis of alpha emitting man-made radionuclides is complex and requires several steps. One of the important steps in such analysis is separation prior to measurement. The goal of separation is to eliminate all possible radioactive and non-radioactive interferences that could otherwise degrade the spectrum or overlap the peaks of interest. Corresponding author. Tel.: +386 1 5885 201; fax: +386 1 5885 346.
E-mail address: [email protected] (R Jakopicˇ). 0969-8043/$ - see front matter r 2007 Published by Elsevier Ltd. doi:10.1016/j.apradiso.2006.12.005
Several investigators have reported different procedures for separation and determination of man-made radionuclides. In the past, radiochemical separations were usually made by liquid–liquid extraction or ion exchange. The separations were tedious, time-consuming and generated a large amount of wastes. In the last 10 years, extraction chromatography has become a very attractive method for separation of the actinides in different environmental and biological samples (Horwitz et al., 1993; Pilvio¨ and Bickel, 1998; Mellado et al., 2001, 2002; Alvarez and Navarro, 1996; Kaye et al., 1995). Simple and fast separations with high chemical yield and good selectivity are achieved. The reason for the good selectivity lies in the organic extractant which is bound on an inert support and selectively retains certain radionuclides. In this paper, we compared separation techniques for plutonium and americium radionuclides using ion exchange (Dowex 1X8) and extraction chromatography using TRU and UTEVA resins (Eichroms Industries) in soil samples and sediments. For extraction chromatography we used a method developed by Toribio et al. (2001) which we slightly modified to fit our purposes. We used leaching in concentrated nitric acid (HNO3) for sample preparation instead of total digestion in a microwave oven and an
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
additional precipitation step on calcium oxalate (CaC2O4) was introduced prior to loading the sample solution onto the column. According to previous studies in our laboratory (Benedik et al., 1999), man-made radionuclides are usually adsorbed superficially and therefore leaching instead of total dissolution may be used. For ion exchange chromatography we used a method developed by Rubio Montero et al. (2000) with a modification in sample preparation by leaching in concentrated HNO3 and using 9 M hydrochloric acid (HCl) to remove thorium from the column. For the reduction of plutonium from Pu(IV) to Pu(III), hydroiodic acid was replaced with ammonium iodide (NH4I). Americium determination was also included in the separation scheme. All the radionuclides were measured by alpha spectrometry and the sources were prepared by micro-precipitation on neodymium fluoride. The results obtained for both procedures were compared in terms of activities and recoveries. 242Pu and 243Am tracers were used for the determination of the recovery. In the final presentation of an analytical result, it is very important to evaluate measurement uncertainty. In the past sometimes the quality of the results was defined by the standard deviation of repeated measurements, or was not even reported (Drolc and Rosˇ , 2002). In a radioactivity measurement the standard deviation is given by the square root of the number of counts (L’Annunziata, 1998), but this is only the statistical uncertainty of counting. Measurements uncertainty must also include other sources of uncertainty such as sample collection, tracer activity concentration, detector efficiency, chemical recovery, etc. The evaluation of uncertainty requires the analyst to look closely at all stages of the method and at all possible sources of uncertainty (Ellison et al., 2000). 