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Radium isotopes as indicators of adsorption–desorption interactions and barite formation in groundwater
Journal of Environmental Radioactivity 46 (1999) 271}286
Radium isotopes as indicators of adsorption}desorption interactions and barite formation in groundwater Paul Martin *, Riaz A. Akber Environmental Research Institute of the Supervising Scientist (ERISS), Jabiru, NT 0886, Australia Queensland University of Technology, GPO Box 2434, Brisbane, Queensland 4001, Australia Received 14 May 1998; received in revised form 16 September 1998; accepted 20 November 1998
Abstract Borewater monitoring near the Ranger U mine tailings dam has revealed deterioration in water quality in several bores since 1981, with increases in sulphate concentrations of up to 1000 times being observed. In some cases, signi"cant increases in Ra concentrations also occurred. In this study, isotopes of Ra, Th and Ac were measured in borewater in 1988}1993. For the most seepage-a!ected bores, the concentrations of Ra, Ra and Ra as well as the nuclidic ratios Ra/Ra, Ra/Ra and Ra/Ra generally increased. Time-series concentration increases were observed for Sr but not for Ba. Ac concentrations increased with time, but not su$ciently to account for the increasing Ra concentrations. It is concluded that increases observed in Ra isotope concentrations arise from competition for cation adsorption sites in the vicinity of the bore rather than by direct transport of Ra from the tailings. Formation of a barite solid phase is occurring in the groundwater and causing the removal of some Ra from solution, with rapid replenishment of the shorter-lived isotopes from their parents. 1999 Elsevier Science Ltd. All rights reserved. Keywords: Radium; Actinum; Thorium; Barite; Groundwater; Uranium mining
1. Introduction Signi"cant changes in groundwater chemistry and, in some cases, Ra concentrations were observed in the vicinity of the Ranger U mine tailings dam since the beginning of the mining operation. In particular, Ra concentrations increased by factors of about 2}4 between 1982 and 1988 for three bores (OB11A, OB13A and OB16; Fig. 1). For those bores a!ected by seepage, concentrations of Ca, Mg *Corresponding author. Tel.: 0061 889 799 711; fax: 0061 889 792 076; e-mail: [email protected] 0265-931X/99/$ - see front matter 1999 Elsevier Science Ltd. All rights reserved. PII: S 02 6 5-9 3 1X (9 8) 0 0 14 7 - 7
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Fig. 1. Ranger bore location plan.
and SO\ increased with time. In this paper, we report the results of a study of the behaviour of Ra isotopes as a function of the concentrations of these major ions. The four natural-series Ra isotopes are: Ra (t "1600 yr) of the U series, Ra (11.4 d) of the Ac series, and Ra (3.66 d) and Ra (5.75 yr) of the Th series. The natural activity ratio U/U is 0.046, and so the activity concentrations of nuclides of the Ac series are normally very low in environmental samples. The Th/U activity ratio is variable in nature due to the quite di!erent geochemical behaviour of Th and U. The di!erences in their half-lives makes the Ra isotopes potentially useful geochemical tools, and they have been used in studies of Ra migration in aquifers (Dean, Bland & Levinson, 1983; Davidson & Dickson, 1986; Dickson, Giblin & Snelling, 1987), determination of in situ retardation factors and distribution coe$cients (Laul, 1992), adsorption}desorption reaction rates
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273
(Krishnaswami, Graustein, Turekian & Dowd, 1982; Krishnaswami, Bhusan & Bhaskaran, 1991) and for geochemical exploration (Dickson, Meakins & Bland, 1983). A number of studies of Ra isotopes in groundwater have found cases where the Ra/Ra and Ra/Ra ratios were greater than the theoretical parent decay supply values of 0.046 and 1.0, respectively. Dickson (1990) reported that for a sample set of over 400 Australian groundwaters the median value for Ra/Ra was 0.075. Kuptsov et al. (1969) measured Ra/Ra ratios between 3 and 105 in carbonate spring waters. Such ratios are often associated with elevated sulphate concentrations (for example, Dickson, 1985; Krishnaswami et al., 1991; Martin & Murray, 1991; Sturchio, Bohlke & Markun, 1993), and were also observed for some seepage-a!ected bores in the present study. In this paper, possible causes of the high ratios measured in the Ranger borewaters are discussed.
2. Study area The ERA-Ranger U mine and mill complex is located approximately 260 km east of Darwin in the Alligator Rivers Region of northern Australia. Excavation of orebody No. 1 commenced in August 1980 and ended in October 1994. Excavation of orebody No. 3 began in 1996. Commercial operation of the mill commenced in August 1981. Current annual production is about 4000 tonnes U O . The climate is tropical monsoonal, with contrasting wet (approximately December}April) and dry (approximately May}November) seasons. Mean annual rainfall is about 1480 mm and the mean annual Class A pan evaporation is about 2590 mm. Release of contaminated water into the o!-site environment is minimised by use of a number of waterbodies with varying water quality, including four retention ponds (RP1}RP4), the minepits and a tailings dam (Fig. 1). Covering approximately 110 ha of land, the tailings dam is the largest of the water management structures. Subaqueous tailings deposition was practiced until 1987, but since then tailings have been kept largely sub-aerial in order to achieve a greater density. At the Ranger mill, U is leached from the ore by H SO using pyrolusite as an oxidant (Noller, 1991). Sulphate is also produced due to the oxidation of sulphidebearing minerals present in the catchment area of the retention ponds. Consequently, on-site waterbodies are generally characterised by high sulphate levels. Sulphate concentrations in tailings dam surface water were approximately 70 and 300 mM in 1983 and 1993, respectively. Corresponding Mn concentrations were approximately 2 and 20 mM, respectively. Tailings are neutralised before placing in the tailings dam. Two aquifers exist in the Ranger on-site area (Whitehead, 1980; Ahmad & Green, 1986), an uncon"ned shallow aquifer varying in thickness from 2 to 20 m and associated with the soil pro"le and underlying laterite layer, and a con"ned to semi-con"ned deep aquifer in underlying weathered and fresh bedrock. A layer of low permeability sandy silty clay is commonly found between the laterite and the weathered bedrock. Where it occurs, this layer forms a semi-con"ning upper boundary to the deep aquifer. Both aquifers are a!ected by the seasonal nature of the rainfall, with water levels in deep aquifer bores falling by about 1}2 m between May and November.
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Table 1 Bore slotted depth, water temperature and Fe, Mn and Pb borewater concentrations Bore
Slotted depth (m)
T (3C)
OB1A OB2A OB4A OB6A OB9A OB10A OB11A OB12A OB13A OB15 OB16 OB17A OB19A OB21A OB22 OB23 OB24 OB29 OB44
16}31 15}30 22}37 14}26 12}30 12}30 11}23 16}31 10}30 10}25 10}20 23}41 * 31}43 25}41 36}51 22}36 35}50 11}16
32 * 34 33 34 31 33 33 33 28 35 * 34 31 * * * 31 28
Fe (lM)
2 :1 3 4 2 :1 3 13 :1 :1 4 * 2 5 2 * :1
Mn (lM)
Pb (nM)
21 0.18
:0.7 * 1.5 * 0.8 5.2 6.8 :1 * 6.3 * 1.1 1.5 * * * * * *
0.27 0.88 0.36 0.20
0.74 0.55 5 3.7 0.18 3.3 0.09 0.13
Time-series changes have been observed for these bores; see text.
A number of observation bores have been drilled in the area for groundwater quality monitoring purposes. In this paper we deal only with bores accessing the deeper aquifer (Fig. 1). All the bores discussed have a PVC casing; where known, bore slotted depths are given in Table 1.
3. Sampling and analytical methods Sampling was timed to coincide with the routine groundwater sampling program of Ranger, most collections occurring in May or November. Containers and lids were washed with 2% HNO and rinsed with demineralised water before use, and rinsed three times with bore water immediately before collecting the sample. Samples were collected approximately 20 min after the start of pumping. The pump rate was approximately 5 l min\. Samples were "ltered through a 0.45 lm "lter before acidi"cation to 1% HNO before analysis. Only one sample was collected for analysis on each sampling occasion. Uncertainties quoted for radionuclide analyses are one standard deviation due to counting statistics only. Since the main focus of the work was Ra behaviour, all samples were analysed for Ra and Ra, and most for the shorter-lived Ra and Ra. Some samples were also analysed for Th and Ac isotopes, primarily in order to investigate whether they were supporting their Ra progeny in the water.
