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An estimation of radiation doses to benthic invertebrates from sediments collected near a Canadian uranium mine
Environment International 27 (2001) 341 – 353 www.elsevier.com/locate/envint
An estimation of radiation doses to benthic invertebrates from sediments collected near a Canadian uranium mine Patricia Thomas*, Karsten Liber1 Toxicology Centre, University of Saskatchewan, 44 Campus Drive, Saskatoon, Saskatchewan, Canada S7N 5B3 Received 6 February 2001; accepted 1 June 2001
Abstract A new method is described for calculating radiation doses to benthic invertebrates from radionuclide concentrations in freshwater sediment. Both internal and external radiation doses were estimated for all 14 principal radionuclides of the uranium-238 decay series. Sediments were collected from three sites downstream of a uranium mining operation in northern Saskatchewan, Canada. Sediments from two sites, located approximately 1.6 and 4.4 km downstream from mining operations, yielded absorbed doses to both larval midges, Chironomus tentans, and adult amphipods, Hyalella azteca, of 59 – 60 and 19 mGy/year, respectively, compared to 3.2 mGy/year for a nearby control site. External beta radiation from protactinium-234 (234Pa) and alpha radiation from uranium (U) contributed most of the dose at the impacted sites, whereas polonium-210 (210Po) was most important at the control site. If a weighting factor of 20 was employed for the greater biological effect of alpha vs. beta and gamma radiation, then total equivalent doses rose to 540 – 560 mGy/year at the site closest to uranium operations. Such equivalent doses are above the 360-mGy/year no-observed-effect level for reproductive effects in vertebrates from gamma radiation exposure. Data are not available to determine the effect of such doses on benthic organisms, but they are high enough to warrant concern. Detrimental effects have been observed in H. azteca at similar uranium concentration in laboratory toxicity tests, but it remains unclear whether the radiotoxicity or the chemotoxicity of uranium is responsible for these effects. D 2001 Elsevier Science Ltd. All rights reserved. Keywords: Uranium mining; Sediment contamination; Radiation dose; Benthic invertebrates
1. Introduction Animals and humans receive natural background radiation doses from cosmic rays, gamma rays arising from rocks and soil, inhalation of radon gas, and ingestion of radionuclides with food, water, and soil. In humans, annual background doses from all of these sources range from 1 to 10 milliSievert (mSv) (Committee on the Biological Effects of Ionizing Radiation (BEIR), 1990; Beak Consultants, 1995). Animals often receive higher doses than humans because they ingest more soil or sediment with food items, which raises their body burden of radionuclides. In addition, they live outdoors, which increases their exposure to cosmic rays and terrestrial gamma radiation. * Corresponding author. Tel.: +1-306-244-4807; fax: +1-306-931-1664. E-mail addresses: [email protected] (P. Thomas), [email protected] (K. Liber). 1 Also corresponding author. Tel.: + 1-306-966-7444; fax: + 1-306931-1664.
Based on their body burdens of five uranium series radionuclides (U, 226Ra, 210Pb, 210Bi, and 210Po), annual background doses of 18 –38 mGy have been calculated to small mammals, birds, and caribou from unpolluted sites in northern Canada (Thomas, 1999, 2000). These dose calculations employed the same weighting factor of 20, as is used in human dose calculations, to account for the greater relative biological effect (RBE) of alpha vs. beta and gamma radiation. Some small mammals living close to uranium mine tailings received doses of up to 230 mGy/year, mostly from the ingestion of 226Ra (Thomas, 2000). These doses would have been even higher if external radiation doses, as well as all 14 principal radionuclides in the uranium decay series, had been considered. One group of animals that may receive elevated radiation doses are the benthic macroinvertebrates, which burrow in sediment. Radionuclides often accumulate in sediment, such as in depositional zones downstream of uranium mining operations. Benthic macroinvertebrates are important components of aquatic food chains and, thus, an
0160-4120/01/$ – see front matter D 2001 Elsevier Science Ltd. All rights reserved. PII: S 0 1 6 0 - 4 1 2 0 ( 0 1 ) 0 0 0 8 5 - X
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effort to estimate radiation doses to such organisms should be made. However, radiation doses to many benthic organisms are difficult to calculate because these organisms are too small to allow for direct radiochemical analyses. It is also difficult to distinguish between radionuclides absorbed into body tissues, those adsorbed to the external body surfaces and those that are within the animal gastrointestinal tract (GIT). This publication describes the calculations of internal and external radiation doses to two benthic macroinvertebrates — the larval midge, Chironomus tentans, and the amphipod, Hyalella azteca. The dose calculations are based solely on the radionuclide concentrations in the sediment which surround these organisms and which they ingest. The work presented here was part of a larger study designed to evaluate the potential impact of effluent from the Rabbit Lake uranium mine in northern Saskatchewan, Canada, on adjacent aquatic communities. Differences in natural benthic communities downstream from the mining operation, correlation of such differences to uranium and other contaminant exposure, and assessment of sediment toxicity will be described elsewhere (Liber et al., in preparation).
2. Methods 2.1. Sediment sampling and analysis Sediment samples for radiochemical analysis were collected from three different locations (Horseshoe Pond, Hidden Bay, and Tributary Creek) near from the Rabbit
Lake uranium mine in northern Saskatchewan, Canada (581103800N, 1034205700W). Horseshoe Pond is located approximately 1.6 km downstream from the mine effluent treatment plant. The pond empties into Horseshoe Creek, which flows into Hidden Bay on Wollaston Lake, approximately 4.4 km further downstream. Tributary Creek, the control site, also empties into Horseshoe Creek, but originates from an area north of Horseshoe Creek and is not impacted by any effluent or mining activity. Sediments were collected from all three sites with a small Ekman grab sampler (15 15 cm). Care was taken to ensure that all sediment samples were collected from the top 5 – 10 cm sediment layer. Three replicate samples were collected from each site and placed in separate 1-l Nalgene bottles for transportation in ice-packed coolers to Saskatoon, SK. Five radionuclides in dry sediments were analyzed by the Saskatchewan Research Council: uranium (U), thorium-230 (230Th), radium-226 (226Ra), lead-210 (210Pb), and polonium-210 (210Po). Uranium was analyzed by delayed neutron counting (detection limit = 0.1 mg/g of sample). 230Th, 226 Ra, and 210Po were chemically extracted and analyzed by alpha spectroscopy (detection limits 0.02, 0.005, and 0.005 Bq/g, respectively). 210Pb was analyzed by precipitation of its decay product bismuth-210 (210Bi) and counted by beta spectroscopy (detection limit = 0.02 Bq/g). Detailed descriptions of these procedures can be found in Canadian Mining and Energy Technologies (CANMET, 1978). Not all radionuclides in the 238U decay series were measured in sediments. However, the entire decay series (Table 1) was considered in the radiation dose calculations
Table 1 Uranium-238 decay series Average energy (MeV) per radiation type Radionuclide
Half-life b
Natural uranium Uranium-238b Thorium-234 Protactinium-234 Uranium-234b Thorium-230 Radium-226 Radon-222(gas) Polonium-218 Lead-214 Bismuth-214 Polonium-214 Lead-210 Bismuth-210 Polonium-210 Lead-206 (stable)
4.5 billion years 24.1 days 1.17 min 244,500 years 80,000 years 1,600 years 3.85 days 3.05 min 26.8 min 19.7 min 0.00015 s 22.2 years 4.97 days 138 days
Alpha 4.55 see natural uranium above none none see natural uranium above 4.74 4.86 ignored 6.11 none none 7.83 none none 5.4
Betaa
Gammaa 2
2.61 10 3
5.92 10 2 8.20 10 1
9.34 10 3 1.13 10 2
1.46 10 2 3.59 10 3 1.09 10 5 1.42 10 5 2.91 10 1 6.48 10 1 8.19 10 7 3.80 10 2 3.89 10 1 8.18 10 8
1.55 10 3 6.74 10 3 3.98 10 4 9.12 10 6 2.48 10 1 1.46 10 0 8.33 10 5 4.81 10 3 none 8.50 10 6
1.19 10
Average energies are calculated from Yield Energy values in ICRP (1983) data tables. a All electron energies from beta particles, as well as conversion and Auger electrons, are included under ‘‘Beta’’ in the table. All photon energies from gamma, as well as X-rays, are included under ‘‘Gamma’’ in the table. b By mass, natural uranium is composed of 99.3% 238U and 0.7% fissile 235U (half-life = 710 million years). 238U and 234U are in equilibrium and, thus, contribute equally to the activity of natural U, although the mass of 234U present is much smaller than that of 238U. Including all three radionuclides, 1 mg/kg of natural U is equivalent to 25.2 Bq/kg.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
by making the following assumptions regarding the chemistry and mobility of each radionuclide in the environment: (1) 234Th, 234Pa, and 234U were in equilibrium with and, thus, had similar activities to 238U; (2) radon gas and its short-lived progeny were in partial equilibrium (30%) with the parent, 226Ra, as Blaylock et al. (1993) has assumed in other studies; and (3) 210Bi was in equilibrium with its parent, 210Pb. 210Po was measured separately and corrected for decay of unsupported 210Po, as well as ingrowth from 210Pb between the date of sampling and the date of analysis. 2.2. Dose calculations Total radiation doses from the 14 principal uranium series radionuclides were calculated to two benthic organisms, based on the mean radionuclide concentrations in sediments from three sites: Horseshoe Pond, Hidden Bay, and Tributary Creek, a control site (Table 2). Eqs. 1 to 3 (below) consider three sources of radionuclides which can irradiate the organism: (1) alpha irradiation from sediment within the GIT; (2) the fraction of radionuclides in the GIT, which are absorbed into body tissues; and (3) external radiation from radionuclides in the sediment outside the organism. From each of these three sources, the energy (in joules) absorbed per kilogram of body tissue (excluding the GIT contents) was determined for the alpha, beta, and gamma radiations from each of the 14 238U decay series radionuclides listed in Table 1.
Table 2 Radionuclide concentrations (Bq/kg dry weight) in sediments collected from three sites near the Rabbit Lake uranium mine in northern Saskatchewan, September 1997 (N = 3) Horseshoe Pond
Hidden Bay
Tributary Creek (control site)
Radionuclides measured Unata Mean S.D. 230 Th Mean S.D. 226 Ra Mean S.D. 210 Pb Mean S.D. 210 Po Mean S.D.
50,484 7,288 140 30 600 157 2,057 2,054 544 342
13,616 979 37 21 53 12 507 51 141 35
156 25 30 17 57 31 153 68 240 156
Sediment characteristics Particle size classes (%)b Sand ( > 53 mm) Silt (2 – 52 mm) Clay ( < 2 mm) Water content (%)
27.5 ± 3.7 36.0 ± 2.2 36.5 ± 1.6 79.5 ± 2.6
35.9 ± 6.6 56.5 ± 5.7 7.6 ± 1.0 71.1 ± 6.6
68.3 ± 28.9 28.8 ± 26.5 2.9 ± 2.3 35.9 ± 11.2
a By dividing natural U activity by 25.2, the concentration in milligram per kilogram (ppm) can be found. b Mean particle size ( ± 1 S.E.) for separate samples collected with a corer from a depth of 0 – 6 cm.