2. Experimental 2.1. Samples Reference materials IAEA-135 (Radionuclides in Irish Sea Sediment), IAEA-368 (Pacific Ocean Sediment), IAEA-300 (Radionuclides in Baltic Sea Sediment), IAEA375 (Radionuclides in Soil), Soil-6 (Radionuclides in Soil), Ribble Sediment (sediment from the tidal zone of the river Ribble, Lancashire, UK) and intertidal sediment from the Cumbrian coast, UK. Ribble sediment is influenced both by tidal-borne radionuclides originating from Sellafield and the manufacture of fuel elements at Springfields upstream. 2.2. Reagents and materials All reagents used in the analysis were of analytical grade. The following reagents were used for leaching and for preparation of the solutions used in the radiochemical separation: 65% HNO3, 36% HCl, 40% HF, 32% H2O2, 25% ammonia (NH3), CH3OH, C2H5OH, NH4SCN,
505
NH4I, H2C2O4 2H2O, ascorbic acid (powder form), NaNO2, NH2OH HCl, TRU Spec and UTEVA Spec Resins (Eichroms Industries, 100–150m particle size resin), anion exchange resin (Dowex 1X8, 100–200 mesh). The following reagent solutions were prepared:
0.6 M ferrous sulphamate—Fe(NH2SO3)2 About 28.5 g of sulphamic acid (NH2SO3H) was added to 75 mL of deionised water, heated and 3.5 g of powdered iron was added and left to dissolve. Then the solution was filtered through a 0.2 mm filter and diluted to 100 ml with deionised water. The solution was prepared fresh weekly. 0.1 M ammonium hydrogen oxalate—NH4HC2O4 About 6.31 g of oxalic acid dihydrate (H2C2O4 2H2O) and 7.11 g of ammonium oxalate hydrate ((NH4)2 C2O4 H2O) were dissolved in 900 mL of deionised water and diluted to 1 L with deionised water. 10 mg m L1 neodymium fluoride substrate solution— NdF3 About 4.2 mg of neodymium (III) oxide (Nd2O3) was dissolved in 454 mL of 1 M HNO3 in a polyethene bottle and then 48 mL of 40% hydrofluoric acid was added.
2.3. Tracers 242
Pu standardised solution with an activity concentration of 1.06 Bq g1 (the expanded uncertainty 0.8%, k ¼ 2) and 243Am standardised solution with an activity concentration of 0.746 Bq g1 (the expanded uncertainty 1.89%, k ¼ 2) were purchased from the National Physical Laboratory, Teddington, Middlesex, UK. 2.4. Instrumentation For the alpha spectrometric measurements, a silicon surface barrier detector (PIPS) with an active area of 450 mm2 and counting efficiency of 31%71% connected to an EG&G ORTEC Maestro Gammavision MCA Emulator multichannel analyser system was used. The counting efficiency of the alpha detector was determined by measuring the activity of a 226Ra alpha source prepared by micro-precipitation (SRM 4967, activity concentration 2729 Bq g1, expanded uncertainty 1.18%, k ¼ 3). 2.5. Sample preparation The samples were dried to constant weight at 105 1C. After drying, the samples were ashed to decompose organic matter in an electric muffle furnace at 550 1C for 20 h. Approximately 1–2 g of ashed sample together with 242Pu and 243Am tracers were leached with 100 mL hot concentrated HNO3 with magnetic stirring overnight. During the leaching the beaker was covered with a watch glass to prevent significant evaporation. After cooling the solution to room temperature, the leachant and the residue were separated by filtration.