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275
Most of the radionuclide analyses were performed by a-ray spectrometry (Hancock & Martin, 1991; Martin, Hancock, Paulka & Akber, 1995). Samples were collected into a 5 l polypropylene container. Chemical separations for radium analysis were generally performed within 2 days after sample collection. Results for Ra, Ra and Ra were corrected to the date and time of sample collection assuming nil support from the Th parent. Ra/Ra ratios were determined by recounting Ra sources after waiting at least 12 months for ingrowth of Th and Ra. The Ra/Ra ratio in tailings dam water was determined by washing the deposit from the source several years after preparation and determining Th on the washings. Ac (t "21.8 yr) was determined by chemical separation of Ac, with subsequent detection of the a-decays of the Th and Ra progeny; Ac was used as a yield tracer. Th (t "1.91 yr) determinations were performed 8 days or less after sample collection, and corrected for ingrowth from Ra. Th (t "18.7 d) was determined by two separate determinations of Ra on the sample, one performed soon after sample collection and the second after a delay of approximately 2 weeks. For some samples Ra and Ra were determined by c-ray spectrometry. In this case, a 20 l sample was collected into a PVC pail and Ra co-precipitated with MnO , followed by a second co-precipitation with Fe(OH) . The precipitate was "ltered and ashed, then cast in a mould with a polyester resin before counting on a Ge c-ray spectrometer (Murray, Marten, Johnston & Martin, 1987). Analyses for Fe, Mn, Ba and Sr were performed by ICP-AES; those for Pb were performed by graphite furnace atomic absorption spectrometry. Borewater temperatures were measured in situ after pumping, using a K-type thermocouple.
4. Results and discussion 4.1. General chemical parameters Groundwaters of the aquifer are Mg>}Na>}HCO\ type, with low concentrations of SO\ and Cl\. Conductivities are generally in the range of 100}400 lS cm\, with pH values of 6.2}7.5. SO\ concentrations in a group of bores to the north of the tailings dam (OB11A, 13A, 15, 16 and, to a lesser extent, 44) increased since 1982 (Fig. 2). Anion equivalences are now dominated by SO\ for these bores. Concentrations of Ca>, Mg>, Na> and K> have all increased. The cation equivalent dominance of Mg> also increased. The pH of water from these bores decreased by 0.5}1.0 pH units over this period, although most of this decrease occurred prior to 1988. Increasing SO\ concentra tions were observed in OB6A (from 1983), OB9A and 10A (from 1986), OB4A (from 1989) and OB2A (from 1991). However, these increases were not as pronounced as those for the bores to the north of the dam. Ranger's monitoring data show that between 1981 to 1993, Mn concentrations in bores OB4A, 11A, 13A, 15 and 29 increased from about 2}15, 1}5, 0.5}7, 1}10 lM and 0.04}1.5 lM, respectively (ERA Ltd, unpublished data). For the remaining bores, Mn concentrations were relatively constant. Table 1 shows measurements of temperature
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Fig. 2. Time-series concentrations of SO\.
and Fe, Mn and Pb concentrations made as part of the present study. Fe concentrations were high and increasing with time in bores OB1A and 4A. Individual Fe measurements were 210 and 470 lM for OB1A in May 1991 and November 1992, respectively, and 140, 310 and 430 lM for OB4A in November 1990, November 1992 and May 1993, respectively. 4.2. Ra isotope results Fig. 3a shows the activity concentrations of Ra for the "ve bores most a!ected by seepage (OB11A, 13A, 15, 16 and 44) between 1988 and 1993. A general trend of increasing Ra concentrations with time is apparent over this period. As discussed further below, there is a possible seasonal variability in Ba concentrations in these borewaters, and so the signi"cance of any apparent trend in Ra concentrations is best established using same-month comparisons. Table 2 lists the available data for Ra isotope activity concentrations for the Novembers of years 1988}1992. For Ra there was a signi"cant (2p) increase for OB11A, 13A, 15 and 16 over this period. Signi"cant increases were also observed for Ra and Ra for bores OB11A, 13A and 15. These results are explicable in terms of increasing groundwater salinity. It is well established that increasing cation concentrations will lead to competition with Ra for adsorption sites and consequently higher groundwater Ra concentrations (Kraemer & Reid, 1984; Langmuir & Melchior, 1985; Ku, Luo, Leslie & Hammond, 1992). The reduction in pH values observed for these bores could also have contributed to this desorption e!ect. Fig. 3b shows that although Ra activity concentrations increased signi"cantly in OB11A over the 1988}1992 period, they #uctuated in OB13A and 15 and decreased in OB16. These results, especially the pronounced Ra decrease in OB16 following a concentration peak in 1989, as well as the decreases observed for Ra and Ra in
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277
Fig. 3. Time-series concentrations of (a) Ra, (b) Ra and (c) Ac.
OB16 (Table 2), do not appear to be consistent with the hypothesis of desorption due to major cation concentration increases. Ranger's monitoring data for tailings dam surface water show Ra concentrations in the range 5.3}33 Bq l\ (n"37) between 1988 and 1993. Measurements of the Ra/Ra ratio in "ltered tailings dam water samples collected in July 1988 and December 1991 were 0.0010$0.0003 and 0.002$0.001, respectively. Consequently, if a signi"cant migration of Ra from the tailings dam were occurring, a decrease in Ra/Ra ratios in borewater would be expected to occur over time. Table 3 shows the percentage change (*Ra) in a number of Ra isotope ratios between November 1988 and the Novembers of subsequent years, calculated as
*Ra"
(R !R ) 66 ;100% R
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Table 2 Ra, Ra, Ra and Ra activity concentrations for Novembers of the given year Bore
Year
Ra (mBq l\)
Ra (mBq l\)
Ra (mBq l\)
Ra (mBq l\)
OB11A
1988 1989 1990 1991 1992 1988 1989 1992 1988 1989 1990 1992 1988 1989 1991 1992 1989
134$3 143$3 192$17 177$3 169$5 66$1 83$3 56$2 62$4 54$2 76$5 48$2 80$2 105$3 37$3 30$1 49$2
51$1 75$3 86$8 94$5 100$4 92$3 96$5 94$4 142$14 159$10 244$17 196$9 129$4 106$7 79$10 66$3 23$4
57$3 86$5 105$10 * 124$7 125$5 198$10 247$16 129$10 150$9 245$19 348$20 247$11 * * 207$17 28$3
8.7$0.7 13.4$1.3 18.2$1.9 * 24.7$2.3 11.2$1.0 15.9$1.1 26.5$2.7 6.8$0.9 8.2$1.1 12.0$1.6 16.2$1.6 21.0$1.6 * * 29.9$3.6 5.6$0.9
OB13A
OB15
OB16
OB44
Table 3 Percentage changes in Ra/Ra, Ra/Ra, Ra/Ra, Ra/Ra and Ra/Ra activity ratios between November 1988 and November of the given year Bore
Year
*Ra86 (%)
*Ra46 (%)
*Ra36 (%)
*Ra48 (%)
*Ra43 (%)
OB11A
1989 1990 1991 1992 1989 1992 1989 1990 1992 1989 1991 1992
38$6 17$3 39$7 55$5 917$4 20$5 29$11 41$10 78$13 937$4 31$13 36$5
40$10 28$6 * 72$11 26$6 130$20 33$8 55$9 250$20 * * 120$20
45$18 46$14 * 125$30 13$12 180$40 40$20 40$20 210$40 * * 280$50
2$9 10$6 * 12$8 53$10 96$15 3$11 10$10 96$17 * * 64$15
*9$13 *17$8 * 929$9 11$12 917$11 93$16 7$16 13$16 * * 944$8
OB13A OB15
OB16
where R " the activity ratio in November 1988 and R " the activity ratio in 66 November 19XX. Over the course of the study the Ra/Ra ratio actually increased signi"cantly in all four bores. The e!ect of direct migration of Ra from the dam must, therefore, be small in magnitude in comparison with other e!ects at this stage.