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2.2.1. Alpha dose from radionuclides within the GIT Internal radiation doses from alpha-emitting radionuclides in sediment within the GIT were determined by Eq. (1): Dose ðmGy=yearÞ ¼ ð1:6 1010 mJ=MeVðBq=kg d:w: 0:5 kg of source w:w: s=yearÞ ðYE AFÞÞ=ðkg of target w:w:Þ ð1Þ where 1. a conversion factor of 1.6 10 10 mJ/MeV converts absorbed energy in million electron volts (MeV) to millijoules (mJ); 2. Bq/kg d.w. is the radionuclide concentration (dry weight) in sediment within the GIT; where 1 Bq = 1 nuclear disintegration/s. These concentrations were multiplied by 0.5 to account for the 50% moisture content assumed for sediment within the GIT. A factor of 50% was chosen to reflect fine sediment in the GIT being slightly more compacted than in the external sediment; 3. kg of source w.w. refers to the mass of the GIT contents (wet weight); 4. there are 3.156 107 s/year; 5. YE are the average ‘‘yield times energy’’ values in million electron volts from Table 1, determined for the alpha radiations from each radionuclide from the International Commission of Radiological Protection (ICRP, 1983); 6. AF is the absorbed fraction, or the fraction of radiation energy emitted from a radionuclide within the GIT that is absorbed by body tissues; and 7. kg of target w.w. is the body mass of the organism minus the GIT contents (wet weight). Factors 2 – 4, listed above from Eq. (1), determine the number of disintegrations from a radionuclide occurring within the GIT per year. Factors 5– 7 determine the energy from each nuclear disintegration in the GIT source absorbed per kilogram of body target tissue (in million electron volts per disintegration per kilogram). When million electron volts are converted to millijoules by Factor 1, the result is the absorbed dose to the organism in milligray per year, where 1 mGy = 1 mJ/kg. Eq. (1) was calculated separately for the alpha radiations from each radionuclide and then summed for all radionuclides. Body length, width, and mass were measured in several organisms of both macroinvertebrate species (the midge, C. tentans, and the amphipod, H. azteca) reared in beakers containing uncontaminated sediment from Wollaston Lake in northern Saskatchewan (Table 3). The mass of the GIT was calculated as the difference between full organisms and those purged of their GIT contents. The mass of the whole body minus the GIT contents represented body tissues and
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Table 3 Mean morphometric measurements of C. tentans and H. azteca C. tentans (fourth instar) Body mass measurements (n = 5) Body mass with full GIT a (mg wet weight) Target tissue for radiation: body mass with purged GIT (mg wet weight) Source tissue: mass of GIT contents (mg wet weight) GIT as percent of body mass with purged GIT Body size measurements (n = 6) Body length (mm) Lateral body width (mm) Diameter of GIT (mm) Percent of GIT within 50 mm of the GIT wallb
H. azteca (adult)
11.68 ± 5.61
4.41 ± 0.67
10.90 ± 5.22
4.18 ± 0.66
0.79 ± 0.44
0.22 ± 0.12
7.2 ± 5.3
5.3 ± 3.0
12,364 ± 1450 796 ± 73 300 ± 122 56
6,494 ± 312 840 ± 32 152 ± 28 88
a
GIT = gastrointestinal tract. The mass of a cylinder of GIT contents (200 mm in diameter for C. tentans and 52 mm in diameter for H. azteca), from which no alpha particle could reach the GIT wall, was subtracted from the total GIT mass. The resulting mass as a percentage of the total GIT mass is the mass within 50 mm from the wall. This percentage was used in internal AF calculations because only alpha particles within this portion of the GIT can irradiate body tissues to varying degrees, depending on their distance from the wall. b
was used as the target mass of body tissue exposed to radionuclides within the GIT. The AF values for alpha radiations were 0.14 for C. tentans and 0.22 for H. azteca. The short range of alpha particles within the GIT contents meant that only a portion of the GIT contents are within a 50-mm range from where an ionizing alpha particle can penetrate into body tissues. The proportion of GIT mass within 50 mm of the GIT wall was 56% in C. tentans and 88% in H. azteca, determined from calculated GIT diameters in Table 3. Within this 50-mm layer, the fraction of energy absorbed by body tissues ranges from 0.5 for particles emitted next to the GIT wall2 to 0 for alpha particle emitted from a distance of 50 mm from the wall (Fig. 1). The average fraction absorbed for all distances (0– 50 mm from the GIT wall) was estimated at 0.25.3 This value was then multiplied by the proportion of GIT contents < 50 mm from the GIT wall, yielding overall AF values of 0.14 (0.25 0.56) for C. tentans and 0.22 (0.25 0.88) for H. azteca. The doses from beta and gamma radiations within the GIT contents were ignored because the AF values from such radiations would be small. Many of the weak beta and 2
Only half of the alpha particles emitted next to the GIT wall will penetrate body tissue. The other half will decay within the GIT contents and, thus, do not deliver any dose to living biological tissue. 3 An exact absorbed fraction or AF value for all alpha radiations arising from within the 50-mm layer requires the integration of the energy deposited along all pathlengths for particles which can penetrate the GIT wall and the integration over all distances 0 – 50 mm from the GIT wall.
gamma radiations from the 238U series radionuclides would be attenuated in the GIT contents. Of the beta and gamma radiations reaching body tissues, little of the energy would be absorbed, given a total body width of < 900 mm in the study organisms. For alpha radiations from each radionuclide, a radiation weighting factor (wR) of 20 was applied to convert absorbed dose to ‘‘equivalent dose’’. The wR factor accounts for the greater RBE of alpha vs. beta and gamma radiations in causing biological damage. A factor of 20 is used for the greater RBE in causing cancer in humans (ICRP, 1991). It may or may not apply to other species and endpoints, but there is currently no better alternative. 2.2.2. Doses from radionuclides absorbed into body tissues from the GIT The second source of radiation dose to the benthic organisms is the radionuclides, which have been absorbed from the GIT and have entered body tissues. Ideally, these radionuclides would be directly measured in body tissues, but the small size of the organisms makes such measurements difficult. At equilibrium in a chronic intake situation, one can assume that the radionuclide concentration in the GIT will equilibrate with body tissues to varying degrees, depending on solubility of the radionuclide. As a first approximation of equilibrium for radionuclides between the GIT and body tissues, human f1 values were used in Eq. (2). Tissue dose ðmGy=yearÞ ¼ 1:6 1010 mJ=MeV Bq=kg d:w: 0:5 f1 s=year YE AF
ð2Þ
where f1 = 0.02 for U; 0.0005 for 234Th, 230Pa, and 230Th; 0.2 for 226Ra; 0.2 for 210Pb; 0.05 for 210Bi; and 0.5 for 210Po (ICRP, 1996). The f1 value estimates the fractional amount of each radionuclide absorbed from the GIT into the blood in adult humans (ICRP, 1996). The f1 values are no less accurate in this situation than animal/sediment concentration ratios (CRs), which range from 0.22 for field-collected organisms (Swanson, 1985) to < 0.136 (Hart et al., 1986) to 0.007 – 0.016 (Beak International, 1998) for uranium. Not only are CRs notoriously site-, species-, and season-specific, but they are unavailable for many of the 238U series radionuclides and require empirical measurements of radionuclide concentrations in very small organisms. Such measurements are likely to include the radionuclides in the GIT and adsorbed to the external surface of the organism, leading to an overestimate of the CR value. The f1 values may also overestimate absorption by invertebrates because the invertebrate GIT is more alkaline than the human GIT. Many radionuclides are more soluble in acid and, thus, may be absorbed more efficiently in humans than in invertebrates.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
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Fig. 1. Diagrammatic representation of the GIT within the body of an organism for AF determination. The length of each arrow represents the range of an alpha particle (i.e., 50 mm) originating within the GIT. (A) Alpha particles can be emitted at any angle from any point within the GIT. (B) AF = 0 because all of the alpha particle energy is deposited within the GIT before it reaches body tissues. (C) AF 0.25 for an alpha particle 25 mm from the GIT over the angle of 180. This approximation is used because most of the alpha energy is deposited near the end of its range. (D) AF = 0.5 for alpha particles at the GIT wall because half of all emitted particles will be completely absorbed by body tissue.
2.2.2.1. The special case of internal doses from radon progeny. One of the areas of uncertainty in 238U series dose calculations is the diffusion of radon within watery sediment and whether or not it can diffuse into and through small benthic organisms. For all practical purposes, there is no gaseous phase in the aquatic environment, only liquid. In a continuous liquid medium, the invertebrate body may pose no barrier to radon diffusion and, thus, radon and its progeny may be present in body tissues at the same concentrations as in external sediments. Alternatively, radon progeny may exist in body tissues only as a result of the decay of the 226Ra already absorbed into body tissues. Both assumptions were tested, assuming a variety of equilibrium states between 226Ra and radon.4 The reported doses from radon progeny in tissues were based on free diffusion of radon at 30% partial equilibrium with 226Ra in external sediment as was assumed for radon progeny in Table 2. Thus, for the short-lived radon progeny, Eq. (2) was used by employing an f 1 of unity and replacing Bq/kg d.w. sediment0.5 with Bq/kg w.w. sediment.
4 NCRP (1991) assumes 100% equilibrium of 226Ra with radon in external sediments but total loss of radon from 226Ra in invertebrate body tissues; Blaylock et al. (1993) assume 30% equilibrium of 226Ra with radon in external sediments.
Once within body tissues, 100% of the alpha energy from each nuclear decay will be absorbed by body tissues so that AF = 1. Beta and gamma radiations have greater penetration power and range than alpha particles and, thus, AF values for dose calculations depend on the overall size and shape of the organism and the kinetic energy of the radiation. For large organisms, AF values of unity for beta and gamma radiations of all energies have been used (Amiro, 1995; Amiro and Zach, 1993). As organisms become smaller, the AF decreases. The AF values for beta and gamma radiations were based on a point source dose distribution function developed by the National Council on Radiation Protection and Measurements (NCRP) (1991) and Blaylock et al. (1993) for uniformly distributed beta and gamma radiations in small aquatic insects and larvae, assumed to be 16-mg ellipsoid organisms (Fig. 2).5 The low-energy beta radiations from most of the decay series (U, 234Th, 230Th, 226Ra, 218Po, 214 Po, 210Pb, and 210Po) are completely absorbed within the organism, so that the internal AF = 1. Higher-energy beta radiations had lower internal AFs, which were 0.25 for
5
Average beta energies from Table 1 were multiplied by three and then estimated from graphs of AF vs. bmax in NCRP (1991) and Blaylock et al. (1993).
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Fig. 2. Geometry and dimensions (in mm) of benthic invertebrates. (a) Model aquatic organism for the point source dose distribution model (NCRP, 1991; Blaylock et al., 1993). (b) Measured and calculated dimensions for C. tentans and H. azteca in this study. 234
Pa, 0.38 for 214Bi, 0.6 for 210Bi, and 0.8 for 214Pb. Internal AFs for gamma radiations were 0.4 for all radionuclides in the decay series, except 214Pb and 214Bi (AF = 0.002 for both) and 210Bi, which does not produce gamma rays. The use of AF values from Blaylock et al. (1993) and NCRP (1991) for the 16-mg model organism may overestimate internal beta and gamma doses to the organisms used in this study by a factor of 2 – 4. The 16-mg organism was modeled as a solid cylinder. In contrast, the midge, C. tentans, and the amphipod, H. azteca, were considered
hollow cylinders around the GIT. Both species were smaller and four times thinner than the 16-mg model organism and C. tentans was twice as long. As shown in Fig. 2, the presence of the GIT contents would attenuate many of the weak beta and gamma radiation as they traverse body tissues. The decreased body thickness would further decrease internal AF values, relative to the model of NCRP (1991) and Blaylock et al. (1993). However, any overestimation in AF values here is accounted for by the principle of reciprocity in Eq. (3) below for external beta and gamma doses.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
2.2.3. Dose from radionuclides in the surrounding sediment External doses were calculated for beta and gamma, but not alpha, radiations in the external sediments. The short range of alpha particles prevents external alpha particles from penetrating the cuticle of most benthic organisms. Doses were calculated by Eq. (3), as follows: External dose ðmGy=yearÞ ¼ 1:61010 mJ=MeVðBq=kg w:w: s=yearÞ ðYE AFÞ
ð3Þ
where Bq/kg w.w. was the radionuclide concentration in wet weight sediment, determined from the moisture content data in Table 2, and YE values in million electron volts per disintegration were from Table 1. By the principle of reciprocity, NCRP (1991) and Blaylock et al. (1993) estimated external AF values to be 1 AF,
347
where AF = the internal AF value used in Eq. (2), for beta and gamma radiation within body tissues. For the lowenergy beta radiations from U, 234Th, 230Th, 226Ra, 218Po, 214 Po, 210Pb, and 210Po, the internal AF = 1 so the external AF = 0. The external AF values for higher-energy beta radiations were 0.75 for 234Pa, 0.62 for 214Bi, 0.4 for 210 Bi, and 0.2 for 214Pb. External AFs for gamma radiations were 0.6 for all radionuclides in the decay series, except 214 Pb and 214Bi (AF = 0.998 for both), with 210Bi not emitting gamma rays. Notably, external AFs increase as the energy of the radiation increases and the size of the organism decreases. External dose estimates, based on AFs from NCRP (1991) and Blaylock et al. (1993), were compared to other methods of external dose calculation (Table 4). An external AF of unity was employed on the assumption that small organisms, completely buried in sediment, receive the same average dose rate as the sediment itself, particularly for high-energy radiations. Dose rate constants from Beak
Table 4 External beta and gamma doses, calculated from dose coefficients from Beak Consultants (1985), AF values from Blaylock et al. (1993), and AF values of unitya Horseshoe Pond
Hidden Bay b
Tributary Creek
Blaylock
AF = 1
Beak Consultants
Blaylock
AF = 1
Beak Consultantsb
Blaylock
AF = 1
Beta doses (mGy/year) Natural U 27.2 234 Th 234 Pa 230 Th 226 Ra 222 Rn 218 Po 0.07 214 Pb 214 Bi 214 Po 210 Pb 0.15 210 Bi 210 Po Subtotal 27.43
– – 32.2 – – – – 0.01 0.08 – – 0.33 – 32.63
0.6 3.12 42.9 0 0 0 0 0.06 0.12 0 0.08 0.83 0 47.76
10.35
– – 12.2 – – – – 0 0.01 – – 0.12 – 12.37
0.23 1.19 16.3 0 0 0 0 0.01 0.02 0 0.03 0.29 0 18.08
0.26
– – 0.312 – – – – 0.003 0.023 – – 0.077 – 0.415
0.01 0.03 0.42 0 0 0 0 0.02 0.04 0 0.02 0.19 0 0.72
Gamma doses (mGy/year) Natural U 1.81 234 Th 234 Pa 230 Th 226 Ra 0 222 Rn 218 Po 0.19 214 Pb 214 Bi 214 Po 210 Pb 210 Bi 210 Po Subtotal 2.01
0.08 0.31 0.37 0 0 0 0 0.05 0.28 0 0.01 – 0 1.09
0.14 0.52 0.61 0 0 0 0 0.05 0.28 0 0.01 – 0 1.6
0.69
0.05 0.2 0.23 0 0 0 0 0.01 0.04 0 0 – 0 0.52
0.02
0.71
0.03 0.12 0.14 0 0 0 0 0.01 0.03 0 0 – 0 0.33
0.08
0.001 0.003 0.004 0 0.001 0 0 0.038 0.228 0 0.001 – 0 0.276
0.001 0.005 0.006 0 0.001 0 0 0.038 0.229 0 0.002 – 0 0.28
Total external dose
32.9
48.16
9.62
10.99
16.1
0.39
0.69
1
Beak Consultants
a
28.73
b
0.01
0.05
10.41
0 0.02
0.02
0.03
0.32
0 0.06
Dashes indicate external doses of zero. Doses listed as 0.00 or 0.000 are < 0.005 and < 0.0005, respectively. b The dose coefficients from Beak (1985) in units of mGy/h per Bq/kg wet weight sediment were as follows: 0.3 for beta and 0.02 for gamma from 238U through 234U; 0.0037 for gamma from 226Ra; 0.21 for beta and 0.6 for gamma from the short-lived radon progeny, and; 0.04 for beta from 210Pb.