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
506
2.6. Separation procedure Methods A and B are schematically presented in Fig. 1. 2.6.1. Method A (extraction chromatography) The filtrate from sample preparation in 2.5 was heated until boiling and approximately 10 g of oxalic acid (H2C2O4 2H2O) was added. The pH of the solution was adjusted to 5.5–6.0 with NH3 solution. The actinides were co-precipitated on CaC2O4, while the iron remained mainly in solution as the oxalate complex (Moreno et al., 1997). For samples that have a low calcium content, additional calcium in the form of calcium chloride (CaCl2) was added to carry down actinides in the CaC2O4 precipitation step. The solution was cooled and the oxalate precipitate was centrifuged for 10 min at 2500 rpm and washed several times with deionised water. The oxalate precipitate was fumed twice with concentrated HNO3 and 32% hydrogen peroxide (H2O2) and evaporated to dryness. The residue was dissolved in 3 M HNO3 and then Fe(NH2SO3)2 was added to reduce plutonium to the trivalent oxidation state. The content of Fe(III) was checked by adding one drop of
1 M ammonium thyocyanate (NH4SCN) solution to 1 drop of sample solution. If the reaction was positive, then up to 200 mg of ascorbic acid (C6H8O6) was added to keep iron in the +2 valency state. Sequential separation of plutonium and americium from the sample solution was performed using UTEVA and TRU resins (100–150m, Eichroms Industries). About 3 g of UTEVA or TRU resin material were mixed with deionised water and used to fill a chromatographic column (1 cm internal diameter, 20 cm long, obtained from Bio-Rad Laboratories). The UTEVA column was conditioned with 60 mL of 3 M HNO3. The sample solution was loaded onto the column, followed by washing with 60 mL 3 M HNO3. Plutonium and americium radioisotopes were eluted from the column, while uranium and thorium radioisotopes were adsorbed and therefore removed. The effluent containing plutonium and americium was evaporated to dryness and redissolved in 2 M HNO3. The solution was loaded onto the TRU column, which was conditioned with 2 M HNO3. The TRU column was washed with 60 mL of 2 M HNO3–0.1 M sodium nitrite. Sodium nitrite (NaNO2) was used to oxidise Pu(III) to Pu(IV) to allow Pu and Am separation. After converting
sample (up to 2 g) + 242Pu and 243Am
Method A
co-precipitation on CaC2O4
leaching with conc. HNO3 and filtration
Method B
evaporation to dryness, dissolution in 1M HNO3,
oxalate destruction,
oxidation state adjustment,
dissolution in 3M HNO3,
load solution in 8M HNO3
oxidation state adjustment Dowex 1X8 column (1) UTEVA column (U, Th)
3M HNO3 (Pu, Am) evaporation to dryness,
8M HNO3 (Am, Fe)
9M HCl (Th, discard)
dissolution in 2M HNO3 9M HCl/0.1M NH4I (Pu) TRU column evaporation to dryness, 2M HNO3/0.1M NaNO2 (discard)
9M HCl, 4M HCl (Am)
0.1M NH4HC2O4 (Pu)
dissolution in 9M HCl
Dowex 1X8 column (2)
9M HCl Am → purification
Fig. 1. Analytical procedures for the sequential determination of plutonium and americium radioisotopes.
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
the column to the chloride system by means of 9 M HCl, americium was stripped with 60 mL of 4 M HCl and plutonium with 60 ml of 0.1 M NH4HC2O4. The approximate rates of loading and stripping solutions were 1 mL min1. 2.6.2. Method B (anion exchange) The filtrate from sample preparation in 2.5 was evaporated to dryness and the residue was dissolved in 1 M HNO3. About 1 g of hydroxylamine hydrochloride (NH2OH HCl) was added and the solution was heated to reduce plutonium to Pu(III). The completeness of the reaction was checked with a drop of 1 M ammonium thyocyanate solution. The solution was cooled and made 8 M in HNO3 by adding concentrated HNO3. Then 0.5 g sodium nitrite was added to ensure conversion of Pu(III) to Pu(IV). The solution was boiled to destroy excess hydroxylamine hydrochloride, cooled and transferred to the top of an anion exchange column of height 10 cm, diameter 1 cm (Dowex 1X8, 100–200 mesh), which was conditioned with 60 ml of 8 M HNO3. After washing Am and Fe from the column with 60 mL of 8 M HNO3 and Th with 100 ml of 9 M HCl, Pu radioisotopes were stripped with 60 mL freshly prepared 9 M HCl–0.1 M NH4I solution. The iodide reduced Pu(IV) to Pu(III) which was not retained by the column. The Am in the 8 M HNO3 effluent solution was evaporated to dryness and dissolved in 9 M HCl. The solution was loaded onto a second anion exchange resin column, which was prepared in 9 M HCl. Iron was retained, while Am passed through the column. The Am fraction was further purified from lanthanides. The solution was evaporated to dryness and dissolved in 1 M HNO3–methanol solution and loaded onto an anion exchange resin column. The column was washed with 60 mL of 1 M HNO3–methanol and 60 mL of 0.1 M HCl–0.5 M ammonium thyocyanate–80% methanol. Prior to stripping Am radioisotopes from the column with 100 mL of 1.5 M HCl–86% methanol, a wash with 30 mL 1 M HNO3–methanol was performed (Bains and Warwick, 1993). 