279
P. Martin, R.A. Akber/J. Environ. Radioactivity 46 (1999) 271}286 Table 4 Th, Th and Th activity concentrations, November 1989 Bore
Th (mBq l\)
Th (mBq l\)
Th (mBq l\)
OB9A OB11A OB13A OB15 OB44
:0.2 :0.2 :0.3 :0.1 :0.3
:0.1 :0.1 :0.1 :0.1 :0.2
:0.4 :0.4 :0.4 :0.9 :0.7
*Ra values for Ra/Ra and Ra/Ra ratios were generally higher than those for the Ra/Ra ratio. Given their relatively short half-lives, Ra and Ra would not generally be expected to migrate long distances in an aquifer (Osmond & Cowart, 1992). If this assumption is correct, then the Ra/Ra ratio would not be expected to vary signi"cantly with time. *Ra values (Table 3) show that bore OB16 has experienced a decreasing Ra/Ra ratio, almost having its value between 1988 and 1992. This ratio also decreased signi"cantly in OB11A over this period. These results will be discussed further below. 4.3. Ac, Th and Ra isotope concentrations Table 3 shows that for the most seepage-a!ected bores the Ra/Ra and Ra/Ra ratios have increased with time. These ratios are now signi"cantly above the theoretical parent decay supply values of 0.046 and 1.0, respectively, in OB11A, 13A, 15, 16 and 44. Such increases could arise from increased support for Ra and Ra from their parents in the water column. In a study of Ra isotopes in borewaters and saline seepages in Western Australia, Dickson (1985) reported Ra/Ra ratios in the range of 0.14}0.62. High Ra/Ra ratios were generally associated with high Cl\ and SO\ concentra tions. Dickson postulated that soluble single and double SO\ salts of Ac may be formed, resulting in support for Ra in the water. His measurements of the Western Australian waters indicated that there was indeed some support for Ra in the water, with most Ac/Ra ratios in the range 0.25}0.8. Ac activity concentrations in Ranger bores OB11A, 13A, 15, 16 and 44 have increased with time (Fig. 3c), showing that some Ac mobilization is occurring. However, in none of the borewaters tested did the Ac activity concentration exceed 20% of that of Ra. On most occasions Ac support in the water could account for no more than 4% of the Ra excess in OB13A, 15 and 16, and about 10}20% of the excess in OB11A and 44. In addition, Ac does not decay directly to Ra, but passes through the intermediate Th isotope Th. In the pH range prevailing in these bores, most of this Th can be expected to be lost from the water column due to adsorption processes (Lieser & Hill, 1992), reducing further the degree of support for Ra in the water. Measurements of the members of the Ac, U and Th decay series (Tables 4 and 5) con"rm that Ra isotopes are in substantial excess over their Th parents in the water.
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Table 5 Th and Ra concentrations, September 1988 Bore
Th (mBq l\)
Ra (mBq l\)
OB11A OB13A OB16
:3 :4 :4
9.3$0.9 12.0$0.8 21$1
The di!erence in degree of Ac mobilization between the Ranger borewaters and the Western Australian waters described by Dickson is probably due to the greater acidity (pH 4}5) of the latter. Lieser et al. (1991) reported the results of Ac adsorption experiments carried out for a groundwater from Gorleben (Germany). They found that salinity had an appreciable e!ect on Ac adsorption at pH 3.5, but not at pH 7.5. Their conclusion was that at pH 7.5 hydrolysis of Ac is practically complete, whereas at pH 3.5 hydrolysis is incomplete and Na> ions can compete with the ionic species of Ac for ion exchange sites. 4.4. Evidence for formation of a barite phase Langmuir and Riese (1985) and Langmuir and Melchior (1985) studied the trace aqueous and solid solution chemistry of Ra, and concluded that, since Ra concentrations in natural waters and waters associated with U mining are probably never high enough to reach saturation with respect to a pure Ra solid, concentrations are limited by adsorption or solid solution formation. On the basis of the results presented above, we conclude that a combination of both of these processes is a!ecting Ra isotope concentrations in the Ranger borewaters. Formation of a solid solution would result in the removal of Ra from the water. Subsequent ingrowth due to the decay of the parents in the source rocks will be more rapid for short-lived Ra isotopes (hereafter designated 1Ra) than for long-lived ones (*Ra). Consequently, the 1Ra/*Ra activity ratio in the water should increase for a period dependent on the isotope half-lives. Analytical data were used in conjunction with the program WATEQP (Appelo and Postma, 1994) to calculate the saturation states of SO\ mineral phases in these waters. The saturation state (X) is given by X"IAP/K where IAP is the ion activity product and K is the solubility product of the mineral at the speci"ed temperature and pressure. Ion activity products are calculated from the activities of the uncomplexed ions in solution, for example, IAP "[Ba>][SO\], where the quantities in square brackets represent activities of the uncomplexed ions in a water sample. Waters for which ) is greater than 1 are termed supersaturated, while those for which ) is less than 1 are subsaturated. Ranger borewaters were found to be
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281
Fig. 4. Ra/Ra vs. the barite saturation state ()).
Fig. 5. Time-series concentrations of (a) Sr and (b) Ba.
subsaturated with respect to anglesite (PbSO ) and celestite (SrSO ). However, for a number of bores the waters were supersaturated with respect to barite (BaSO ). Fig. 4 shows the activity ratio Ra/Ra plotted against the calculated barite saturation state in borewater samples for which su$cient data are available. The Ra/Ra ratio is generally signi"cantly greater than 0.046 for those bores with ) greater than about 1}2, in agreement with the prediction above of increasing 1Ra/*Ra ratios following formation of a solid solution.
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Sr concentrations measured for bores OB4A, 9A, 11A, 13A, 15, 16 and 44 increased with time (Fig. 5a). This behaviour is to be expected as Sr will be removed from adsorption sites by increased salinity in a similar mechanism to that a!ecting Ra. In contrast, Ba concentrations for OB11A, 13A, 15, 16 and 44 have tended to #uctuate, generally being lower in November (i.e. late dry season/early wet season) than in May (late wet season/early dry season), and showed no apparent increase with time since 1988 (Fig. 5b). In light of the saturation state calculations discussed above, this is good evidence that a barite phase is being formed in these groundwaters. Although the most recent samples from OB4A and 6A showed calculated values for ) greater than 1, an increase in the Ra/Ra ratio was not observed (Fig. 4). We interpret this as indicating that barite formation did not occur in the vicinity of these bores during the study. The fact that Ba concentrations increased in water from bore OB4A and were relatively steady in OB6A (Fig. 5b) supports this conclusion. 4.5. Implications for modelling of Ra behaviour Results presented above demonstrate that the primary source of Ra in borewater from the Ranger site is the aquifer rocks in the vicinity of the bore, rather than transport from the tailings dam. Further, for some bores increasing SO\ concentra tions led to the presence of a barite phase with co-precipitation of Ra. It follows that a model based upon retardation coe$cients cannot predict the behaviour of radium in this system. The potential in#uences on Ra isotope activity concentrations are: (1) Ingrowth via decay of the ¹h parent in the water: Th isotope measurements showed that there is little direct support for Ra from dissolved parents in the Ranger borewaters. (2) Release of Ra following dissolution of aquifer minerals: Increases in Mn and Fe concentrations observed in a few bores (Table 1) may be due to dissolution of native oxides, although direct transport of Mn from the tailings dam cannot be ruled out. Dissolution of Mn or Fe oxides could be expected to increase Ra isotope concentrations in the water, but be accompanied either by no change in or by reduction in 1Ra/*Ra ratios. Consequently, if it is occurring, then the e!ect on the ratios must be small compared with other factors in the bores to the north of the tailings dam. (3) Ingrowth from the parent in the solid by a-recoil processes: The a-recoil is a physical process and is unlikely to be a!ected substantially by a change in water chemistry. For this reason, we will assume here that supply rates of di!erent Ra isotopes by this process remained constant. (4) Adsorption onto and desorption from solid surface sites: Increasing groundwater salinity should result in increased Ra activity concentrations in the water. The e!ect on Ra isotope ratios would depend upon the adsorption and desorption rate constants, as discussed below. (5) Removal of Ra from solution by a precipitating mineral phase: This process would result in a decrease in activity concentrations of all Ra isotopes in solution in the short term. Subsequent ingrowth of short-lived isotopes would result in an increase in 1Ra/*Ra ratios.
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283
Table 6 Mean (standard deviation) for Ra activity concentration and Ra/Ra, Ra/Ra and Ra/Ra activity ratios for a number of bores. For n "1 the uncertainty is one standard deviation due to counting statistics only Bore
Ra (mBq l\)
Ra/Ra
n
Ra/Ra
Ra/Ra
Ra/Ra
n
OB1A OB2A OB4A OB6A OB7A OB9A OB10A OB12A OB17A OB18A OB19A OB20 OB21A OB22 OB23 OB24 OB28 OB29 OB30 OB41 OB43
36 (2) 25 (6) 96 (19) 330 (48) 104 (6) 68 (20) 73 (14) 34 (1) 84 (34) 70 (6) 120 (120) 41$2 200$40 160 (50) 170 (90) 24$1 97$3 61 (33) 20$2 34$3 31$2
0.7 (0.3) 1.0 (0.5) 0.53 (0.04) 0.18 (0.02) 0.13 (0.04) 0.27 (0.03) 0.10 (0.03) 2.1 (0.1) 1.3 (0.7) 1.5 (0.2) 0.58 (0.08) 0.68$0.06 0.26$0.03 0.06 (0.02) 0.54 (0.05) 0.29$0.03 0.24$0.04 0.62 (0.35) 0.52$0.11 0.56$0.11 0.50$0.12
2 2 7 11 2 7 6 2 5 3 2 1 1 3 2 1 1 3 1 1 1
0.033$0.008 0.029$0.004 0.037 (0.004) 0.043 (0.006) 0.026$0.004 0.046 (0.016) 0.032 (0.006) 0.105$0.008 0.029 (0.014) 0.047 (0.004) 0.044$0.003 9 * 0.059 (0.011) 0.021$0.007 0.028$0.019 * 0.062 (0.013) * * *
0.66$0.12 0.77$0.05 0.62 (0.12) 1.06 (0.23) 1.21$0.15 0.84 (0.20) 1.03 (0.31) 1.04$0.05 0.82 (0.21) 0.65 (0.06) 1.03$0.05 * * 0.61 (0.23) 0.40$0.07 1.20$0.39 * 1.13 (0.16) * * *
0.32$0.06 1.06$0.04 0.33 (0.06) 0.19 (0.04) 0.128$0.015 0.23 (0.05) 0.09 (0.02) 2.23$0.06 0.82 (0.36) 1.00 (0.13) 0.538$0.013 * * 0.036 (0.006) 0.23$0.04 0.35$0.10 * 0.69 (0.46) * * *
1 1 6 9 1 6 4 1 4 3 1 * * 3 1 1 * 2 * * *
89 (75)
0.60 (0.51)
0.043 (0.021)
0.87 (0.25)
0.55 (0.57)
Mean (s.d.)
n"number of samples for Ra and Ra Ra. n"number of samples for Ra/Ra, Ra/ Ra and Ra/Ra.