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Consultants (1985) were also applied, which assumed a sediment density of 1.5 g/cm3.
3. Results The sediment data clearly showed that all radionuclides measured were elevated at Horseshoe Pond, the site closest to the outflow of uranium mining effluent (Table 2). The downstream site at Hidden Bay had lower concentrations
and the control site, Tributary Creek, the lowest of all. Notably, uranium and 210Pb concentrations were high in both Horseshoe Pond and Hidden Bay, whereas 230Th, 226 Ra, and 210Po concentrations were similar in Hidden Bay and Tributary Creek. Calculated doses from all individual radionuclides, according to the source and type of radiation, are shown in Appendix A for C. tentans and Appendix B for H. azteca. Absorbed dose estimates to invertebrates from all 238U series radionuclides were an order of magnitude higher for
Fig. 3. Absorbed (a) and equivalent (b) doses to larval midges (C. tentans) and adult amphipods (H. azteca) by radionuclide. HP = Horseshoe Pond, HB = Hidden Bay, and TC = Tributary Creek. TC enlarged for clarity.
Fig. 4. Absorbed (a) and equivalent (b) doses to larval midges (C. tentans) and adult amphipods (H. azteca) by radiation source and type. HP = Horseshoe Pond, HB = Hidden Bay, and TC = Tributary Creek. TC enlarged for clarity.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
Horseshoe Creek and Hidden Bay than for the control site, Tributary Creek (Fig. 3a). Total doses from radionuclides within the three sources (GIT, body tissues, and external sediment) added up to 60, 19, and 3.2 mGy/year for both C. tentans and H. azteca from the Horseshoe Pond, Hidden Bay, and Tributary Creek sites, respectively. Most of the absorbed dose at the impacted sites was from external beta radiation, while it was internal alpha radiation at the control site (Fig. 4a). The most important contributors to absorbed doses for the two invertebrate species at Horseshoe Pond and Hidden Bay were 234Pa (54 – 65%), assumed to be in equilibrium with uranium, followed by uranium itself (25 – 31%). At Tributary Creek, most of the dose was internal alpha radiation in body tissues from 210Po (52%) and the short-lived radon progeny, 218 Po and 214Po (24%) (Fig. 3a). When a radiation weighting factor of 20 was employed for alpha radiation, equivalent doses from Horseshoe Pond, Hidden Bay, and Tributary Creek sediment were 540, 140, and 54 mGy/year, respectively, for C. tentans, and 560, 140, and 54 mGy/year, respectively, for H. azteca (Fig. 3b). For alpha radiations within the GIT, doses were 13 –14% higher in H. azteca vs. C. tentans because the GIT diameter was smaller, raising the AF. Alpha radiation from U in the GIT and body tissues accounted for 64– 70% of the total equivalent dose in both species using sediment from the Horseshoe Pond and Hidden Bay sites. In contrast, internal alpha radiation was 2% for U and 62% for 210Po for the Tributary Creek control site (Fig. 3b). External doses were calculated by three different methods (Table 4). The dose coefficients from Beak Consultants (1985) yielded the lowest dose estimates. The external doses based on an external AF of unity were 44 –46% higher than doses based on AFs from Blaylock et al. (1993). The greatest differences were for low-energy beta and gamma radiations, which are mostly absorbed from internal rather than external sources. There was almost no difference in the doses from the high-energy gamma radiations from 214Pb (0.249 MeV) and 214Bi (1.51 MeV). Thus, the use of the easier-to-calculate external AF of unity is justified for those gamma radiations, which are >0.25 MeV, or as a screening method for doses to extremely small organisms because their size decreases internal AFs and raises external AFs near unity. The dose estimates based on the AFs from Blaylock et al. (1993) were used in all other tables and figures. Overall, external beta radiation from 234Pa, followed by alpha radiation from U, contributed the majority of the total absorbed dose at the two impacted sites, whereas alpha radiation from 210Po, 218Po, and 214Po in body tissues was most important at the control site. Gamma radiation constituted 2– 3% at all three sites, being highest at the control site, primarily from the short-lived radon decay product, 214 Bi. If a radiation weighting factor of 20 was applied for alpha radiation, internal alpha radiation was estimated to contribute 91 – 99% of the total equivalent dose, either from
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U at the two uranium-contaminated sites, or from 210Po and the alpha-emitting radon progeny at the control site.
4. Discussion This study was one of the first to consider the doses from alpha emitters within the GIT as a separate source of irradiation to aquatic organisms. Modeling the GIT as a separate source is equivalent to what Pentreath and Woodhead (2000) refer to as a Type B dosimetric model (i.e., a hollow cylinder geometry). The results show that the GIT is an important source of alpha irradiation for small aquatic organisms, adding a 10% absorbed dose increment and a 25% equivalent dose increment to benthic invertebrates in the vicinity of uranium mine operations. The results further show that the smaller organism (H. azteca, 4.4 mg) receives a greater dose from alpha radiations within the GIT than the larger organism (C. tentans, 12 mg). The difference in dose is more evident in equivalent dose calculations when a weighting factor is used for alpha radiation. As the diameter of the GIT decreases, the AF value for ingested alpha emitters increases because more alpha particles within the GIT reach body tissues. The AF value is 0.22 in H. azteca (4.4 mg), 0.14 in C. tentans (11.7 mg), and 0.005 in humans (70 kg, ICRP, 1979). If the GIT weight and diameter are known, AF values for internal alpha radiation doses can be calculated for other species and would be expected to play an increasingly important role in the dose to the smallest organisms. The invertebrate species considered here are ‘‘collectors – gatherers’’. They preferentially consume organic matter from sediment rather than whole sediment. Because organic matter has a greater capacity to adsorb radionuclides than whole sediment, radionuclide concentrations in the GIT may be even higher than the whole sediment concentrations, assumed to represent GIT contents in Eq. (1). The doses calculated in this study are highly dependent on assumptions of equilibrium made for a number of the 238 U decay series radionuclides. Thus, 234Pa delivered the majority of the absorbed dose to both invertebrate species because of its assumed equilibrium with uranium. Rapid establishment of equilibrium between uranium and 234Pa may also be a matter of some importance in terrestrial environments, potentially contaminated with depleted uranium, because of the increased external beta radiation exposure from 234Pa. The degree to which radon progeny are in equilibrium with 226Ra in sediment deserves further investigation, as does whether or not radon can diffuse into body tissues in aquatic environments. The NCRP (1991) assumes 100% equilibrium in sediment, but also claims that radon can escape to the atmosphere from lake sediments and that no radon would be present in invertebrate body tissues. In this study, only 30% equilibrium was assumed between radon
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and 226Ra in sediment, which may underestimate external dose from 214Bi. By assuming that radon and its progeny can equilibrate between sediment and invertebrate body tissue, the internal alpha dose from 218Po and 214Po may be overestimated. Appendices A and B show that alpha radiation from 218Po and 214Po in body tissues account for only 5 – 9% of the equivalent dose at the U contaminated sites but 29% at the Tributary Creek control site. Given the importance of these radionuclides in estimating human dose, it would seem unlikely that they should be ignored when estimating doses to benthic organisms. The doses calculated for the benthic invertebrates in Horseshoe Pond reached 540– 560 mGy/year, assuming a weighting factor of 20 for alpha. These doses were dominated by the alpha radiations in the GIT and body tissues from natural U concentrations of 50,484 Bq/kg in the sediment. These dose estimates were one to two orders of magnitude lower than NCRP (1991) dose estimates to insects near Tailings Creek in northern Saskatchewan where 238 U concentrations were only 150 Bq/kg. The differences in dose estimates were due to: (1) much higher 226Ra and 210 Po concentrations, measured or assumed in the Tailings Creek sediments; (2) the assumption of 100% equilibrium among 226Ra, radon, short-lived radon progeny, and 210Pb; and (3) the measurements and/or transfer parameters used to establish radionuclide concentrations in benthic invertebrate tissue. Equivalent doses of 540 –560 mGy/year can be compared to benchmark dose limits, which have been estimated to protect animal populations from detrimental effects, primarily such effects as early mortality, reduced reproductive success, and scorable cytogenetic damage (Pentreath, 1991; Pentreath and Woodhead, 2000). Dose limits from gamma radiation are estimated at 1 – 10 mGy/day or 360 –3600 mGy/year to ensure that mortality and reproductive effects will not occur in the most sensitive species, usually mice (International Atomic Energy Agency (IAEA), 1992; United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR), 1996). However, species differ in radiosensitivity and benthic macroinvertebrates are likely to be less radiosensitive than mice and other mammals (Rose, 1992). Doses of 10 mGy/day (3600 mGy/year) are generally assumed to protect aquatic organisms (NCRP, 1991; Pentreath, 1991; IAEA, 1992; UNSCEAR, 1996). For example, the number of egg capsules was reduced in the pond snail Physa at doses of 6.5 mGy/day. However, overall egg numbers only decreased at doses of 240 –1200 mGy/day. Birth rates in the cladoceran Daphnia did not drop until doses reached >4600 mGy/day (Marshall and Cooley in Blaylock et al., 1993). The potential radiation impacts of sediment-associated uranium on benthic organisms in Horseshoe Pond and Hidden Bay depends on the weighting factor used for alpha radiation and the endpoint considered. Trivedi and Gentner (2000) suggest that ecodosimetry weighting factors (eR) could be determined for three different endpoints for a
given organism: Level 1 for effects on individual organisms; Level 2 for effects on ecosystem sustainability; and Level 3 for genetic/hereditary effects. RBE values in the literature range from < 1 to 370 for alpha radiation, depending on the species and the endpoint considered. The RBEs are generally lower for genetic effects and in vitro studies, and higher for effects in vivo, such as the high RBE of 370 found for mouse oocyte survival (Samuels, 1966). Kocher and Trabalka (2000) suggest that a different weighting factor for animals will require the use of a unit for equivalent dose other than the Sievert. In this study, the equivalent doses to C. tentans and H. azteca in Horseshoe Pond (540 – 560 mGy/year) surpassed the lowest reproductive dose limit of 360 mGy/year, solely because a radiation weighting factor of 20 was used for alpha radiation. However, the 360-mGy/year limit is conservative, as it is generally applied to mammals rather than aquatic organisms. The long half-life of 4.5 billion years for 238U means that high mass concentrations of uranium must be ingested before it can deliver much of an alpha radiation dose. At high concentrations, however, the chemical toxicity of uranium may be more detrimental than the radiological dose. The uranium concentrations in sediment from Horseshoe Pond (2000 mg/kg or 50,484 Bq/kg), but not Hidden Bay (540 mg/kg or 13,616 Bq/kg), were capable of exerting toxic effects on H. azteca.6 Radiological effects are possible at the radiation doses calculated in this study, but are strongly dependent on the assumptions made regarding the weighting factor used for alpha radiation.