2.7. Source preparation Alpha-counting sources for this work were prepared by neodymium fluoride micro-precipitation (Hindman, 1983). Solutions eluted from the columns were evaporated to dryness and fumed twice with 1 mL concentrated HNO3 and 1 mL 32% H2O2 to destroy NH4I and any organic compounds (resin) that might have been present in the eluate. Then 2 mL of 1 M HNO3 was added and the solution was transferred to a polypropylene centrifuge tube. About 50 mg Nd3+ solution and 0.5 mL 40% hydrofluoric acid were added, swirled and allowed to stand in an ice bath. After 30 min the neodymium fluoride suspension with plutonium radioisotopes was filtered through a 25 mm, 0.1 mm membrane filter (Supors Membrane Disc Filters, Gelman Laboratory) pre-washed
507
with 80% ethanol (C2H5OH) solution, deionised water and conditioned with 5 mL of neodymium substrate solution. The centrifuge tubes were rinsed with 5 ml of 0.58 M hydrofluoric acid solution, twice with deionised water and 5 mL 80% ethanol solution and the rinse solutions filtered. The filters were dried, mounted on an aluminium planchet and measured by alpha spectrometry. 3. Results and discussion The results obtained by sequential determination of plutonium and americium radionuclides are presented in Table 1 and in Figs. 2–8. The individual and the average chemical recoveries for each sample are presented in Table 1 and Figs. 7 and 8. Chemical recoveries are higher for method A than for method B, as shown in the bar chart in Figs. 7 and 8. Furthermore, the recoveries for plutonium radioisotopes are generally higher than for americium. For Pu the average recoveries in all samples used for method A are around 60% and for method B 42%. For americium the average recoveries are around 57% for method A and 43% for method B, respectively. For some samples (IAEA 300, IAEA 368, Ribble sediment), higher dispersion in recovery values was observed between replicates. The lowest chemical recoveries were obtained for americium radioisotopes in Cumbrian and Ribble sediment. Both these sediments have a higher content of iron than other samples, which was probably not removed completely in the preconcentration step and caused pre-elution of americium or interfered with adsorption of americium onto the TRU column. However, further work should be done to investigate this problem so as to allow higher recoveries. From Table 2 it can be seen that most of the results obtained by the two methods are in good agreement with the reference or literature values and previous work in our laboratory (Tavcˇar et al., 2005). For plutonium and americium radioisotopes in Cumbrian sediment lower results were observed; on the other hand, higher values were observed in Ribble sediment. In the case of 238Pu in Soil-6, the result was under the detection limit due to the low available mass of the sample. From the results in Table 2, it can also be concluded that leaching in this case for the determination of plutonium and americium radioisotopes was complete and therefore there was no need for total dissolution of the sample, suggesting negligible refractory artificial actinides in the samples. Leaching in concentrated HNO3 is also faster and simpler compared to total dissolution, thus saving time and chemicals. In Figs. 9 and 10, alpha spectra of Am radioisotopes isolated from IAEA-135 reference sample and Pu radioisotopes isolated from Ribble sediment are presented. In the case of the plutonium spectrum from IAEA-135 in Fig. 9, three peaks are observed, one belonging to tracer 242 Pu which was added at the beginning of the procedure to monitor the losses of plutonium during separation (leaching, radiochemical separation, source preparation) and
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
508
Table 1 Individual and average chemical recoveries for Pu and Am radioisotopes Sample: Pu
Method A
Method B
Individual Pu recovery (%)
Average Pu recovery (%)
Individual Pu recovery (%)
Soil-6 IAEA 135 IAEA 300 IAEA 368 IAEA 375 Cumbrian s. Ribble s.