Following the model and notation of Krishnaswami et al. (1982), we can represent the adsorption/desorption process by "rst-order adsorption (k ) and desorption (k ) rate constants of dimension time\. The values of k and k will depend on the system under consideration, and in particular on the major ion chemistry of the water phase. In general, k will be several orders of magnitude smaller than k and most of the Ra potentially available for dissolution in the groundwater will reside on adsorption sites, although this may not be the case for highly saline waters (Laul, 1992). Where k is of the order of or less than the decay constant (j) of a short-lived Ra isotope, the adsorption process will result in the 1Ra/*Ra ratio in the water phase being higher than that available from supply by parent decay. Table 6 shows the means of all readings for Ra/Ra and Ra/Ra in Ranger bores except OB11A, 13A, 15, 16, and 44. These values should then represent the situation in Ranger bores when the major processes occurring are supply by parent decay and adsorption/desorption reactions. The mean values for Ra/Ra and Ra/Ra over these bores are 0.043 and 0.87, respectively. Ninety-"ve percent con"dence
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intervals for these means (0.043$0.012 and 0.87$0.14) included the theoretical parent decay supply values of 0.046 and 1.0, respectively. For a few bores (OB1A, 2A, 4A and 23), both of these ratios were substantially lower than the theoretical parent decay supply values (Table 6). Such ratios are not explicable in terms of the supply-adsorption model discussed above. One possible explanation is that dissolution of aquifer minerals is a signi"cant source of the supply of Ra. Alternatively, low ratios may indicate the presence of colloids. If colloids are present and incorporate Ra onto sites with a long retention time (i.e. equivalent to a small value of k ), then low values of 1Ra/*Ra ratios in the measured sample could result. Samples from OB1A and 4A showed high Fe and Mn concentrations, which may indicate that one of these processes is important for these bores, though this was not the case for OB2A and 23. Based on the data in Table 6, we conclude that for the Ranger system adsorption/desorption reactions are rapid relative to the Ra and Ra radioactive decay processes, and that the Ra isotope ratios on adsorption sites are not substantially di!erent from those in the water column. This being the case, we can infer the proportion of the total Ra activity which is in the precipitated phase (P ) from the values for Ra/Ra ratios in the water (R) as follows: 0.046 P "1! . R Values calculated for P for May 1993 were 0.78$0.02, 0.89$0.01, 0.80$0.06, 0.95$0.01 and 0.57$0.16 for OB11A, 13A, 15, 16 and 44, respectively. These calculations assume that e!ects such as dissolution of aquifer minerals do not have a signi"cant e!ect on the Ra/Ra ratio in these bores. This may not be the case for OB11A, 13A and 15, as they showed increasing time-series Mn concentrations, and so the values for P probably represent minima for these bores. Use of the Ra/Ra ratio as a separate quantitative measure of Ra removal does not seem practical due to several complicating factors. One of these is the e!ect of reduced support for Ra in the water column following Ra removal; evidence for such reduced support includes the reduction observed in the Ra/Ra ratio in OB11A and 16 (Table 3). Other complicating factors include Ra decay and ingrowth in the barite and water phases over the course of the study, a probably variable supply of Ra to the water phase via a-recoil following decay of its parent in the barite phase, and possible seasonal e!ects. The mean Ra/Ra ratio of 0.60 for bores listed in Table 6 is signi"cantly lower than the average Th/U activity ratio of 1.2 for crustal rocks (Gascoyne, 1992). In general, higher Ra/Ra and Ra/Ra ratios are observed for bores to the west of the tailings dam (i.e. further from the two major orebodies; see Fig. 1). Dickson et al. (1983) proposed a scheme for rating the signi"cance of groundwaters with respect to U exploration, in which Ra/Ra ratios greater than approximately 0.6 would generally result in a rating of &&poor''. The data in Table 6 are in broad agreement with this value, although the mean Ra/Ra ratio of 0.69 for OB29 is surprisingly high, despite its proximity to orebody 3.
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285
5. Conclusions This study demonstrates the usefulness of isotopes of Ra, and of their Th and Ac parents, in studying adsorption}desorption interactions and SO\ mineral formation in groundwater systems. The short-lived isotopes Ra and Ra are particularly useful in studying adsorption}desorption interactions because they generally travel only short distances in an aquifer and are quickly replenished following disturbance of a steady-state condition. In the Ranger case, increasing major cation concentrations lead to increasing Ra and Ra concentrations. However, a reduction in the Ra/Ra ratio with time in two bores appears to have been caused by removal of Ra to a solid phase, with subsequent lack of support for Ra in the water. The Ra/Ra ratio proved to be a good indicator of the recent (compared with the half-life of Ra) formation of a Ra solid solution. For groundwater to the north of the Ranger tailings dam, we conclude that the mineral phase involved is barite, on the basis of saturation state calculations and the time-series concentration behaviour of Ba and Sr. Increases in the Ra/Ra ratio above the expected value of 0.046 were generally observed when the calculated barite saturation states were greater than about 1}2. Time-series increases in Ac concentrations were observed in seepagea!ected bores, but these were insu$cient to account for the high Ra/Ra ratios. Calculations based on the isotope ratio imply that up to 95% of the &&available'' Ra has been removed by solid solution in barite. Ratios lower than the theoretical parent decay supply values for Ra/Ra and Ra/Ra observed for samples from a few bores indicates a possible limitation on this method of detection of Ra solid solution. The presence of high Fe and Mn concentrations in water from two of these bores may mean that dissolution of aquifer minerals or the presence of colloids play a role in this case. In the Ranger case, the above in#uences are at present great enough to mask any Ra concentration increases which may be due to direct transport from the tailings dam.
Acknowledgements pH and major ion water quality data used in this paper were obtained from ERARanger environmental monitoring reports. We also thank Mr P. Cusbert, Mr G. Hancock, Dr C. leGras, Mr R. Marten, Ms T. Mitchell, Ms S. Paulka, Mr J. P"tzner and Mr B. Ryan for their assistance with sample collection and analysis.
References Ahmad, M., & Green, D. C. (1986). Groundwater regimes and isotopic studies, Ranger mine area, Northern Territory. Australian Journal of Earth Sciences, 33, 391}399. Appelo, C. A. J., & Postma, D. (1994). Geochemistry, groundwater and pollution. A.A. Balkema, Rotterdam, Netherlands. Davidson, M. R., & Dickson, B. L. (1986). A porous #ow model for steady state transport of radium in groundwater. =ater Resources Research, 22, 34}44.