5. Conclusions A new method for calculating radiation doses to benthic invertebrates from measured sediment concentrations was described, which included dose contributions from sediment within the GIT. Doses from all radionuclides in the uranium-238 decay series were calculated to two representative benthic invertebrate species, based on radionuclide concentrations in sediment collected from sites near the Rabbit Lake uranium mining operation in northern Saskatchewan. Absorbed dose estimates were up to 20 times higher at the two uranium-contaminated sites compared to the control site. Most of the dose increment was from alpha radiation from uranium within the GIT, of which only 2% may be absorbed into body tissues, and from external beta radiation from 234Pa in the surrounding sediment, assumed to be at equilibrium with U in the sediment. The radiological doses estimated here, coupled
6 Spiked sediment toxicity tests with UO2(NO3)26H2O yielded 10-day LC50 values of 10,551 mg/kg d.w. for C. tentans and 2442 mg/kg d.w. for H. azteca; 10-day LC25 values were 4859 and 1449 mg/kg, respectively. For H. azteca, the 10-day EC50 for growth impairment was estimated at 964 mg/kg d.w. (Liber and White Sobey, 2000).
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
with the additional chemotoxic stress of high uranium concentrations, may affect the survival, growth, or reproduction of these or similar species in the wild. The importance of the radiological effect depends on a radiation weighting factor of 20 for alpha radiation being biologically relevant to effects in invertebrate species. Additional work is needed to properly address these questions, particularly for the alpha-emitting radionuclides that prevail around uranium mines.
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Acknowledgments Funding for this research was provided by Saskatchewan Environment and Resource Management, Cameco, Cogema Resources, Cigar Lake Mining, the Canadian Nuclear Safety Commission, and the Department of Fisheries and Oceans. We would like to thank Dr. Sylvain St. Pierre, Mrs. Maisie Shiell, and two anonymous reviewers for their helpful comments on earlier versions of this manuscript.
Appendix A. Absorbed doses (Abs in mGy/year) and equivalent doses (Equiv in mGy/year20 for alpha) to C. tentans larvae from radionuclide concentrations in sediment Total Percent of Eq. (1) Eq. (2) Eq. (3) dose total dose Alpha in GIT Horseshoe Pond Natural U 5.9 234 Th 234 Pa 230 Th 1.7E 02 226 Ra 7.5E 02 222 Rn 218 Po 2.8E 02 214 Pb 214 Bi 214 Po 3.6E 02 210 Pb 210 Bi 210 Po 7.5E 02 Total 6.11 Hidden Bay Natural U 1.6E + 00 234 Th 234 Pa 230 Th 4.5E 03 226 Ra 6.6E 03 222 Rn 218 Po 2.5E 03 214 Pb 214 Bi 214 Po 3.2E 03 210 Pb 210 Bi 210 Po 2.0E 02 Total 1.62 Tributary Creek natural U 1.8E 02 234 Th 234 Pa 230 Th 3.3E 03 226 Ra 7.1E 03
Alpha in tissue
Beta in tissue
Gamma in tissue
12
0.03 3.8E 03 1.3E 02 2.6E 06 1.1E 03 2.0E 06 2.6E 06 4.3E 02 4.6E 02 1.5E 07 3.9E 02 6.1E 02 5.6E 08 0.24
0 2.4E 04 2.9E 04 1.1E 07 8.2E 04 3.0E 05 6.8E 07 9.2E 05 5.4E 04 6.2E 06 2.0E 03 0.0E + 00 2.3E 06 0.007
8.2E 03 1.0E 03 3.5E 03 6.8E 07 9.6E 05 2.5E 07 3.3E 07 5.4E 03 5.7E 03 1.9E 08 9.7E 03 1.5E 02 1.5E 08 0.05
7.2E 04 6.4E 05 7.8E 05 2.9E 08 7.2E 05 3.7E 06 8.5E 08 1.2E 05 6.8E 05 7.7E 07 4.9E 04 0.0E + 00 6.1E 07 0.002
9.4E 05 1.2E 05 4.0E 05 5.0E 07 1.0E 04
8.2E 06 7.4E 07 8.9E 07 2.1E 08 7.8E 05
8.4E 04 1.5E + 00 1.1E + 00
1.5E + 00
3.7E + 00 19.35
3.1E + 00
2.2E 04 1.3E 01 1.4E 01
1.8E 01
9.6E 01 4.54
3.6E 02
1.6E 04 1.4E 01
Beta external
32
1.1E 02 7.5E 02
Gamma external
Abs
Equiv
Abs
Equiv
2.9E 06 1.06
17.571 0.297 32.505 0.018 1.551 0.000 1.167 0.100 0.393 1.495 0.048 0.392 3.781 59.32
349.238 0.297 32.505 0.357 30.945 0.000 23.331 0.100 0.393 29.899 0.048 0.392 75.620 543.12
29.6 0.5 54.8 0.0 2.6 0.0 2.0 0.2 0.7 2.5 0.1 0.7 6.4 100
64.3 0.1 6.0 0.1 5.7 0.0 4.3 0.0 0.1 5.5 0.0 0.1 13.9 100
3.1E 02 1.1E 01 1.3E 01 5.0E 05 3.1E 04 5.5E 06 1.3E 07 5.7E 03 3.4E 02 1.2E 06 2.1E 03 0.0E + 00 1.0E 06 0.32
4.748 0.112 12.358 0.005 0.137 0.000 0.144 0.013 0.049 0.185 0.012 0.130 0.981 18.87
94.204 0.112 12.358 0.094 2.734 0.000 2.885 0.013 0.049 3.697 0.012 0.130 19.612 135.90
25.2 0.6 65.5 0.0 0.7 0.0 0.8 0.1 0.3 1.0 0.1 0.7 5.2 100
69.3 0.1 9.1 0.1 2.0 0.0 2.1 0.0 0.0 2.7 0.0 0.1 14.4 100
7.9E 04 2.8E 03 3.4E 03 8.1E 05 7.5E 04
0.055 0.003 0.314 0.004 0.148
1.081 0.003 0.314 0.069 2.940
1.7 0.1 9.8 0.1 4.6
2.0 0.0 0.6 0.1 5.5
0.082 2.9E 01 3.5E 01 1.3E 04 2.5E 03 4.4E 05 1.0E 06 4.6E 02 2.7E 01 9.3E 06 6.1E 03
3.3E 01 32.55
1.2E + 01
1.4E 03 9.3E 03
1.2E 01 12.35
3.1E 01
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222
Rn Po 214 Pb 214 Bi 214 Po 210 Pb 210 Bi 210 Po Total 218
2.7E 03
3.4E 01
3.4E 03
4.3E 01
3.3E 02 0.068
1.6E + 00 2.58
6.0E 07 7.9E 07 1.3E 02 1.4E 02 4.5E 08 2.9E 03 4.5E 03 2.5E 08 0.034
8.8E 06 2.0E 07 2.7E 05 1.6E 04 1.8E 06 1.5E 04 0.0E + 00 1.0E 06 4.4E 04
3.2E 03 2.2E 02
7.7E 02 0.41
1.3E 05 3.0E 07 1.4E 02 8.1E 02 2.8E 06 1.4E 03 0.0E + 00 4.0E 06 0.10
0.000 0.341 0.030 0.117 0.437 0.005 0.082 1.669 3.20
0.000 6.817 0.030 0.117 8.736 0.005 0.082 33.383 53.58
0.0 10.6 0.9 3.6 13.6 0.1 2.5 52.1 100
0.0 12.7 0.1 0.2 16.3 0.0 0.2 62.3 100
Appendix B. Absorbed doses (Abs in mGy/year) and equivalent doses (Equiv in mGy/year20 for alpha) to H. azteca adults from radionuclide concentrations in sediment Total Percent of Eq. (1) Eq. (2) Eq. (3) dose total dose Alpha in GIT Horseshoe Pond Natural U 6.7E + 00 234 Th 234 Pa 230 Th 1.9E 02 226 Ra 8.5E 02 222 Rn 218 Po 3.2E 02 214 Pb 214 Bi 214 Po 4.1E 02 210 Pb 210 Bi 210 Po 8.6E 02 Total 6.98 Hidden Bay Natural U 1.8E + 00 234 Th 234 Pa 230 Th 5.1E 03 226 Ra 7.5E 03 222 Rn 218 Po 2.8E 03 214 Pb 214 Bi 214 Po 3.6E 03 210 Pb 210 Bi 210 Po 2.2E 02 Total 1.85 Tributary Creek Natural U 2.1E 02 234 Th 234 Pa 230 Th 3.7E 03
Alpha in tissue
Beta in tissue
Gamma in tissue
1.2E + 01
3.0E 02 3.8E 03 1.3E 02 2.6E 06 1.1E 03 2.0E 06 2.6E 06 4.3E 02 4.6E 02 1.5E 07 3.9E 02 6.1E 02 5.6E 08 0.24
2.7E 03 2.4E 04 2.9E 04 1.1E 07 8.2E 04 3.0E 05 6.8E 07 9.2E 05 5.4E 04 6.2E 06 2.0E 03 0.0E + 00 2.3E 06 0.007
8.2E 03 1.0E 03 3.5E 03 6.8E 07 9.6E 05 2.5E 07 3.3E 07 5.4E 03 5.7E 03 1.9E 08 9.7E 03 1.5E 02 1.5E 08 0.05
7.2E 04 6.4E 05 7.8E 05 2.9E 08 7.2E 05 3.7E 06 8.5E 08 1.2E 05 6.8E 05 7.7E 07 4.9E 04 0.0E + 00 6.1E 07 0.002
9.4E 05 1.2E 05 4.0E 05 5.0E 07
8.2E 06 7.4E 07 8.9E 07 2.1E 08
8.4E 04 1.5E + 00 1.1E + 00
1.5E + 00
3.7E + 00 19.35
3.1E + 00
2.2E 04 1.3E 01 1.4E 01
1.8E 01
9.6E 01 4.54
3.6E 02
1.6E 04
Beta external
3.2E + 01
1.1E 02 7.5E 02
Gamma external
Abs
Equiv
Abs
Equiv
2.9E 06 1.06
18.401 0.297 32.505 0.020 1.562 0.000 1.171 0.100 0.393 1.500 0.048 0.392 3.792 60.18
365.846 0.297 32.505 0.405 31.156 0.000 23.411 0.100 0.393 30.001 0.048 0.392 75.832 560.39
30.6 0.5 54.0 0.0 2.6 0.0 1.9 0.2 0.7 2.5 0.1 0.7 6.3 100
65.3 0.1 5.8 0.1 5.6 0.0 4.2 0.0 0.1 5.4 0.0 0.1 13.5 100
3.1E 02 1.1E 01 1.3E 01 5.0E 05 3.1E 04 5.5E 06 1.3E 07 5.7E 03 3.4E 02 1.2E 06 2.1E 03 0.0E + 00 1.0E 06 0.32
4.972 0.112 12.358 0.005 0.138 0.000 0.145 0.013 0.049 0.185 0.012 0.130 0.983 19.10
98.684 0.112 12.358 0.107 2.752 0.000 2.892 0.013 0.049 3.706 0.012 0.130 19.668 140.48
26.0 0.6 64.7 0.0 0.7 0.0 0.8 0.1 0.3 1.0 0.1 0.7 5.1 100
70.2 0.1 8.8 0.1 2.0 0.0 2.1 0.0 0.0 2.6 0.0 0.1 14.0 100
7.9E 04 2.8E 03 3.4E 03 8.1E 05
0.057 0.003 0.314 0.004
1.133 0.003 0.314 0.078
1.8 0.1 9.8 0.1
2.1 0.0 0.6 0.1
8.2E 02 2.9E 01 3.5E 01 1.3E 04 2.5E 03 4.4E 05 1.0E 06 4.6E 02 2.7E 01 9.3E 06 6.1E 03
3.3E 01 32.55
1.2E + 01
1.4E 03 9.3E 03
1.2E 01 12.35
3.1E 01
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Ra Rn 218 Po 214 Pb 214 Bi 214 Po 210 Pb 210 Bi 210 Po Total
8.1E 03
1.4E 01
3.1E 03
3.4E 01
222
3.9E 03
4.3E 01
3.8E 02 0.077
1.6E + 00 2.58
1.0E 04 6.0E 07 7.9E 07 1.3E 02 1.4E 02 4.5E 08 2.9E 03 4.5E 03 2.5E 08 0.034
7.8E 05 8.8E 06 2.0E 07 2.7E 05 1.6E 04 1.8E 06 1.5E 04 0.0E + 00 1.0E 06 4.4E 04
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3.2E 03 2.2E 02
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An estimation of radiation doses to benthic invertebrates from sediments collected near a Canadian uranium mine Patricia Thomas*, Karsten Liber1 Toxicology Centre, University of Saskatchewan, 44 Campus Drive, Saskatoon, Saskatchewan, Canada S7N 5B3 Received 6 February 2001; accepted 1 June 2001
Abstract A new method is described for calculating radiation doses to benthic invertebrates from radionuclide concentrations in freshwater sediment. Both internal and external radiation doses were estimated for all 14 principal radionuclides of the uranium-238 decay series. Sediments were collected from three sites downstream of a uranium mining operation in northern Saskatchewan, Canada. Sediments from two sites, located approximately 1.6 and 4.4 km downstream from mining operations, yielded absorbed doses to both larval midges, Chironomus tentans, and adult amphipods, Hyalella azteca, of 59 – 60 and 19 mGy/year, respectively, compared to 3.2 mGy/year for a nearby control site. External beta radiation from protactinium-234 (234Pa) and alpha radiation from uranium (U) contributed most of the dose at the impacted sites, whereas polonium-210 (210Po) was most important at the control site. If a weighting factor of 20 was employed for the greater biological effect of alpha vs. beta and gamma radiation, then total equivalent doses rose to 540 – 560 mGy/year at the site closest to uranium operations. Such equivalent doses are above the 360-mGy/year no-observed-effect level for reproductive effects in vertebrates from gamma radiation exposure. Data are not available to determine the effect of such doses on benthic organisms, but they are high enough to warrant concern. Detrimental effects have been observed in H. azteca at similar uranium concentration in laboratory toxicity tests, but it remains unclear whether the radiotoxicity or the chemotoxicity of uranium is responsible for these effects. D 2001 Elsevier Science Ltd. All rights reserved. Keywords: Uranium mining; Sediment contamination; Radiation dose; Benthic invertebrates
1. Introduction Animals and humans receive natural background radiation doses from cosmic rays, gamma rays arising from rocks and soil, inhalation of radon gas, and ingestion of radionuclides with food, water, and soil. In humans, annual background doses from all of these sources range from 1 to 10 milliSievert (mSv) (Committee on the Biological Effects of Ionizing Radiation (BEIR), 1990; Beak Consultants, 1995). Animals often receive higher doses than humans because they ingest more soil or sediment with food items, which raises their body burden of radionuclides. In addition, they live outdoors, which increases their exposure to cosmic rays and terrestrial gamma radiation. * Corresponding author. Tel.: +1-306-244-4807; fax: +1-306-931-1664. E-mail addresses: [email protected] (P. Thomas), [email protected] (K. Liber). 1 Also corresponding author. Tel.: + 1-306-966-7444; fax: + 1-306931-1664.