63 46 78 74 60 62 78
7178 4473 62713 66711 5277 5778 64719
43 36 40 41 53 32 50
Sample: Am
Method A
Soil-6 IAEA 135 IAEA 300 IAEA 368 IAEA 375 Cumbrian s. Ribble s.
78 42 49 58 49 51 51
73 66
56
46
44 43 47 52 53 31 48
40
38
50 41 51
51
40
4173 4075 4675 4576 5271 3271 4675
Method B
Individual Am recovery (%)
Average Am recovery (%)
Individual Am recovery (%)
62 61 68 81 50 50 37
5874 6879 6077 71715 4973 5273 4176
29 45 50 47 46 28 30
57 74 63 60 51 54 45
Average Pu recovery (%)
55 54
53
46
40 55 74 58 45 24 27
Average Am recovery (%)
55
42
41 41 42
43
36
42711 5077 55717 4979 4472 2673 3175
The uncertainties correspond to the standard deviation of the average at the 71 s level.
239/240Pu
241Am
reference value method A method B previous work
Bq/kg
1.5 1
reference value method A merhod B previous work
4 Bq/kg
2
5
3 2 1
0.5
0
0
IAEA-300
IAEA-300
IAEA-368
IAEA-375
Soil-6
sample Fig. 2. A comparison of activity concentration in Bq kg IAEA-300, IAEA-368, IAEA-375 and Soil-6.
1
of
241
Am in
IAEA-375 sample
Soil-6
Fig. 4. A comparison of activity concentration in Bq kg1 of IAEA-300, IAEA-375 and Soil-6.
239/240
Pu in
239/240Pu 241Am
1200 800
reference value method A method B previous work
1000 800 Bq/kg
1600
Bq/kg
1200 reference value method A method B previous work
600 400
400
200
0
0 IAEA-135
Cumbrian Sediment
Ribble Sediment
sample
Fig. 3. A comparison of activity concentration in Bq kg1 of Cumbrian, Ribble sediment and IAEA-135.
241
Am in
Cumbrian Sediment
Ribble IAEA-368 Sediment sample
IAEA-135
Fig. 5. A comparison of activity concentration in Bq kg1 of IAEA-135, IAEA-368, Ribble and Cumbrian sediment.
239/240
Pu in
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
of radioactive wastes resulting from fuel fabrication and reprocessing activities.
238Pu
250
reference value method A method B previous work
Bq/kg
200 150 100 50
3.1. Evaluation of measurement uncertainty The procedure for evaluation of the uncertainty of the result of 239/40Pu activity concentration is presented below. The activity concentration of 239/40Pu is calculated from Eq. (1)
0 Cumbrian Sediment
IAEA-135 Ribble Sediment sample
IAEA-368
Fig. 6. A comparison of activity concentration in Bq kg1 of Cumbrian, Ribble sediment, IAEA-135 and IAEA-368.
90 method A
238
Pu in
method B
80 Pu recovery (%)
70 60 50 40
20 10 0 Soil-6
Cumbrian Ribble Sediment Sediment
Sample
Fig. 7. Pu recoveries in various reference materials obtained by two methods.