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Dean, J. R., Bland, C. J., & Levinson, A. A. (1983). The measurement of Ra/Ra activity ratios in groundwater as a uranium exploration technique. Journal of Geochemical Exploration, 19, 187}193. Dickson, B. L. (1985). Radium isotopes in saline seepages, south-western Yilgarn, Western Australia. Geochimica et Cosmochimica Acta, 49, 361}368. Dickson, B. L. (1990). Radium in groundwater. ¹he environmental behaviour of radium (Vol. 1, Chap. 4}2, pp. 335}372). Vienna:. IAEA Dickson, B. L., Giblin, A. M., & Snelling, A. A. (1987). The source of radium in anomalous accumulations near sandstone escarpments, Australia. Applied Geochemistry, 2, 385}398. Dickson, B. L., Meakins, R. L., & Bland, C. J. (1983). Evaluation of radioactive anomalies using radium isotopes in groundwaters. Journal of Geochemical Exploration, 19, 195}205. Gascoyne, M. (1992). Geochemistry of the actinides and their daughters. In M. Ivanovich, & R. S. Harmon (Eds.), ;ranium series disequilibrium: Applications to environmental problems (2nd ed.). Oxford: Clarendon Press. Hancock, G. J., & Martin, P. (1991). Determination of Ra in environmental samples by alpha-particle spectrometry. Applied Radiation and Isotopes, 42, 63}69. Kraemer, T. F., & Reid, D. F. (1984). The occurrence and behavior of radium in saline formation water of the U.S. Gulf Coast region. Isotope Geoscience, 2, 153}174. Krishnaswami, S., Bhusan, R., & Baskaran, M. (1991). Radium isotopes and Rn in shallow brines, Kharaghoda (India). Chemical Geology (Isot. Geosci. Sect.) 87, 125}136. Krishnaswami, S., Graustein, W. C., Turekian, K. K., & Dowd, J. F. (1982). Radium, thorium and radioactive lead isotopes in groundwaters: Application to the in situ determination of adsorptiondesorption rate constants and retardation factors. =ater Resources Research, 18, 1663}1675. Ku, T.-L., Luo, S., Leslie, B. W., & Hammond, D. E. (1992). Decay-series disequilibria applied to the study of rock-water interaction and geothermal systems. In: M. Ivanovich & R.S. Harmon (Eds.) ;ranium series disequilibrium: Applications to environmental problems (2nd ed.) Oxford: Clarendon Press. Kuptsov, V. M., Cherdyntsev, V. V., & Zverev, V. L. (1969). Radioisotopes in water of neovolcanic regions of the Caucasus. Geochemistry International, 6, 919}922. Langmuir, D., & Melchior, D. (1985). The geochemistry of Ca, Sr, Ba and Ra sulfates in some deep brines from the Palo Duro Basin, Texas. Geochimica et Cosmochimica Acta, 49, 2423}2432. Langmuir, D., & Riese, A. C. (1985). The thermodynamic properties of radium. Geochimica et Cosmochimica Acta, 49, 1593}1601. Laul, J. C. (1992). Natural radionuclides in groundwaters. Journal of Radioanalytical and Nuclear Chemistry Articles, 156, 235}242. Lieser, K. H., & Hill, R. (1992). Hydrolysis and colloid formation of thorium in water and consequences for its migration behaviour * comparison with uranium. Radiochimica Acta, 56, 37}45. Lieser, K. H., Hill, R., MuK hlenweg, U., Singh, R. N., Tu Shu-de, & Steinkop! Th. (1991). Actinides in the environment. Journal of Radioanalytical and Nuclear Chemistry Articles, 147, 117}131. Martin, P., Hancock, G. J., Paulka, S., & Akber, R. A. (1995). Determination of Ac by alpha-particle spectrometry. Applied Radiation and Isotopes, 46, 1065}1070. Martin, P., & Murray, A. S. (1991). An investigation of radium isotopes in Nabarlek borewaters. Internal Report 38, Supervising Scientist for the Alligator Rivers Region, Canberra. Murray, A. S., Marten, R., Johnston, A., & Martin, P. (1987). Analysis for naturally occurring radionuclides at environmental concentrations by gamma spectrometry. Journal of Radioanalytical and Nuclear Chemistry Articles, 115, 263}288. Noller, B. N. (1991). Non-radiological contaminants from uranium mining and milling at Ranger, Jabiru, Northern Territory, Australia. Environmental Monitoring and Assessment, 19, 383}400. Osmond, J. K., & Cowart, J. B. (1992). Ground water. In M. Ivanovich, R. S. Harmon (Eds.), ;ranium series disequilibrium: Applications to environmental problems (2nd ed.). Oxford: Clarendon Press. Sturchio, N. C., Bohlke, J. K., & Markun, F. J. (1993). Radium isotope geochemistry of thermal waters, Yellowstone National Park, Wyoming, USA. Geochimica et Cosmochimica Acta, 57, 1203}1214. Whitehead, B. R. (1980). A compilation and interpretation of hydrogeological data, Ranger on site area, September 1980. Northern Territory Geological Survey Technical Report GS 80/33.
Radium isotopes as indicators of adsorption}desorption interactions and barite formation in groundwater Paul Martin *, Riaz A. Akber Environmental Research Institute of the Supervising Scientist (ERISS), Jabiru, NT 0886, Australia Queensland University of Technology, GPO Box 2434, Brisbane, Queensland 4001, Australia Received 14 May 1998; received in revised form 16 September 1998; accepted 20 November 1998
Abstract Borewater monitoring near the Ranger U mine tailings dam has revealed deterioration in water quality in several bores since 1981, with increases in sulphate concentrations of up to 1000 times being observed. In some cases, signi"cant increases in Ra concentrations also occurred. In this study, isotopes of Ra, Th and Ac were measured in borewater in 1988}1993. For the most seepage-a!ected bores, the concentrations of Ra, Ra and Ra as well as the nuclidic ratios Ra/Ra, Ra/Ra and Ra/Ra generally increased. Time-series concentration increases were observed for Sr but not for Ba. Ac concentrations increased with time, but not su$ciently to account for the increasing Ra concentrations. It is concluded that increases observed in Ra isotope concentrations arise from competition for cation adsorption sites in the vicinity of the bore rather than by direct transport of Ra from the tailings. Formation of a barite solid phase is occurring in the groundwater and causing the removal of some Ra from solution, with rapid replenishment of the shorter-lived isotopes from their parents. 1999 Elsevier Science Ltd. All rights reserved. Keywords: Radium; Actinum; Thorium; Barite; Groundwater; Uranium mining
1. Introduction Signi"cant changes in groundwater chemistry and, in some cases, Ra concentrations were observed in the vicinity of the Ranger U mine tailings dam since the beginning of the mining operation. In particular, Ra concentrations increased by factors of about 2}4 between 1982 and 1988 for three bores (OB11A, OB13A and OB16; Fig. 1). For those bores a!ected by seepage, concentrations of Ca, Mg *Corresponding author. Tel.: 0061 889 799 711; fax: 0061 889 792 076; e-mail: [email protected] 0265-931X/99/$ - see front matter 1999 Elsevier Science Ltd. All rights reserved. PII: S 02 6 5-9 3 1X (9 8) 0 0 14 7 - 7
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Fig. 1. Ranger bore location plan.
and SO\ increased with time. In this paper, we report the results of a study of the behaviour of Ra isotopes as a function of the concentrations of these major ions. The four natural-series Ra isotopes are: Ra (t "1600 yr) of the U series, Ra (11.4 d) of the Ac series, and Ra (3.66 d) and Ra (5.75 yr) of the Th series. The natural activity ratio U/U is 0.046, and so the activity concentrations of nuclides of the Ac series are normally very low in environmental samples. The Th/U activity ratio is variable in nature due to the quite di!erent geochemical behaviour of Th and U. The di!erences in their half-lives makes the Ra isotopes potentially useful geochemical tools, and they have been used in studies of Ra migration in aquifers (Dean, Bland & Levinson, 1983; Davidson & Dickson, 1986; Dickson, Giblin & Snelling, 1987), determination of in situ retardation factors and distribution coe$cients (Laul, 1992), adsorption}desorption reaction rates
P. Martin, R.A. Akber/J. Environ. Radioactivity 46 (1999) 271}286
273
(Krishnaswami, Graustein, Turekian & Dowd, 1982; Krishnaswami, Bhusan & Bhaskaran, 1991) and for geochemical exploration (Dickson, Meakins & Bland, 1983). A number of studies of Ra isotopes in groundwater have found cases where the Ra/Ra and Ra/Ra ratios were greater than the theoretical parent decay supply values of 0.046 and 1.0, respectively. Dickson (1990) reported that for a sample set of over 400 Australian groundwaters the median value for Ra/Ra was 0.075. Kuptsov et al. (1969) measured Ra/Ra ratios between 3 and 105 in carbonate spring waters. Such ratios are often associated with elevated sulphate concentrations (for example, Dickson, 1985; Krishnaswami et al., 1991; Martin & Murray, 1991; Sturchio, Bohlke & Markun, 1993), and were also observed for some seepage-a!ected bores in the present study. In this paper, possible causes of the high ratios measured in the Ranger borewaters are discussed.