Based on their body burdens of five uranium series radionuclides (U, 226Ra, 210Pb, 210Bi, and 210Po), annual background doses of 18 –38 mGy have been calculated to small mammals, birds, and caribou from unpolluted sites in northern Canada (Thomas, 1999, 2000). These dose calculations employed the same weighting factor of 20, as is used in human dose calculations, to account for the greater relative biological effect (RBE) of alpha vs. beta and gamma radiation. Some small mammals living close to uranium mine tailings received doses of up to 230 mGy/year, mostly from the ingestion of 226Ra (Thomas, 2000). These doses would have been even higher if external radiation doses, as well as all 14 principal radionuclides in the uranium decay series, had been considered. One group of animals that may receive elevated radiation doses are the benthic macroinvertebrates, which burrow in sediment. Radionuclides often accumulate in sediment, such as in depositional zones downstream of uranium mining operations. Benthic macroinvertebrates are important components of aquatic food chains and, thus, an
0160-4120/01/$ – see front matter D 2001 Elsevier Science Ltd. All rights reserved. PII: S 0 1 6 0 - 4 1 2 0 ( 0 1 ) 0 0 0 8 5 - X
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effort to estimate radiation doses to such organisms should be made. However, radiation doses to many benthic organisms are difficult to calculate because these organisms are too small to allow for direct radiochemical analyses. It is also difficult to distinguish between radionuclides absorbed into body tissues, those adsorbed to the external body surfaces and those that are within the animal gastrointestinal tract (GIT). This publication describes the calculations of internal and external radiation doses to two benthic macroinvertebrates — the larval midge, Chironomus tentans, and the amphipod, Hyalella azteca. The dose calculations are based solely on the radionuclide concentrations in the sediment which surround these organisms and which they ingest. The work presented here was part of a larger study designed to evaluate the potential impact of effluent from the Rabbit Lake uranium mine in northern Saskatchewan, Canada, on adjacent aquatic communities. Differences in natural benthic communities downstream from the mining operation, correlation of such differences to uranium and other contaminant exposure, and assessment of sediment toxicity will be described elsewhere (Liber et al., in preparation).
2. Methods 2.1. Sediment sampling and analysis Sediment samples for radiochemical analysis were collected from three different locations (Horseshoe Pond, Hidden Bay, and Tributary Creek) near from the Rabbit
Lake uranium mine in northern Saskatchewan, Canada (581103800N, 1034205700W). Horseshoe Pond is located approximately 1.6 km downstream from the mine effluent treatment plant. The pond empties into Horseshoe Creek, which flows into Hidden Bay on Wollaston Lake, approximately 4.4 km further downstream. Tributary Creek, the control site, also empties into Horseshoe Creek, but originates from an area north of Horseshoe Creek and is not impacted by any effluent or mining activity. Sediments were collected from all three sites with a small Ekman grab sampler (15 15 cm). Care was taken to ensure that all sediment samples were collected from the top 5 – 10 cm sediment layer. Three replicate samples were collected from each site and placed in separate 1-l Nalgene bottles for transportation in ice-packed coolers to Saskatoon, SK. Five radionuclides in dry sediments were analyzed by the Saskatchewan Research Council: uranium (U), thorium-230 (230Th), radium-226 (226Ra), lead-210 (210Pb), and polonium-210 (210Po). Uranium was analyzed by delayed neutron counting (detection limit = 0.1 mg/g of sample). 230Th, 226 Ra, and 210Po were chemically extracted and analyzed by alpha spectroscopy (detection limits 0.02, 0.005, and 0.005 Bq/g, respectively). 210Pb was analyzed by precipitation of its decay product bismuth-210 (210Bi) and counted by beta spectroscopy (detection limit = 0.02 Bq/g). Detailed descriptions of these procedures can be found in Canadian Mining and Energy Technologies (CANMET, 1978). Not all radionuclides in the 238U decay series were measured in sediments. However, the entire decay series (Table 1) was considered in the radiation dose calculations
Table 1 Uranium-238 decay series Average energy (MeV) per radiation type Radionuclide
Half-life b
Natural uranium Uranium-238b Thorium-234 Protactinium-234 Uranium-234b Thorium-230 Radium-226 Radon-222(gas) Polonium-218 Lead-214 Bismuth-214 Polonium-214 Lead-210 Bismuth-210 Polonium-210 Lead-206 (stable)
4.5 billion years 24.1 days 1.17 min 244,500 years 80,000 years 1,600 years 3.85 days 3.05 min 26.8 min 19.7 min 0.00015 s 22.2 years 4.97 days 138 days
Alpha 4.55 see natural uranium above none none see natural uranium above 4.74 4.86 ignored 6.11 none none 7.83 none none 5.4
Betaa
Gammaa 2
2.61 10 3
5.92 10 2 8.20 10 1
9.34 10 3 1.13 10 2
1.46 10 2 3.59 10 3 1.09 10 5 1.42 10 5 2.91 10 1 6.48 10 1 8.19 10 7 3.80 10 2 3.89 10 1 8.18 10 8
1.55 10 3 6.74 10 3 3.98 10 4 9.12 10 6 2.48 10 1 1.46 10 0 8.33 10 5 4.81 10 3 none 8.50 10 6
1.19 10
Average energies are calculated from Yield Energy values in ICRP (1983) data tables. a All electron energies from beta particles, as well as conversion and Auger electrons, are included under ‘‘Beta’’ in the table. All photon energies from gamma, as well as X-rays, are included under ‘‘Gamma’’ in the table. b By mass, natural uranium is composed of 99.3% 238U and 0.7% fissile 235U (half-life = 710 million years). 238U and 234U are in equilibrium and, thus, contribute equally to the activity of natural U, although the mass of 234U present is much smaller than that of 238U. Including all three radionuclides, 1 mg/kg of natural U is equivalent to 25.2 Bq/kg.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
by making the following assumptions regarding the chemistry and mobility of each radionuclide in the environment: (1) 234Th, 234Pa, and 234U were in equilibrium with and, thus, had similar activities to 238U; (2) radon gas and its short-lived progeny were in partial equilibrium (30%) with the parent, 226Ra, as Blaylock et al. (1993) has assumed in other studies; and (3) 210Bi was in equilibrium with its parent, 210Pb. 210Po was measured separately and corrected for decay of unsupported 210Po, as well as ingrowth from 210Pb between the date of sampling and the date of analysis. 2.2. Dose calculations Total radiation doses from the 14 principal uranium series radionuclides were calculated to two benthic organisms, based on the mean radionuclide concentrations in sediments from three sites: Horseshoe Pond, Hidden Bay, and Tributary Creek, a control site (Table 2). Eqs. 1 to 3 (below) consider three sources of radionuclides which can irradiate the organism: (1) alpha irradiation from sediment within the GIT; (2) the fraction of radionuclides in the GIT, which are absorbed into body tissues; and (3) external radiation from radionuclides in the sediment outside the organism. From each of these three sources, the energy (in joules) absorbed per kilogram of body tissue (excluding the GIT contents) was determined for the alpha, beta, and gamma radiations from each of the 14 238U decay series radionuclides listed in Table 1.
Table 2 Radionuclide concentrations (Bq/kg dry weight) in sediments collected from three sites near the Rabbit Lake uranium mine in northern Saskatchewan, September 1997 (N = 3) Horseshoe Pond
Hidden Bay
Tributary Creek (control site)
Radionuclides measured Unata Mean S.D. 230 Th Mean S.D. 226 Ra Mean S.D. 210 Pb Mean S.D. 210 Po Mean S.D.
50,484 7,288 140 30 600 157 2,057 2,054 544 342
13,616 979 37 21 53 12 507 51 141 35
156 25 30 17 57 31 153 68 240 156
Sediment characteristics Particle size classes (%)b Sand ( > 53 mm) Silt (2 – 52 mm) Clay ( < 2 mm) Water content (%)
27.5 ± 3.7 36.0 ± 2.2 36.5 ± 1.6 79.5 ± 2.6
35.9 ± 6.6 56.5 ± 5.7 7.6 ± 1.0 71.1 ± 6.6
68.3 ± 28.9 28.8 ± 26.5 2.9 ± 2.3 35.9 ± 11.2
a By dividing natural U activity by 25.2, the concentration in milligram per kilogram (ppm) can be found. b Mean particle size ( ± 1 S.E.) for separate samples collected with a corer from a depth of 0 – 6 cm.