90
method A
method B
80 Am recovery (%)
A239=40 Pu ¼
P239=40 Pu ; Zchem det ms t
(1)
where A239=40 Pu is the activity concentration of 239/40Pu in Bq kg1, P239=40 Pu is the peak area of 239/40Pu, Zchem is the chemical recovery, edet is the detector efficiency, t is the time of measurement in s and ms is the sample mass in kg. The uncertainty associated with the mass u(ms) of the sample was estimated using data from the calibration certificate and the analytical balance manufacturer’s recommendations on uncertainty estimation as 0.1 mg. We assumed a rectangular distribution and converted to a normal distribution by dividing by 31/2 Eq. (2) 0:1 mg uðms Þ ¼ pffiffiffi ¼ 0:058 mg: 3
30
IAEA-135 IAEA-300 IAEA-368 IAEA-375
509
(2)
The uncertainty in chemical recovery u(Zchem) for 239/40Pu consists of the uncertainty of the tracer activity contribution uðA242 Pu Þ, the peak area of plutonium tracer uðP242 Pu Þ, the mass of plutonium tracer added to the sample uðm242 Pu Þ and the detector efficiency u(edet) and it is calculated from Eq. (3) uðZchem Þ ¼ Zchem sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi uðm242 Pu Þ 2 uðZdet Þ 2 uðA242 Pu Þ 2 uðP242 Pu Þ 2 . þ þ þ m242 Pu Zdet A242 Pu P242 Pu ð3Þ
70 60
The uncertainty of the tracer activity uðA242 Pu Þ was taken from the certificate. The uncertainty of the peak area uðP242 Pu Þ was calculated as the square root of the peak area pffiffiffiffiffiffiffiffiffiffiffiffi (4) uðP242 Pu Þ ¼ P242 Pu .
50 40 30 20 10 0 IAEA-135
IAEA-300
IAEA-368
IAEA-375
Soil-6
Cumbrian Sediment
Ribble Sediment
Sample
Fig. 8. Am recoveries in various reference materials obtained by two methods.
238
239/240
other two to Pu and Pu present in the sample. The doublet peak of 239/240Pu cannot be resolved by alpha spectrometry due to closeness of the alpha energies of 239 Pu and 240Pu. In the Am spectrum isolated from Ribble sediment in Fig. 9, two peaks are observed, one of 243 Am tracer and other of 241Am, which is present in the sample. The highest activities were observed in Ribble and Cumbrian sediments. This was expected since both sediments originated from areas contaminated by disposal
The uncertainty in detector efficiency consists of the uncertainty of 226Ra activity uðA226 Ra Þ, uncertainty of peak area uðP226 Ra Þ, mass uðm226 Ra Þof 226Ra added and the recovery of micro-precipitation u(Zmicro) and is calculated from Eq. (5) det ¼
P226 Ra . A226 Ra Zmicro m226 Ra t
(5)
In the determination of 239/40Pu, the combined uncertainty is calculated from Eq. (6) uðA239=40 Pu Þ ¼ A239=40 Pu sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi uðP239=40 Pu Þ 2 uðZchem Þ 2 uðdet Þ 2 uðms Þ 2 . þ þ þ P239=40 Pu Zchem det ms ð6Þ
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511
510
Table 2 Activity concentrations in Bq kg1 dry weight for selected Pu and Am radioisotopes Sample
Activity concentration (Bq kg1)
IAEA 135
238
Isotope Pu
239/240
Pu Am
241 238
Soil-6
Pu
239/240
Pu 241 Am IAEA 300
238
Pu
239/240
Pu 241 Am
IAEA 368
238
Pu
239/240
Pu Am
241
IAEA 375
238
Pu
239/240
Pu 241 Am Cumbrian sediment
238
Pu
239/240
Pu Am
241
Ribble sediment
238
Pu
239/240
Pu 241 Am
This work Method A
This work Method B
4274(2) 209713 321720
3974(2) 215715 291715
o 0.2(3) 1.0070.11 0.4070.09
o0.2(4) 1.0770.10 0.3670.07
Previous work (Tavcˇar et al., 2005)
Reference or literature value 43 (42–45)a 213 (205–226)a 318 (310–325)
4674 217712 313714 o 0.2 1.0370.16 0.3970.08
o0.036 1.0470.07 0.39–0.45b 0.15 3.55 (3.44–3.65)a 1.38 (1.2–1.5)a
0.1270.04(4) 3.6270.30 1.3570.16
0.1370.03(3) 3.4770.36 1.2170.12
o 0.2 3.4970.23 1.2170.18
9.370.7(2) 3072 1.370.3
8.770.8(3) 3272 1.270.2
8.271.0 3175 1.470.2
8.570.9a 3172a 1.3 (1.21.5)
0.10470.047(3) 0.4070.04 0.1270.02
0.09070.020(4) 0.3770.05 0.1070.02
0.08070.030 0.3470.06 0.1570.03
0.07170.014 0.3070.04 0.1370.07
196710(2) 917751 1474791
194714(2) 878753 1430782
193719 930774 1508799
21174c 995719c 1590713c
3372(2) 186712 300732
3272(3) 195713 311721
3273 187710 288723
3375 16178 283–351d
Pu is decay corrected to the reference date. Uncertainties are given at 71 s. Number of determinations are given in parentheses. Certified value. b Moreno et al. (1997). c Adsley et al. (1998). d Benedik et al. (1999).