2. Study area The ERA-Ranger U mine and mill complex is located approximately 260 km east of Darwin in the Alligator Rivers Region of northern Australia. Excavation of orebody No. 1 commenced in August 1980 and ended in October 1994. Excavation of orebody No. 3 began in 1996. Commercial operation of the mill commenced in August 1981. Current annual production is about 4000 tonnes U O . The climate is tropical monsoonal, with contrasting wet (approximately December}April) and dry (approximately May}November) seasons. Mean annual rainfall is about 1480 mm and the mean annual Class A pan evaporation is about 2590 mm. Release of contaminated water into the o!-site environment is minimised by use of a number of waterbodies with varying water quality, including four retention ponds (RP1}RP4), the minepits and a tailings dam (Fig. 1). Covering approximately 110 ha of land, the tailings dam is the largest of the water management structures. Subaqueous tailings deposition was practiced until 1987, but since then tailings have been kept largely sub-aerial in order to achieve a greater density. At the Ranger mill, U is leached from the ore by H SO using pyrolusite as an oxidant (Noller, 1991). Sulphate is also produced due to the oxidation of sulphidebearing minerals present in the catchment area of the retention ponds. Consequently, on-site waterbodies are generally characterised by high sulphate levels. Sulphate concentrations in tailings dam surface water were approximately 70 and 300 mM in 1983 and 1993, respectively. Corresponding Mn concentrations were approximately 2 and 20 mM, respectively. Tailings are neutralised before placing in the tailings dam. Two aquifers exist in the Ranger on-site area (Whitehead, 1980; Ahmad & Green, 1986), an uncon"ned shallow aquifer varying in thickness from 2 to 20 m and associated with the soil pro"le and underlying laterite layer, and a con"ned to semi-con"ned deep aquifer in underlying weathered and fresh bedrock. A layer of low permeability sandy silty clay is commonly found between the laterite and the weathered bedrock. Where it occurs, this layer forms a semi-con"ning upper boundary to the deep aquifer. Both aquifers are a!ected by the seasonal nature of the rainfall, with water levels in deep aquifer bores falling by about 1}2 m between May and November.
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Table 1 Bore slotted depth, water temperature and Fe, Mn and Pb borewater concentrations Bore
Slotted depth (m)
T (3C)
OB1A OB2A OB4A OB6A OB9A OB10A OB11A OB12A OB13A OB15 OB16 OB17A OB19A OB21A OB22 OB23 OB24 OB29 OB44
16}31 15}30 22}37 14}26 12}30 12}30 11}23 16}31 10}30 10}25 10}20 23}41 * 31}43 25}41 36}51 22}36 35}50 11}16
32 * 34 33 34 31 33 33 33 28 35 * 34 31 * * * 31 28
Fe (lM)
2 :1 3 4 2 :1 3 13 :1 :1 4 * 2 5 2 * :1
Mn (lM)
Pb (nM)
21 0.18
:0.7 * 1.5 * 0.8 5.2 6.8 :1 * 6.3 * 1.1 1.5 * * * * * *
0.27 0.88 0.36 0.20
0.74 0.55 5 3.7 0.18 3.3 0.09 0.13
Time-series changes have been observed for these bores; see text.
A number of observation bores have been drilled in the area for groundwater quality monitoring purposes. In this paper we deal only with bores accessing the deeper aquifer (Fig. 1). All the bores discussed have a PVC casing; where known, bore slotted depths are given in Table 1.
3. Sampling and analytical methods Sampling was timed to coincide with the routine groundwater sampling program of Ranger, most collections occurring in May or November. Containers and lids were washed with 2% HNO and rinsed with demineralised water before use, and rinsed three times with bore water immediately before collecting the sample. Samples were collected approximately 20 min after the start of pumping. The pump rate was approximately 5 l min\. Samples were "ltered through a 0.45 lm "lter before acidi"cation to 1% HNO before analysis. Only one sample was collected for analysis on each sampling occasion. Uncertainties quoted for radionuclide analyses are one standard deviation due to counting statistics only. Since the main focus of the work was Ra behaviour, all samples were analysed for Ra and Ra, and most for the shorter-lived Ra and Ra. Some samples were also analysed for Th and Ac isotopes, primarily in order to investigate whether they were supporting their Ra progeny in the water.
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275
Most of the radionuclide analyses were performed by a-ray spectrometry (Hancock & Martin, 1991; Martin, Hancock, Paulka & Akber, 1995). Samples were collected into a 5 l polypropylene container. Chemical separations for radium analysis were generally performed within 2 days after sample collection. Results for Ra, Ra and Ra were corrected to the date and time of sample collection assuming nil support from the Th parent. Ra/Ra ratios were determined by recounting Ra sources after waiting at least 12 months for ingrowth of Th and Ra. The Ra/Ra ratio in tailings dam water was determined by washing the deposit from the source several years after preparation and determining Th on the washings. Ac (t "21.8 yr) was determined by chemical separation of Ac, with subsequent detection of the a-decays of the Th and Ra progeny; Ac was used as a yield tracer. Th (t "1.91 yr) determinations were performed 8 days or less after sample collection, and corrected for ingrowth from Ra. Th (t "18.7 d) was determined by two separate determinations of Ra on the sample, one performed soon after sample collection and the second after a delay of approximately 2 weeks. For some samples Ra and Ra were determined by c-ray spectrometry. In this case, a 20 l sample was collected into a PVC pail and Ra co-precipitated with MnO , followed by a second co-precipitation with Fe(OH) . The precipitate was "ltered and ashed, then cast in a mould with a polyester resin before counting on a Ge c-ray spectrometer (Murray, Marten, Johnston & Martin, 1987). Analyses for Fe, Mn, Ba and Sr were performed by ICP-AES; those for Pb were performed by graphite furnace atomic absorption spectrometry. Borewater temperatures were measured in situ after pumping, using a K-type thermocouple.
4. Results and discussion 4.1. General chemical parameters Groundwaters of the aquifer are Mg>}Na>}HCO\ type, with low concentrations of SO\ and Cl\. Conductivities are generally in the range of 100}400 lS cm\, with pH values of 6.2}7.5. SO\ concentrations in a group of bores to the north of the tailings dam (OB11A, 13A, 15, 16 and, to a lesser extent, 44) increased since 1982 (Fig. 2). Anion equivalences are now dominated by SO\ for these bores. Concentrations of Ca>, Mg>, Na> and K> have all increased. The cation equivalent dominance of Mg> also increased. The pH of water from these bores decreased by 0.5}1.0 pH units over this period, although most of this decrease occurred prior to 1988. Increasing SO\ concentra tions were observed in OB6A (from 1983), OB9A and 10A (from 1986), OB4A (from 1989) and OB2A (from 1991). However, these increases were not as pronounced as those for the bores to the north of the dam. Ranger's monitoring data show that between 1981 to 1993, Mn concentrations in bores OB4A, 11A, 13A, 15 and 29 increased from about 2}15, 1}5, 0.5}7, 1}10 lM and 0.04}1.5 lM, respectively (ERA Ltd, unpublished data). For the remaining bores, Mn concentrations were relatively constant. Table 1 shows measurements of temperature
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Fig. 2. Time-series concentrations of SO\.
and Fe, Mn and Pb concentrations made as part of the present study. Fe concentrations were high and increasing with time in bores OB1A and 4A. Individual Fe measurements were 210 and 470 lM for OB1A in May 1991 and November 1992, respectively, and 140, 310 and 430 lM for OB4A in November 1990, November 1992 and May 1993, respectively. 4.2. Ra isotope results Fig. 3a shows the activity concentrations of Ra for the "ve bores most a!ected by seepage (OB11A, 13A, 15, 16 and 44) between 1988 and 1993. A general trend of increasing Ra concentrations with time is apparent over this period. As discussed further below, there is a possible seasonal variability in Ba concentrations in these borewaters, and so the signi"cance of any apparent trend in Ra concentrations is best established using same-month comparisons. Table 2 lists the available data for Ra isotope activity concentrations for the Novembers of years 1988}1992. For Ra there was a signi"cant (2p) increase for OB11A, 13A, 15 and 16 over this period. Signi"cant increases were also observed for Ra and Ra for bores OB11A, 13A and 15. These results are explicable in terms of increasing groundwater salinity. It is well established that increasing cation concentrations will lead to competition with Ra for adsorption sites and consequently higher groundwater Ra concentrations (Kraemer & Reid, 1984; Langmuir & Melchior, 1985; Ku, Luo, Leslie & Hammond, 1992). The reduction in pH values observed for these bores could also have contributed to this desorption e!ect. Fig. 3b shows that although Ra activity concentrations increased signi"cantly in OB11A over the 1988}1992 period, they #uctuated in OB13A and 15 and decreased in OB16. These results, especially the pronounced Ra decrease in OB16 following a concentration peak in 1989, as well as the decreases observed for Ra and Ra in
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277
Fig. 3. Time-series concentrations of (a) Ra, (b) Ra and (c) Ac.