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2.2.1. Alpha dose from radionuclides within the GIT Internal radiation doses from alpha-emitting radionuclides in sediment within the GIT were determined by Eq. (1): Dose ðmGy=yearÞ ¼ ð1:6 1010 mJ=MeVðBq=kg d:w: 0:5 kg of source w:w: s=yearÞ ðYE AFÞÞ=ðkg of target w:w:Þ ð1Þ where 1. a conversion factor of 1.6 10 10 mJ/MeV converts absorbed energy in million electron volts (MeV) to millijoules (mJ); 2. Bq/kg d.w. is the radionuclide concentration (dry weight) in sediment within the GIT; where 1 Bq = 1 nuclear disintegration/s. These concentrations were multiplied by 0.5 to account for the 50% moisture content assumed for sediment within the GIT. A factor of 50% was chosen to reflect fine sediment in the GIT being slightly more compacted than in the external sediment; 3. kg of source w.w. refers to the mass of the GIT contents (wet weight); 4. there are 3.156 107 s/year; 5. YE are the average ‘‘yield times energy’’ values in million electron volts from Table 1, determined for the alpha radiations from each radionuclide from the International Commission of Radiological Protection (ICRP, 1983); 6. AF is the absorbed fraction, or the fraction of radiation energy emitted from a radionuclide within the GIT that is absorbed by body tissues; and 7. kg of target w.w. is the body mass of the organism minus the GIT contents (wet weight). Factors 2 – 4, listed above from Eq. (1), determine the number of disintegrations from a radionuclide occurring within the GIT per year. Factors 5– 7 determine the energy from each nuclear disintegration in the GIT source absorbed per kilogram of body target tissue (in million electron volts per disintegration per kilogram). When million electron volts are converted to millijoules by Factor 1, the result is the absorbed dose to the organism in milligray per year, where 1 mGy = 1 mJ/kg. Eq. (1) was calculated separately for the alpha radiations from each radionuclide and then summed for all radionuclides. Body length, width, and mass were measured in several organisms of both macroinvertebrate species (the midge, C. tentans, and the amphipod, H. azteca) reared in beakers containing uncontaminated sediment from Wollaston Lake in northern Saskatchewan (Table 3). The mass of the GIT was calculated as the difference between full organisms and those purged of their GIT contents. The mass of the whole body minus the GIT contents represented body tissues and
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Table 3 Mean morphometric measurements of C. tentans and H. azteca C. tentans (fourth instar) Body mass measurements (n = 5) Body mass with full GIT a (mg wet weight) Target tissue for radiation: body mass with purged GIT (mg wet weight) Source tissue: mass of GIT contents (mg wet weight) GIT as percent of body mass with purged GIT Body size measurements (n = 6) Body length (mm) Lateral body width (mm) Diameter of GIT (mm) Percent of GIT within 50 mm of the GIT wallb
H. azteca (adult)
11.68 ± 5.61
4.41 ± 0.67
10.90 ± 5.22
4.18 ± 0.66
0.79 ± 0.44
0.22 ± 0.12
7.2 ± 5.3
5.3 ± 3.0
12,364 ± 1450 796 ± 73 300 ± 122 56
6,494 ± 312 840 ± 32 152 ± 28 88
a
GIT = gastrointestinal tract. The mass of a cylinder of GIT contents (200 mm in diameter for C. tentans and 52 mm in diameter for H. azteca), from which no alpha particle could reach the GIT wall, was subtracted from the total GIT mass. The resulting mass as a percentage of the total GIT mass is the mass within 50 mm from the wall. This percentage was used in internal AF calculations because only alpha particles within this portion of the GIT can irradiate body tissues to varying degrees, depending on their distance from the wall. b
was used as the target mass of body tissue exposed to radionuclides within the GIT. The AF values for alpha radiations were 0.14 for C. tentans and 0.22 for H. azteca. The short range of alpha particles within the GIT contents meant that only a portion of the GIT contents are within a 50-mm range from where an ionizing alpha particle can penetrate into body tissues. The proportion of GIT mass within 50 mm of the GIT wall was 56% in C. tentans and 88% in H. azteca, determined from calculated GIT diameters in Table 3. Within this 50-mm layer, the fraction of energy absorbed by body tissues ranges from 0.5 for particles emitted next to the GIT wall2 to 0 for alpha particle emitted from a distance of 50 mm from the wall (Fig. 1). The average fraction absorbed for all distances (0– 50 mm from the GIT wall) was estimated at 0.25.3 This value was then multiplied by the proportion of GIT contents < 50 mm from the GIT wall, yielding overall AF values of 0.14 (0.25 0.56) for C. tentans and 0.22 (0.25 0.88) for H. azteca. The doses from beta and gamma radiations within the GIT contents were ignored because the AF values from such radiations would be small. Many of the weak beta and 2
Only half of the alpha particles emitted next to the GIT wall will penetrate body tissue. The other half will decay within the GIT contents and, thus, do not deliver any dose to living biological tissue. 3 An exact absorbed fraction or AF value for all alpha radiations arising from within the 50-mm layer requires the integration of the energy deposited along all pathlengths for particles which can penetrate the GIT wall and the integration over all distances 0 – 50 mm from the GIT wall.
gamma radiations from the 238U series radionuclides would be attenuated in the GIT contents. Of the beta and gamma radiations reaching body tissues, little of the energy would be absorbed, given a total body width of < 900 mm in the study organisms. For alpha radiations from each radionuclide, a radiation weighting factor (wR) of 20 was applied to convert absorbed dose to ‘‘equivalent dose’’. The wR factor accounts for the greater RBE of alpha vs. beta and gamma radiations in causing biological damage. A factor of 20 is used for the greater RBE in causing cancer in humans (ICRP, 1991). It may or may not apply to other species and endpoints, but there is currently no better alternative. 2.2.2. Doses from radionuclides absorbed into body tissues from the GIT The second source of radiation dose to the benthic organisms is the radionuclides, which have been absorbed from the GIT and have entered body tissues. Ideally, these radionuclides would be directly measured in body tissues, but the small size of the organisms makes such measurements difficult. At equilibrium in a chronic intake situation, one can assume that the radionuclide concentration in the GIT will equilibrate with body tissues to varying degrees, depending on solubility of the radionuclide. As a first approximation of equilibrium for radionuclides between the GIT and body tissues, human f1 values were used in Eq. (2). Tissue dose ðmGy=yearÞ ¼ 1:6 1010 mJ=MeV Bq=kg d:w: 0:5 f1 s=year YE AF
ð2Þ
where f1 = 0.02 for U; 0.0005 for 234Th, 230Pa, and 230Th; 0.2 for 226Ra; 0.2 for 210Pb; 0.05 for 210Bi; and 0.5 for 210Po (ICRP, 1996). The f1 value estimates the fractional amount of each radionuclide absorbed from the GIT into the blood in adult humans (ICRP, 1996). The f1 values are no less accurate in this situation than animal/sediment concentration ratios (CRs), which range from 0.22 for field-collected organisms (Swanson, 1985) to < 0.136 (Hart et al., 1986) to 0.007 – 0.016 (Beak International, 1998) for uranium. Not only are CRs notoriously site-, species-, and season-specific, but they are unavailable for many of the 238U series radionuclides and require empirical measurements of radionuclide concentrations in very small organisms. Such measurements are likely to include the radionuclides in the GIT and adsorbed to the external surface of the organism, leading to an overestimate of the CR value. The f1 values may also overestimate absorption by invertebrates because the invertebrate GIT is more alkaline than the human GIT. Many radionuclides are more soluble in acid and, thus, may be absorbed more efficiently in humans than in invertebrates.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
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Fig. 1. Diagrammatic representation of the GIT within the body of an organism for AF determination. The length of each arrow represents the range of an alpha particle (i.e., 50 mm) originating within the GIT. (A) Alpha particles can be emitted at any angle from any point within the GIT. (B) AF = 0 because all of the alpha particle energy is deposited within the GIT before it reaches body tissues. (C) AF 0.25 for an alpha particle 25 mm from the GIT over the angle of 180. This approximation is used because most of the alpha energy is deposited near the end of its range. (D) AF = 0.5 for alpha particles at the GIT wall because half of all emitted particles will be completely absorbed by body tissue.
2.2.2.1. The special case of internal doses from radon progeny. One of the areas of uncertainty in 238U series dose calculations is the diffusion of radon within watery sediment and whether or not it can diffuse into and through small benthic organisms. For all practical purposes, there is no gaseous phase in the aquatic environment, only liquid. In a continuous liquid medium, the invertebrate body may pose no barrier to radon diffusion and, thus, radon and its progeny may be present in body tissues at the same concentrations as in external sediments. Alternatively, radon progeny may exist in body tissues only as a result of the decay of the 226Ra already absorbed into body tissues. Both assumptions were tested, assuming a variety of equilibrium states between 226Ra and radon.4 The reported doses from radon progeny in tissues were based on free diffusion of radon at 30% partial equilibrium with 226Ra in external sediment as was assumed for radon progeny in Table 2. Thus, for the short-lived radon progeny, Eq. (2) was used by employing an f 1 of unity and replacing Bq/kg d.w. sediment0.5 with Bq/kg w.w. sediment.
4 NCRP (1991) assumes 100% equilibrium of 226Ra with radon in external sediments but total loss of radon from 226Ra in invertebrate body tissues; Blaylock et al. (1993) assume 30% equilibrium of 226Ra with radon in external sediments.
Once within body tissues, 100% of the alpha energy from each nuclear decay will be absorbed by body tissues so that AF = 1. Beta and gamma radiations have greater penetration power and range than alpha particles and, thus, AF values for dose calculations depend on the overall size and shape of the organism and the kinetic energy of the radiation. For large organisms, AF values of unity for beta and gamma radiations of all energies have been used (Amiro, 1995; Amiro and Zach, 1993). As organisms become smaller, the AF decreases. The AF values for beta and gamma radiations were based on a point source dose distribution function developed by the National Council on Radiation Protection and Measurements (NCRP) (1991) and Blaylock et al. (1993) for uniformly distributed beta and gamma radiations in small aquatic insects and larvae, assumed to be 16-mg ellipsoid organisms (Fig. 2).5 The low-energy beta radiations from most of the decay series (U, 234Th, 230Th, 226Ra, 218Po, 214 Po, 210Pb, and 210Po) are completely absorbed within the organism, so that the internal AF = 1. Higher-energy beta radiations had lower internal AFs, which were 0.25 for
5
Average beta energies from Table 1 were multiplied by three and then estimated from graphs of AF vs. bmax in NCRP (1991) and Blaylock et al. (1993).
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Fig. 2. Geometry and dimensions (in mm) of benthic invertebrates. (a) Model aquatic organism for the point source dose distribution model (NCRP, 1991; Blaylock et al., 1993). (b) Measured and calculated dimensions for C. tentans and H. azteca in this study. 234
Pa, 0.38 for 214Bi, 0.6 for 210Bi, and 0.8 for 214Pb. Internal AFs for gamma radiations were 0.4 for all radionuclides in the decay series, except 214Pb and 214Bi (AF = 0.002 for both) and 210Bi, which does not produce gamma rays. The use of AF values from Blaylock et al. (1993) and NCRP (1991) for the 16-mg model organism may overestimate internal beta and gamma doses to the organisms used in this study by a factor of 2 – 4. The 16-mg organism was modeled as a solid cylinder. In contrast, the midge, C. tentans, and the amphipod, H. azteca, were considered
hollow cylinders around the GIT. Both species were smaller and four times thinner than the 16-mg model organism and C. tentans was twice as long. As shown in Fig. 2, the presence of the GIT contents would attenuate many of the weak beta and gamma radiation as they traverse body tissues. The decreased body thickness would further decrease internal AF values, relative to the model of NCRP (1991) and Blaylock et al. (1993). However, any overestimation in AF values here is accounted for by the principle of reciprocity in Eq. (3) below for external beta and gamma doses.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
2.2.3. Dose from radionuclides in the surrounding sediment External doses were calculated for beta and gamma, but not alpha, radiations in the external sediments. The short range of alpha particles prevents external alpha particles from penetrating the cuticle of most benthic organisms. Doses were calculated by Eq. (3), as follows: External dose ðmGy=yearÞ ¼ 1:61010 mJ=MeVðBq=kg w:w: s=yearÞ ðYE AFÞ
ð3Þ
where Bq/kg w.w. was the radionuclide concentration in wet weight sediment, determined from the moisture content data in Table 2, and YE values in million electron volts per disintegration were from Table 1. By the principle of reciprocity, NCRP (1991) and Blaylock et al. (1993) estimated external AF values to be 1 AF,
347
where AF = the internal AF value used in Eq. (2), for beta and gamma radiation within body tissues. For the lowenergy beta radiations from U, 234Th, 230Th, 226Ra, 218Po, 214 Po, 210Pb, and 210Po, the internal AF = 1 so the external AF = 0. The external AF values for higher-energy beta radiations were 0.75 for 234Pa, 0.62 for 214Bi, 0.4 for 210 Bi, and 0.2 for 214Pb. External AFs for gamma radiations were 0.6 for all radionuclides in the decay series, except 214 Pb and 214Bi (AF = 0.998 for both), with 210Bi not emitting gamma rays. Notably, external AFs increase as the energy of the radiation increases and the size of the organism decreases. External dose estimates, based on AFs from NCRP (1991) and Blaylock et al. (1993), were compared to other methods of external dose calculation (Table 4). An external AF of unity was employed on the assumption that small organisms, completely buried in sediment, receive the same average dose rate as the sediment itself, particularly for high-energy radiations. Dose rate constants from Beak
Table 4 External beta and gamma doses, calculated from dose coefficients from Beak Consultants (1985), AF values from Blaylock et al. (1993), and AF values of unitya Horseshoe Pond
Hidden Bay b
Tributary Creek
Blaylock
AF = 1
Beak Consultants
Blaylock
AF = 1
Beak Consultantsb
Blaylock
AF = 1
Beta doses (mGy/year) Natural U 27.2 234 Th 234 Pa 230 Th 226 Ra 222 Rn 218 Po 0.07 214 Pb 214 Bi 214 Po 210 Pb 0.15 210 Bi 210 Po Subtotal 27.43
– – 32.2 – – – – 0.01 0.08 – – 0.33 – 32.63
0.6 3.12 42.9 0 0 0 0 0.06 0.12 0 0.08 0.83 0 47.76
10.35
– – 12.2 – – – – 0 0.01 – – 0.12 – 12.37
0.23 1.19 16.3 0 0 0 0 0.01 0.02 0 0.03 0.29 0 18.08
0.26
– – 0.312 – – – – 0.003 0.023 – – 0.077 – 0.415
0.01 0.03 0.42 0 0 0 0 0.02 0.04 0 0.02 0.19 0 0.72
Gamma doses (mGy/year) Natural U 1.81 234 Th 234 Pa 230 Th 226 Ra 0 222 Rn 218 Po 0.19 214 Pb 214 Bi 214 Po 210 Pb 210 Bi 210 Po Subtotal 2.01
0.08 0.31 0.37 0 0 0 0 0.05 0.28 0 0.01 – 0 1.09
0.14 0.52 0.61 0 0 0 0 0.05 0.28 0 0.01 – 0 1.6
0.69
0.05 0.2 0.23 0 0 0 0 0.01 0.04 0 0 – 0 0.52
0.02
0.71
0.03 0.12 0.14 0 0 0 0 0.01 0.03 0 0 – 0 0.33
0.08
0.001 0.003 0.004 0 0.001 0 0 0.038 0.228 0 0.001 – 0 0.276
0.001 0.005 0.006 0 0.001 0 0 0.038 0.229 0 0.002 – 0 0.28
Total external dose
32.9
48.16
9.62
10.99
16.1
0.39
0.69
1
Beak Consultants
a
28.73
b
0.01
0.05
10.41
0 0.02
0.02
0.03
0.32
0 0.06
Dashes indicate external doses of zero. Doses listed as 0.00 or 0.000 are < 0.005 and < 0.0005, respectively. b The dose coefficients from Beak (1985) in units of mGy/h per Bq/kg wet weight sediment were as follows: 0.3 for beta and 0.02 for gamma from 238U through 234U; 0.0037 for gamma from 226Ra; 0.21 for beta and 0.6 for gamma from the short-lived radon progeny, and; 0.04 for beta from 210Pb.