238
a
1360
220 241Am 243Am
tracer
110
Pu
680 242
Pu tracer
238
340
55
3880.00
239/240
1020 Counts
Counts
165
4497.00
5028.00 Energy (keV)
5473.00
5831.00
2951.00
4309.00
Pu
5667.00
7024.00
8382.00
Energy (keV)
Fig. 9. Alpha spectrum of americium radioisotopes isolated from IAEA135 reference material.
Fig. 10. Alpha spectrum of plutonim radioisotopes isolated from Ribble sediment.
Standard measurement uncertainty and relative standard uncertainty in the determination of 239/40Pu in IAEA 368 are summarised in Table 3 and schematically presented in Fig. 11.
two different radiochemical procedures and measured by alpha spectrometry. Thin sources for alpha measurements were prepared by the micro-precipitation technique. The results demonstrated the usefulness of a simultaneous procedure and therefore a shorter time for the analysis. The major source of uncertainty in the result of the measurement was identified as the contribution from detector efficiency uncertainty, chemical recovery uncertainty and uncertainty in peak area.
4. Conclusion Plutonium and americium radioisotopes in various samples were determined in the same sample aliquot using
ARTICLE IN PRESS R Jakopicˇ et al. / Applied Radiation and Isotopes 65 (2007) 504–511 Table 3 Standard measurement uncertainty and relative standard uncertainty in the determination of 239/40Pu activity concentration in IAEA 368 Value x
Standard Relative standard measurement measurement uncertainty u(x) uncertainty u(x)/x
environment, mass balances and modelling of the environmental processes and risk analysis). The authors wish to express their thanks to Dr. P.J. Day, University of Manchester, who kindly supplied the Ribble Sediment sample. References
Detector efficiency 0.1149 g m(226Ra) P(226Ra) 17189 A(226Ra) 2729 Bq g1 t 300 s Zcopr 0.60 0.31 edet
5.8 10 g 131 11 Bq g1 — 0.02 0.01
5 10 0.008 0.004 — 0.03 0.03
Chemical recovery 0.03468 g m(242Pu) P(242Pu) 1728 A(242Pu) 1.06 Bq g1 t 208000 s edet 0.31 Zchem 0.74
5.8 105 g 42 0.00424 Bq g1 — 0.01 0.03
0.002 0.02 0.004 — 0.03 0.04
Activity concentration 0.6065 g ms P(239/40Pu) 843 t 208000 s 0.31 edet Zchem 0.74
5.8 105 g 29 — 0.01 0.03
1 104 0.03 — 0.03 0.04
A(239/40Pu)
2 Bq kg1
0.06
30 Bq kg1
511
5
4
Fig. 11. Relative distribution of uncertainties in measurement of activity concentration in IAEA 368.