OB16 (Table 2), do not appear to be consistent with the hypothesis of desorption due to major cation concentration increases. Ranger's monitoring data for tailings dam surface water show Ra concentrations in the range 5.3}33 Bq l\ (n"37) between 1988 and 1993. Measurements of the Ra/Ra ratio in "ltered tailings dam water samples collected in July 1988 and December 1991 were 0.0010$0.0003 and 0.002$0.001, respectively. Consequently, if a signi"cant migration of Ra from the tailings dam were occurring, a decrease in Ra/Ra ratios in borewater would be expected to occur over time. Table 3 shows the percentage change (*Ra) in a number of Ra isotope ratios between November 1988 and the Novembers of subsequent years, calculated as
*Ra"
(R !R ) 66 ;100% R
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Table 2 Ra, Ra, Ra and Ra activity concentrations for Novembers of the given year Bore
Year
Ra (mBq l\)
Ra (mBq l\)
Ra (mBq l\)
Ra (mBq l\)
OB11A
1988 1989 1990 1991 1992 1988 1989 1992 1988 1989 1990 1992 1988 1989 1991 1992 1989
134$3 143$3 192$17 177$3 169$5 66$1 83$3 56$2 62$4 54$2 76$5 48$2 80$2 105$3 37$3 30$1 49$2
51$1 75$3 86$8 94$5 100$4 92$3 96$5 94$4 142$14 159$10 244$17 196$9 129$4 106$7 79$10 66$3 23$4
57$3 86$5 105$10 * 124$7 125$5 198$10 247$16 129$10 150$9 245$19 348$20 247$11 * * 207$17 28$3
8.7$0.7 13.4$1.3 18.2$1.9 * 24.7$2.3 11.2$1.0 15.9$1.1 26.5$2.7 6.8$0.9 8.2$1.1 12.0$1.6 16.2$1.6 21.0$1.6 * * 29.9$3.6 5.6$0.9
OB13A
OB15
OB16
OB44
Table 3 Percentage changes in Ra/Ra, Ra/Ra, Ra/Ra, Ra/Ra and Ra/Ra activity ratios between November 1988 and November of the given year Bore
Year
*Ra86 (%)
*Ra46 (%)
*Ra36 (%)
*Ra48 (%)
*Ra43 (%)
OB11A
1989 1990 1991 1992 1989 1992 1989 1990 1992 1989 1991 1992
38$6 17$3 39$7 55$5 917$4 20$5 29$11 41$10 78$13 937$4 31$13 36$5
40$10 28$6 * 72$11 26$6 130$20 33$8 55$9 250$20 * * 120$20
45$18 46$14 * 125$30 13$12 180$40 40$20 40$20 210$40 * * 280$50
2$9 10$6 * 12$8 53$10 96$15 3$11 10$10 96$17 * * 64$15
*9$13 *17$8 * 929$9 11$12 917$11 93$16 7$16 13$16 * * 944$8
OB13A OB15
OB16
where R " the activity ratio in November 1988 and R " the activity ratio in 66 November 19XX. Over the course of the study the Ra/Ra ratio actually increased signi"cantly in all four bores. The e!ect of direct migration of Ra from the dam must, therefore, be small in magnitude in comparison with other e!ects at this stage.
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P. Martin, R.A. Akber/J. Environ. Radioactivity 46 (1999) 271}286 Table 4 Th, Th and Th activity concentrations, November 1989 Bore
Th (mBq l\)
Th (mBq l\)
Th (mBq l\)
OB9A OB11A OB13A OB15 OB44
:0.2 :0.2 :0.3 :0.1 :0.3
:0.1 :0.1 :0.1 :0.1 :0.2
:0.4 :0.4 :0.4 :0.9 :0.7
*Ra values for Ra/Ra and Ra/Ra ratios were generally higher than those for the Ra/Ra ratio. Given their relatively short half-lives, Ra and Ra would not generally be expected to migrate long distances in an aquifer (Osmond & Cowart, 1992). If this assumption is correct, then the Ra/Ra ratio would not be expected to vary signi"cantly with time. *Ra values (Table 3) show that bore OB16 has experienced a decreasing Ra/Ra ratio, almost having its value between 1988 and 1992. This ratio also decreased signi"cantly in OB11A over this period. These results will be discussed further below. 4.3. Ac, Th and Ra isotope concentrations Table 3 shows that for the most seepage-a!ected bores the Ra/Ra and Ra/Ra ratios have increased with time. These ratios are now signi"cantly above the theoretical parent decay supply values of 0.046 and 1.0, respectively, in OB11A, 13A, 15, 16 and 44. Such increases could arise from increased support for Ra and Ra from their parents in the water column. In a study of Ra isotopes in borewaters and saline seepages in Western Australia, Dickson (1985) reported Ra/Ra ratios in the range of 0.14}0.62. High Ra/Ra ratios were generally associated with high Cl\ and SO\ concentra tions. Dickson postulated that soluble single and double SO\ salts of Ac may be formed, resulting in support for Ra in the water. His measurements of the Western Australian waters indicated that there was indeed some support for Ra in the water, with most Ac/Ra ratios in the range 0.25}0.8. Ac activity concentrations in Ranger bores OB11A, 13A, 15, 16 and 44 have increased with time (Fig. 3c), showing that some Ac mobilization is occurring. However, in none of the borewaters tested did the Ac activity concentration exceed 20% of that of Ra. On most occasions Ac support in the water could account for no more than 4% of the Ra excess in OB13A, 15 and 16, and about 10}20% of the excess in OB11A and 44. In addition, Ac does not decay directly to Ra, but passes through the intermediate Th isotope Th. In the pH range prevailing in these bores, most of this Th can be expected to be lost from the water column due to adsorption processes (Lieser & Hill, 1992), reducing further the degree of support for Ra in the water. Measurements of the members of the Ac, U and Th decay series (Tables 4 and 5) con"rm that Ra isotopes are in substantial excess over their Th parents in the water.
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Table 5 Th and Ra concentrations, September 1988 Bore
Th (mBq l\)
Ra (mBq l\)
OB11A OB13A OB16
:3 :4 :4
9.3$0.9 12.0$0.8 21$1
The di!erence in degree of Ac mobilization between the Ranger borewaters and the Western Australian waters described by Dickson is probably due to the greater acidity (pH 4}5) of the latter. Lieser et al. (1991) reported the results of Ac adsorption experiments carried out for a groundwater from Gorleben (Germany). They found that salinity had an appreciable e!ect on Ac adsorption at pH 3.5, but not at pH 7.5. Their conclusion was that at pH 7.5 hydrolysis of Ac is practically complete, whereas at pH 3.5 hydrolysis is incomplete and Na> ions can compete with the ionic species of Ac for ion exchange sites. 4.4. Evidence for formation of a barite phase Langmuir and Riese (1985) and Langmuir and Melchior (1985) studied the trace aqueous and solid solution chemistry of Ra, and concluded that, since Ra concentrations in natural waters and waters associated with U mining are probably never high enough to reach saturation with respect to a pure Ra solid, concentrations are limited by adsorption or solid solution formation. On the basis of the results presented above, we conclude that a combination of both of these processes is a!ecting Ra isotope concentrations in the Ranger borewaters. Formation of a solid solution would result in the removal of Ra from the water. Subsequent ingrowth due to the decay of the parents in the source rocks will be more rapid for short-lived Ra isotopes (hereafter designated 1Ra) than for long-lived ones (*Ra). Consequently, the 1Ra/*Ra activity ratio in the water should increase for a period dependent on the isotope half-lives. Analytical data were used in conjunction with the program WATEQP (Appelo and Postma, 1994) to calculate the saturation states of SO\ mineral phases in these waters. The saturation state (X) is given by X"IAP/K where IAP is the ion activity product and K is the solubility product of the mineral at the speci"ed temperature and pressure. Ion activity products are calculated from the activities of the uncomplexed ions in solution, for example, IAP "[Ba>][SO\], where the quantities in square brackets represent activities of the uncomplexed ions in a water sample. Waters for which ) is greater than 1 are termed supersaturated, while those for which ) is less than 1 are subsaturated. Ranger borewaters were found to be
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281
Fig. 4. Ra/Ra vs. the barite saturation state ()).
Fig. 5. Time-series concentrations of (a) Sr and (b) Ba.
subsaturated with respect to anglesite (PbSO ) and celestite (SrSO ). However, for a number of bores the waters were supersaturated with respect to barite (BaSO ). Fig. 4 shows the activity ratio Ra/Ra plotted against the calculated barite saturation state in borewater samples for which su$cient data are available. The Ra/Ra ratio is generally signi"cantly greater than 0.046 for those bores with ) greater than about 1}2, in agreement with the prediction above of increasing 1Ra/*Ra ratios following formation of a solid solution.