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Consultants (1985) were also applied, which assumed a sediment density of 1.5 g/cm3.
3. Results The sediment data clearly showed that all radionuclides measured were elevated at Horseshoe Pond, the site closest to the outflow of uranium mining effluent (Table 2). The downstream site at Hidden Bay had lower concentrations
and the control site, Tributary Creek, the lowest of all. Notably, uranium and 210Pb concentrations were high in both Horseshoe Pond and Hidden Bay, whereas 230Th, 226 Ra, and 210Po concentrations were similar in Hidden Bay and Tributary Creek. Calculated doses from all individual radionuclides, according to the source and type of radiation, are shown in Appendix A for C. tentans and Appendix B for H. azteca. Absorbed dose estimates to invertebrates from all 238U series radionuclides were an order of magnitude higher for
Fig. 3. Absorbed (a) and equivalent (b) doses to larval midges (C. tentans) and adult amphipods (H. azteca) by radionuclide. HP = Horseshoe Pond, HB = Hidden Bay, and TC = Tributary Creek. TC enlarged for clarity.
Fig. 4. Absorbed (a) and equivalent (b) doses to larval midges (C. tentans) and adult amphipods (H. azteca) by radiation source and type. HP = Horseshoe Pond, HB = Hidden Bay, and TC = Tributary Creek. TC enlarged for clarity.
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
Horseshoe Creek and Hidden Bay than for the control site, Tributary Creek (Fig. 3a). Total doses from radionuclides within the three sources (GIT, body tissues, and external sediment) added up to 60, 19, and 3.2 mGy/year for both C. tentans and H. azteca from the Horseshoe Pond, Hidden Bay, and Tributary Creek sites, respectively. Most of the absorbed dose at the impacted sites was from external beta radiation, while it was internal alpha radiation at the control site (Fig. 4a). The most important contributors to absorbed doses for the two invertebrate species at Horseshoe Pond and Hidden Bay were 234Pa (54 – 65%), assumed to be in equilibrium with uranium, followed by uranium itself (25 – 31%). At Tributary Creek, most of the dose was internal alpha radiation in body tissues from 210Po (52%) and the short-lived radon progeny, 218 Po and 214Po (24%) (Fig. 3a). When a radiation weighting factor of 20 was employed for alpha radiation, equivalent doses from Horseshoe Pond, Hidden Bay, and Tributary Creek sediment were 540, 140, and 54 mGy/year, respectively, for C. tentans, and 560, 140, and 54 mGy/year, respectively, for H. azteca (Fig. 3b). For alpha radiations within the GIT, doses were 13 –14% higher in H. azteca vs. C. tentans because the GIT diameter was smaller, raising the AF. Alpha radiation from U in the GIT and body tissues accounted for 64– 70% of the total equivalent dose in both species using sediment from the Horseshoe Pond and Hidden Bay sites. In contrast, internal alpha radiation was 2% for U and 62% for 210Po for the Tributary Creek control site (Fig. 3b). External doses were calculated by three different methods (Table 4). The dose coefficients from Beak Consultants (1985) yielded the lowest dose estimates. The external doses based on an external AF of unity were 44 –46% higher than doses based on AFs from Blaylock et al. (1993). The greatest differences were for low-energy beta and gamma radiations, which are mostly absorbed from internal rather than external sources. There was almost no difference in the doses from the high-energy gamma radiations from 214Pb (0.249 MeV) and 214Bi (1.51 MeV). Thus, the use of the easier-to-calculate external AF of unity is justified for those gamma radiations, which are >0.25 MeV, or as a screening method for doses to extremely small organisms because their size decreases internal AFs and raises external AFs near unity. The dose estimates based on the AFs from Blaylock et al. (1993) were used in all other tables and figures. Overall, external beta radiation from 234Pa, followed by alpha radiation from U, contributed the majority of the total absorbed dose at the two impacted sites, whereas alpha radiation from 210Po, 218Po, and 214Po in body tissues was most important at the control site. Gamma radiation constituted 2– 3% at all three sites, being highest at the control site, primarily from the short-lived radon decay product, 214 Bi. If a radiation weighting factor of 20 was applied for alpha radiation, internal alpha radiation was estimated to contribute 91 – 99% of the total equivalent dose, either from
349
U at the two uranium-contaminated sites, or from 210Po and the alpha-emitting radon progeny at the control site.
4. Discussion This study was one of the first to consider the doses from alpha emitters within the GIT as a separate source of irradiation to aquatic organisms. Modeling the GIT as a separate source is equivalent to what Pentreath and Woodhead (2000) refer to as a Type B dosimetric model (i.e., a hollow cylinder geometry). The results show that the GIT is an important source of alpha irradiation for small aquatic organisms, adding a 10% absorbed dose increment and a 25% equivalent dose increment to benthic invertebrates in the vicinity of uranium mine operations. The results further show that the smaller organism (H. azteca, 4.4 mg) receives a greater dose from alpha radiations within the GIT than the larger organism (C. tentans, 12 mg). The difference in dose is more evident in equivalent dose calculations when a weighting factor is used for alpha radiation. As the diameter of the GIT decreases, the AF value for ingested alpha emitters increases because more alpha particles within the GIT reach body tissues. The AF value is 0.22 in H. azteca (4.4 mg), 0.14 in C. tentans (11.7 mg), and 0.005 in humans (70 kg, ICRP, 1979). If the GIT weight and diameter are known, AF values for internal alpha radiation doses can be calculated for other species and would be expected to play an increasingly important role in the dose to the smallest organisms. The invertebrate species considered here are ‘‘collectors – gatherers’’. They preferentially consume organic matter from sediment rather than whole sediment. Because organic matter has a greater capacity to adsorb radionuclides than whole sediment, radionuclide concentrations in the GIT may be even higher than the whole sediment concentrations, assumed to represent GIT contents in Eq. (1). The doses calculated in this study are highly dependent on assumptions of equilibrium made for a number of the 238 U decay series radionuclides. Thus, 234Pa delivered the majority of the absorbed dose to both invertebrate species because of its assumed equilibrium with uranium. Rapid establishment of equilibrium between uranium and 234Pa may also be a matter of some importance in terrestrial environments, potentially contaminated with depleted uranium, because of the increased external beta radiation exposure from 234Pa. The degree to which radon progeny are in equilibrium with 226Ra in sediment deserves further investigation, as does whether or not radon can diffuse into body tissues in aquatic environments. The NCRP (1991) assumes 100% equilibrium in sediment, but also claims that radon can escape to the atmosphere from lake sediments and that no radon would be present in invertebrate body tissues. In this study, only 30% equilibrium was assumed between radon
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and 226Ra in sediment, which may underestimate external dose from 214Bi. By assuming that radon and its progeny can equilibrate between sediment and invertebrate body tissue, the internal alpha dose from 218Po and 214Po may be overestimated. Appendices A and B show that alpha radiation from 218Po and 214Po in body tissues account for only 5 – 9% of the equivalent dose at the U contaminated sites but 29% at the Tributary Creek control site. Given the importance of these radionuclides in estimating human dose, it would seem unlikely that they should be ignored when estimating doses to benthic organisms. The doses calculated for the benthic invertebrates in Horseshoe Pond reached 540– 560 mGy/year, assuming a weighting factor of 20 for alpha. These doses were dominated by the alpha radiations in the GIT and body tissues from natural U concentrations of 50,484 Bq/kg in the sediment. These dose estimates were one to two orders of magnitude lower than NCRP (1991) dose estimates to insects near Tailings Creek in northern Saskatchewan where 238 U concentrations were only 150 Bq/kg. The differences in dose estimates were due to: (1) much higher 226Ra and 210 Po concentrations, measured or assumed in the Tailings Creek sediments; (2) the assumption of 100% equilibrium among 226Ra, radon, short-lived radon progeny, and 210Pb; and (3) the measurements and/or transfer parameters used to establish radionuclide concentrations in benthic invertebrate tissue. Equivalent doses of 540 –560 mGy/year can be compared to benchmark dose limits, which have been estimated to protect animal populations from detrimental effects, primarily such effects as early mortality, reduced reproductive success, and scorable cytogenetic damage (Pentreath, 1991; Pentreath and Woodhead, 2000). Dose limits from gamma radiation are estimated at 1 – 10 mGy/day or 360 –3600 mGy/year to ensure that mortality and reproductive effects will not occur in the most sensitive species, usually mice (International Atomic Energy Agency (IAEA), 1992; United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR), 1996). However, species differ in radiosensitivity and benthic macroinvertebrates are likely to be less radiosensitive than mice and other mammals (Rose, 1992). Doses of 10 mGy/day (3600 mGy/year) are generally assumed to protect aquatic organisms (NCRP, 1991; Pentreath, 1991; IAEA, 1992; UNSCEAR, 1996). For example, the number of egg capsules was reduced in the pond snail Physa at doses of 6.5 mGy/day. However, overall egg numbers only decreased at doses of 240 –1200 mGy/day. Birth rates in the cladoceran Daphnia did not drop until doses reached >4600 mGy/day (Marshall and Cooley in Blaylock et al., 1993). The potential radiation impacts of sediment-associated uranium on benthic organisms in Horseshoe Pond and Hidden Bay depends on the weighting factor used for alpha radiation and the endpoint considered. Trivedi and Gentner (2000) suggest that ecodosimetry weighting factors (eR) could be determined for three different endpoints for a
given organism: Level 1 for effects on individual organisms; Level 2 for effects on ecosystem sustainability; and Level 3 for genetic/hereditary effects. RBE values in the literature range from < 1 to 370 for alpha radiation, depending on the species and the endpoint considered. The RBEs are generally lower for genetic effects and in vitro studies, and higher for effects in vivo, such as the high RBE of 370 found for mouse oocyte survival (Samuels, 1966). Kocher and Trabalka (2000) suggest that a different weighting factor for animals will require the use of a unit for equivalent dose other than the Sievert. In this study, the equivalent doses to C. tentans and H. azteca in Horseshoe Pond (540 – 560 mGy/year) surpassed the lowest reproductive dose limit of 360 mGy/year, solely because a radiation weighting factor of 20 was used for alpha radiation. However, the 360-mGy/year limit is conservative, as it is generally applied to mammals rather than aquatic organisms. The long half-life of 4.5 billion years for 238U means that high mass concentrations of uranium must be ingested before it can deliver much of an alpha radiation dose. At high concentrations, however, the chemical toxicity of uranium may be more detrimental than the radiological dose. The uranium concentrations in sediment from Horseshoe Pond (2000 mg/kg or 50,484 Bq/kg), but not Hidden Bay (540 mg/kg or 13,616 Bq/kg), were capable of exerting toxic effects on H. azteca.6 Radiological effects are possible at the radiation doses calculated in this study, but are strongly dependent on the assumptions made regarding the weighting factor used for alpha radiation.