239/40
Pu
Acknowledgement This work was financially supported by Ministry of Education, Science and Sport of Slovenia (Project group P1-0143, Cycling of nutrients and contaminants in the
Adsley, I., Andrew, D., Arnold, D., Bojanowski, R., Bourlat, Y., Byrne, A.R., Crespo, M-T., Desmond, J., De Felice, P., Fazio, A., Gasco´n, J.L., Grieve, R.S., Holmes, A.S., Jerome, S.M., Korun, M., Magnoni, M., Odell, K.J., Popplewell, D.S., Poupaki, I., Sutton, G., Toole, J., Wakerley, M.W., Wershofen, H., Woods, M.J., Youngman, M.J., 1998. The characterisation of an intertidal sediment from the Cumbrian coastline. Appl. Radiat. Isot. 49, 1295–1300. Alvarez, A., Navarro, N., 1996. Methods for actinides and Sr-90 determination in urine samples. Appl. Radiat. Isot. 47, 869–873. Bains, M.E.D., Warwick, P.E., 1993. The separation of actinides from lanthanides by anion exchange in methanol/hydrogen chloride medium, and its application to routine separation. Sci. Tot. Environ. 130–131, 437–445. Benedik, L., Pintar, H., Byrne, A.R., 1999. The leachability of some natural and man-made radionuclides from soil and sediments on acid attack. J. Radioanal. Nucl. Chem. 240, 859–865. Drolc, A., Rosˇ , M., 2002. Evaluation of measurement uncertainty in the determination of total phosphorous using standardized spectroscopic method ISO 6878. Acta Chim. Slov. 49, 409–423. Ellison, S.L., Rosslein, M., Williams, A., 2000. Quantifying uncertainty in analytical measurement. EURACHEM/CITAC. Hindman, F.D., 1983. Neodymium fluoride mounting for alpha spectrometric determination of uranium, plutonium and americium. Anal. Chem. 55, 2460–2461. Horwitz, E.P., Chiarizia, R., Dietz, M.L., Diamond, H., Nelson, D.M., 1993. Separation and preconcentration of actinides from acidic media by extraction chromatography. Anal. Chim. Acta 281, 361–372. Kaye, J.H., Strebin, R.S., Orr, R.D., 1995. Rapid, quantitative analysis of americium, curium and plutonium isotopes in Hanford samples using extraction chromatography and precipitation plating. J. Radioanal. Nucl. Chem. Lett. 194, 191–196. L’Annunziata, M.F., 1998. Handbook of Radioactivity Analysis. Academic press, London. Mellado, J., Llaurado´, M., Rauret, G., 2001. Determination of Pu, Am, U, Th and Sr in marine sediment by extraction chromatography. Anal. Chim. Acta 443, 81–90. Mellado, J., Llaurado´, M., Rauret, G., 2002. Determination of actinides and strontium in fish samples by extraction. Anal. Chim. Acta 458, 367–374. Moreno, J., Vajda, N., Danesi, P.R., Larosa, J.J., Zeiller, E., Sinojmeri, M., 1997. Combined procedure for the determination of 90Sr, 241Am and Pu radionuclides in soil samples. J. Radioanal. Nucl. Chem. 226, 279–284. Pilvio¨, R., Bickel, M., 1998. Separation of actinides from bone ash matrix with extraction chromatography. J. Alloys Comp. 271–273, 49–53. Rubio Montero, M.P., Martı´ n Sa´nchez, A., Crespo Va´zquez, M.T., Gasco´n Murillo, J.L., 2000. Analysis of plutonium in soil samples. Appl. Radiat. Isot. 53, 259–264. Tavcˇar, P., Jakopicˇ, R., Benedik, L., 2005. Sequential determination of 241 Am, 237Np, Pu radioisotopes and 90Sr in soil and sediment samples. Acta Chim. Slov. 52, 60–66. Toribio, M., Garcı´ a, J.F., Rauret, G., Pilvio¨, R., Bickel, M., 2001. Plutonium determination in mineral soils and sediments by a procedure involving microwave digestion and extraction chromatography. Anal. Chim. Acta 447, 179–189.