5. Conclusions A new method for calculating radiation doses to benthic invertebrates from measured sediment concentrations was described, which included dose contributions from sediment within the GIT. Doses from all radionuclides in the uranium-238 decay series were calculated to two representative benthic invertebrate species, based on radionuclide concentrations in sediment collected from sites near the Rabbit Lake uranium mining operation in northern Saskatchewan. Absorbed dose estimates were up to 20 times higher at the two uranium-contaminated sites compared to the control site. Most of the dose increment was from alpha radiation from uranium within the GIT, of which only 2% may be absorbed into body tissues, and from external beta radiation from 234Pa in the surrounding sediment, assumed to be at equilibrium with U in the sediment. The radiological doses estimated here, coupled
6 Spiked sediment toxicity tests with UO2(NO3)26H2O yielded 10-day LC50 values of 10,551 mg/kg d.w. for C. tentans and 2442 mg/kg d.w. for H. azteca; 10-day LC25 values were 4859 and 1449 mg/kg, respectively. For H. azteca, the 10-day EC50 for growth impairment was estimated at 964 mg/kg d.w. (Liber and White Sobey, 2000).
P. Thomas, K. Liber / Environment International 27 (2001) 341–353
with the additional chemotoxic stress of high uranium concentrations, may affect the survival, growth, or reproduction of these or similar species in the wild. The importance of the radiological effect depends on a radiation weighting factor of 20 for alpha radiation being biologically relevant to effects in invertebrate species. Additional work is needed to properly address these questions, particularly for the alpha-emitting radionuclides that prevail around uranium mines.
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Acknowledgments Funding for this research was provided by Saskatchewan Environment and Resource Management, Cameco, Cogema Resources, Cigar Lake Mining, the Canadian Nuclear Safety Commission, and the Department of Fisheries and Oceans. We would like to thank Dr. Sylvain St. Pierre, Mrs. Maisie Shiell, and two anonymous reviewers for their helpful comments on earlier versions of this manuscript.
Appendix A. Absorbed doses (Abs in mGy/year) and equivalent doses (Equiv in mGy/year20 for alpha) to C. tentans larvae from radionuclide concentrations in sediment Total Percent of Eq. (1) Eq. (2) Eq. (3) dose total dose Alpha in GIT Horseshoe Pond Natural U 5.9 234 Th 234 Pa 230 Th 1.7E 02 226 Ra 7.5E 02 222 Rn 218 Po 2.8E 02 214 Pb 214 Bi 214 Po 3.6E 02 210 Pb 210 Bi 210 Po 7.5E 02 Total 6.11 Hidden Bay Natural U 1.6E + 00 234 Th 234 Pa 230 Th 4.5E 03 226 Ra 6.6E 03 222 Rn 218 Po 2.5E 03 214 Pb 214 Bi 214 Po 3.2E 03 210 Pb 210 Bi 210 Po 2.0E 02 Total 1.62 Tributary Creek natural U 1.8E 02 234 Th 234 Pa 230 Th 3.3E 03 226 Ra 7.1E 03
Alpha in tissue
Beta in tissue
Gamma in tissue
12
0.03 3.8E 03 1.3E 02 2.6E 06 1.1E 03 2.0E 06 2.6E 06 4.3E 02 4.6E 02 1.5E 07 3.9E 02 6.1E 02 5.6E 08 0.24
0 2.4E 04 2.9E 04 1.1E 07 8.2E 04 3.0E 05 6.8E 07 9.2E 05 5.4E 04 6.2E 06 2.0E 03 0.0E + 00 2.3E 06 0.007
8.2E 03 1.0E 03 3.5E 03 6.8E 07 9.6E 05 2.5E 07 3.3E 07 5.4E 03 5.7E 03 1.9E 08 9.7E 03 1.5E 02 1.5E 08 0.05
7.2E 04 6.4E 05 7.8E 05 2.9E 08 7.2E 05 3.7E 06 8.5E 08 1.2E 05 6.8E 05 7.7E 07 4.9E 04 0.0E + 00 6.1E 07 0.002
9.4E 05 1.2E 05 4.0E 05 5.0E 07 1.0E 04
8.2E 06 7.4E 07 8.9E 07 2.1E 08 7.8E 05
8.4E 04 1.5E + 00 1.1E + 00
1.5E + 00
3.7E + 00 19.35
3.1E + 00
2.2E 04 1.3E 01 1.4E 01
1.8E 01
9.6E 01 4.54
3.6E 02
1.6E 04 1.4E 01
Beta external
32
1.1E 02 7.5E 02
Gamma external
Abs
Equiv
Abs
Equiv
2.9E 06 1.06
17.571 0.297 32.505 0.018 1.551 0.000 1.167 0.100 0.393 1.495 0.048 0.392 3.781 59.32
349.238 0.297 32.505 0.357 30.945 0.000 23.331 0.100 0.393 29.899 0.048 0.392 75.620 543.12
29.6 0.5 54.8 0.0 2.6 0.0 2.0 0.2 0.7 2.5 0.1 0.7 6.4 100
64.3 0.1 6.0 0.1 5.7 0.0 4.3 0.0 0.1 5.5 0.0 0.1 13.9 100
3.1E 02 1.1E 01 1.3E 01 5.0E 05 3.1E 04 5.5E 06 1.3E 07 5.7E 03 3.4E 02 1.2E 06 2.1E 03 0.0E + 00 1.0E 06 0.32
4.748 0.112 12.358 0.005 0.137 0.000 0.144 0.013 0.049 0.185 0.012 0.130 0.981 18.87
94.204 0.112 12.358 0.094 2.734 0.000 2.885 0.013 0.049 3.697 0.012 0.130 19.612 135.90
25.2 0.6 65.5 0.0 0.7 0.0 0.8 0.1 0.3 1.0 0.1 0.7 5.2 100
69.3 0.1 9.1 0.1 2.0 0.0 2.1 0.0 0.0 2.7 0.0 0.1 14.4 100
7.9E 04 2.8E 03 3.4E 03 8.1E 05 7.5E 04
0.055 0.003 0.314 0.004 0.148
1.081 0.003 0.314 0.069 2.940
1.7 0.1 9.8 0.1 4.6
2.0 0.0 0.6 0.1 5.5
0.082 2.9E 01 3.5E 01 1.3E 04 2.5E 03 4.4E 05 1.0E 06 4.6E 02 2.7E 01 9.3E 06 6.1E 03
3.3E 01 32.55
1.2E + 01
1.4E 03 9.3E 03
1.2E 01 12.35
3.1E 01
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222
Rn Po 214 Pb 214 Bi 214 Po 210 Pb 210 Bi 210 Po Total 218
2.7E 03
3.4E 01
3.4E 03
4.3E 01
3.3E 02 0.068
1.6E + 00 2.58
6.0E 07 7.9E 07 1.3E 02 1.4E 02 4.5E 08 2.9E 03 4.5E 03 2.5E 08 0.034
8.8E 06 2.0E 07 2.7E 05 1.6E 04 1.8E 06 1.5E 04 0.0E + 00 1.0E 06 4.4E 04
3.2E 03 2.2E 02
7.7E 02 0.41
1.3E 05 3.0E 07 1.4E 02 8.1E 02 2.8E 06 1.4E 03 0.0E + 00 4.0E 06 0.10
0.000 0.341 0.030 0.117 0.437 0.005 0.082 1.669 3.20
0.000 6.817 0.030 0.117 8.736 0.005 0.082 33.383 53.58
0.0 10.6 0.9 3.6 13.6 0.1 2.5 52.1 100
0.0 12.7 0.1 0.2 16.3 0.0 0.2 62.3 100
Appendix B. Absorbed doses (Abs in mGy/year) and equivalent doses (Equiv in mGy/year20 for alpha) to H. azteca adults from radionuclide concentrations in sediment Total Percent of Eq. (1) Eq. (2) Eq. (3) dose total dose Alpha in GIT Horseshoe Pond Natural U 6.7E + 00 234 Th 234 Pa 230 Th 1.9E 02 226 Ra 8.5E 02 222 Rn 218 Po 3.2E 02 214 Pb 214 Bi 214 Po 4.1E 02 210 Pb 210 Bi 210 Po 8.6E 02 Total 6.98 Hidden Bay Natural U 1.8E + 00 234 Th 234 Pa 230 Th 5.1E 03 226 Ra 7.5E 03 222 Rn 218 Po 2.8E 03 214 Pb 214 Bi 214 Po 3.6E 03 210 Pb 210 Bi 210 Po 2.2E 02 Total 1.85 Tributary Creek Natural U 2.1E 02 234 Th 234 Pa 230 Th 3.7E 03
Alpha in tissue
Beta in tissue
Gamma in tissue
1.2E + 01
3.0E 02 3.8E 03 1.3E 02 2.6E 06 1.1E 03 2.0E 06 2.6E 06 4.3E 02 4.6E 02 1.5E 07 3.9E 02 6.1E 02 5.6E 08 0.24
2.7E 03 2.4E 04 2.9E 04 1.1E 07 8.2E 04 3.0E 05 6.8E 07 9.2E 05 5.4E 04 6.2E 06 2.0E 03 0.0E + 00 2.3E 06 0.007
8.2E 03 1.0E 03 3.5E 03 6.8E 07 9.6E 05 2.5E 07 3.3E 07 5.4E 03 5.7E 03 1.9E 08 9.7E 03 1.5E 02 1.5E 08 0.05
7.2E 04 6.4E 05 7.8E 05 2.9E 08 7.2E 05 3.7E 06 8.5E 08 1.2E 05 6.8E 05 7.7E 07 4.9E 04 0.0E + 00 6.1E 07 0.002
9.4E 05 1.2E 05 4.0E 05 5.0E 07
8.2E 06 7.4E 07 8.9E 07 2.1E 08
8.4E 04 1.5E + 00 1.1E + 00
1.5E + 00
3.7E + 00 19.35
3.1E + 00
2.2E 04 1.3E 01 1.4E 01
1.8E 01
9.6E 01 4.54
3.6E 02
1.6E 04
Beta external
3.2E + 01
1.1E 02 7.5E 02
Gamma external
Abs
Equiv
Abs
Equiv
2.9E 06 1.06
18.401 0.297 32.505 0.020 1.562 0.000 1.171 0.100 0.393 1.500 0.048 0.392 3.792 60.18
365.846 0.297 32.505 0.405 31.156 0.000 23.411 0.100 0.393 30.001 0.048 0.392 75.832 560.39
30.6 0.5 54.0 0.0 2.6 0.0 1.9 0.2 0.7 2.5 0.1 0.7 6.3 100
65.3 0.1 5.8 0.1 5.6 0.0 4.2 0.0 0.1 5.4 0.0 0.1 13.5 100
3.1E 02 1.1E 01 1.3E 01 5.0E 05 3.1E 04 5.5E 06 1.3E 07 5.7E 03 3.4E 02 1.2E 06 2.1E 03 0.0E + 00 1.0E 06 0.32
4.972 0.112 12.358 0.005 0.138 0.000 0.145 0.013 0.049 0.185 0.012 0.130 0.983 19.10
98.684 0.112 12.358 0.107 2.752 0.000 2.892 0.013 0.049 3.706 0.012 0.130 19.668 140.48
26.0 0.6 64.7 0.0 0.7 0.0 0.8 0.1 0.3 1.0 0.1 0.7 5.1 100
70.2 0.1 8.8 0.1 2.0 0.0 2.1 0.0 0.0 2.6 0.0 0.1 14.0 100
7.9E 04 2.8E 03 3.4E 03 8.1E 05
0.057 0.003 0.314 0.004
1.133 0.003 0.314 0.078
1.8 0.1 9.8 0.1
2.1 0.0 0.6 0.1
8.2E 02 2.9E 01 3.5E 01 1.3E 04 2.5E 03 4.4E 05 1.0E 06 4.6E 02 2.7E 01 9.3E 06 6.1E 03
3.3E 01 32.55
1.2E + 01
1.4E 03 9.3E 03
1.2E 01 12.35
3.1E 01
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Ra Rn 218 Po 214 Pb 214 Bi 214 Po 210 Pb 210 Bi 210 Po Total
8.1E 03
1.4E 01
3.1E 03
3.4E 01
222
3.9E 03
4.3E 01
3.8E 02 0.077
1.6E + 00 2.58
1.0E 04 6.0E 07 7.9E 07 1.3E 02 1.4E 02 4.5E 08 2.9E 03 4.5E 03 2.5E 08 0.034
7.8E 05 8.8E 06 2.0E 07 2.7E 05 1.6E 04 1.8E 06 1.5E 04 0.0E + 00 1.0E 06 4.4E 04
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