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Sr concentrations measured for bores OB4A, 9A, 11A, 13A, 15, 16 and 44 increased with time (Fig. 5a). This behaviour is to be expected as Sr will be removed from adsorption sites by increased salinity in a similar mechanism to that a!ecting Ra. In contrast, Ba concentrations for OB11A, 13A, 15, 16 and 44 have tended to #uctuate, generally being lower in November (i.e. late dry season/early wet season) than in May (late wet season/early dry season), and showed no apparent increase with time since 1988 (Fig. 5b). In light of the saturation state calculations discussed above, this is good evidence that a barite phase is being formed in these groundwaters. Although the most recent samples from OB4A and 6A showed calculated values for ) greater than 1, an increase in the Ra/Ra ratio was not observed (Fig. 4). We interpret this as indicating that barite formation did not occur in the vicinity of these bores during the study. The fact that Ba concentrations increased in water from bore OB4A and were relatively steady in OB6A (Fig. 5b) supports this conclusion. 4.5. Implications for modelling of Ra behaviour Results presented above demonstrate that the primary source of Ra in borewater from the Ranger site is the aquifer rocks in the vicinity of the bore, rather than transport from the tailings dam. Further, for some bores increasing SO\ concentra tions led to the presence of a barite phase with co-precipitation of Ra. It follows that a model based upon retardation coe$cients cannot predict the behaviour of radium in this system. The potential in#uences on Ra isotope activity concentrations are: (1) Ingrowth via decay of the ¹h parent in the water: Th isotope measurements showed that there is little direct support for Ra from dissolved parents in the Ranger borewaters. (2) Release of Ra following dissolution of aquifer minerals: Increases in Mn and Fe concentrations observed in a few bores (Table 1) may be due to dissolution of native oxides, although direct transport of Mn from the tailings dam cannot be ruled out. Dissolution of Mn or Fe oxides could be expected to increase Ra isotope concentrations in the water, but be accompanied either by no change in or by reduction in 1Ra/*Ra ratios. Consequently, if it is occurring, then the e!ect on the ratios must be small compared with other factors in the bores to the north of the tailings dam. (3) Ingrowth from the parent in the solid by a-recoil processes: The a-recoil is a physical process and is unlikely to be a!ected substantially by a change in water chemistry. For this reason, we will assume here that supply rates of di!erent Ra isotopes by this process remained constant. (4) Adsorption onto and desorption from solid surface sites: Increasing groundwater salinity should result in increased Ra activity concentrations in the water. The e!ect on Ra isotope ratios would depend upon the adsorption and desorption rate constants, as discussed below. (5) Removal of Ra from solution by a precipitating mineral phase: This process would result in a decrease in activity concentrations of all Ra isotopes in solution in the short term. Subsequent ingrowth of short-lived isotopes would result in an increase in 1Ra/*Ra ratios.
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283
Table 6 Mean (standard deviation) for Ra activity concentration and Ra/Ra, Ra/Ra and Ra/Ra activity ratios for a number of bores. For n "1 the uncertainty is one standard deviation due to counting statistics only Bore
Ra (mBq l\)
Ra/Ra
n
Ra/Ra
Ra/Ra
Ra/Ra
n
OB1A OB2A OB4A OB6A OB7A OB9A OB10A OB12A OB17A OB18A OB19A OB20 OB21A OB22 OB23 OB24 OB28 OB29 OB30 OB41 OB43
36 (2) 25 (6) 96 (19) 330 (48) 104 (6) 68 (20) 73 (14) 34 (1) 84 (34) 70 (6) 120 (120) 41$2 200$40 160 (50) 170 (90) 24$1 97$3 61 (33) 20$2 34$3 31$2
0.7 (0.3) 1.0 (0.5) 0.53 (0.04) 0.18 (0.02) 0.13 (0.04) 0.27 (0.03) 0.10 (0.03) 2.1 (0.1) 1.3 (0.7) 1.5 (0.2) 0.58 (0.08) 0.68$0.06 0.26$0.03 0.06 (0.02) 0.54 (0.05) 0.29$0.03 0.24$0.04 0.62 (0.35) 0.52$0.11 0.56$0.11 0.50$0.12
2 2 7 11 2 7 6 2 5 3 2 1 1 3 2 1 1 3 1 1 1
0.033$0.008 0.029$0.004 0.037 (0.004) 0.043 (0.006) 0.026$0.004 0.046 (0.016) 0.032 (0.006) 0.105$0.008 0.029 (0.014) 0.047 (0.004) 0.044$0.003 9 * 0.059 (0.011) 0.021$0.007 0.028$0.019 * 0.062 (0.013) * * *
0.66$0.12 0.77$0.05 0.62 (0.12) 1.06 (0.23) 1.21$0.15 0.84 (0.20) 1.03 (0.31) 1.04$0.05 0.82 (0.21) 0.65 (0.06) 1.03$0.05 * * 0.61 (0.23) 0.40$0.07 1.20$0.39 * 1.13 (0.16) * * *
0.32$0.06 1.06$0.04 0.33 (0.06) 0.19 (0.04) 0.128$0.015 0.23 (0.05) 0.09 (0.02) 2.23$0.06 0.82 (0.36) 1.00 (0.13) 0.538$0.013 * * 0.036 (0.006) 0.23$0.04 0.35$0.10 * 0.69 (0.46) * * *
1 1 6 9 1 6 4 1 4 3 1 * * 3 1 1 * 2 * * *
89 (75)
0.60 (0.51)
0.043 (0.021)
0.87 (0.25)
0.55 (0.57)
Mean (s.d.)
n"number of samples for Ra and Ra Ra. n"number of samples for Ra/Ra, Ra/ Ra and Ra/Ra.
Following the model and notation of Krishnaswami et al. (1982), we can represent the adsorption/desorption process by "rst-order adsorption (k ) and desorption (k ) rate constants of dimension time\. The values of k and k will depend on the system under consideration, and in particular on the major ion chemistry of the water phase. In general, k will be several orders of magnitude smaller than k and most of the Ra potentially available for dissolution in the groundwater will reside on adsorption sites, although this may not be the case for highly saline waters (Laul, 1992). Where k is of the order of or less than the decay constant (j) of a short-lived Ra isotope, the adsorption process will result in the 1Ra/*Ra ratio in the water phase being higher than that available from supply by parent decay. Table 6 shows the means of all readings for Ra/Ra and Ra/Ra in Ranger bores except OB11A, 13A, 15, 16, and 44. These values should then represent the situation in Ranger bores when the major processes occurring are supply by parent decay and adsorption/desorption reactions. The mean values for Ra/Ra and Ra/Ra over these bores are 0.043 and 0.87, respectively. Ninety-"ve percent con"dence
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intervals for these means (0.043$0.012 and 0.87$0.14) included the theoretical parent decay supply values of 0.046 and 1.0, respectively. For a few bores (OB1A, 2A, 4A and 23), both of these ratios were substantially lower than the theoretical parent decay supply values (Table 6). Such ratios are not explicable in terms of the supply-adsorption model discussed above. One possible explanation is that dissolution of aquifer minerals is a signi"cant source of the supply of Ra. Alternatively, low ratios may indicate the presence of colloids. If colloids are present and incorporate Ra onto sites with a long retention time (i.e. equivalent to a small value of k ), then low values of 1Ra/*Ra ratios in the measured sample could result. Samples from OB1A and 4A showed high Fe and Mn concentrations, which may indicate that one of these processes is important for these bores, though this was not the case for OB2A and 23. Based on the data in Table 6, we conclude that for the Ranger system adsorption/desorption reactions are rapid relative to the Ra and Ra radioactive decay processes, and that the Ra isotope ratios on adsorption sites are not substantially di!erent from those in the water column. This being the case, we can infer the proportion of the total Ra activity which is in the precipitated phase (P ) from the values for Ra/Ra ratios in the water (R) as follows: 0.046 P "1! . R Values calculated for P for May 1993 were 0.78$0.02, 0.89$0.01, 0.80$0.06, 0.95$0.01 and 0.57$0.16 for OB11A, 13A, 15, 16 and 44, respectively. These calculations assume that e!ects such as dissolution of aquifer minerals do not have a signi"cant e!ect on the Ra/Ra ratio in these bores. This may not be the case for OB11A, 13A and 15, as they showed increasing time-series Mn concentrations, and so the values for P probably represent minima for these bores. Use of the Ra/Ra ratio as a separate quantitative measure of Ra removal does not seem practical due to several complicating factors. One of these is the e!ect of reduced support for Ra in the water column following Ra removal; evidence for such reduced support includes the reduction observed in the Ra/Ra ratio in OB11A and 16 (Table 3). Other complicating factors include Ra decay and ingrowth in the barite and water phases over the course of the study, a probably variable supply of Ra to the water phase via a-recoil following decay of its parent in the barite phase, and possible seasonal e!ects. The mean Ra/Ra ratio of 0.60 for bores listed in Table 6 is signi"cantly lower than the average Th/U activity ratio of 1.2 for crustal rocks (Gascoyne, 1992). In general, higher Ra/Ra and Ra/Ra ratios are observed for bores to the west of the tailings dam (i.e. further from the two major orebodies; see Fig. 1). Dickson et al. (1983) proposed a scheme for rating the signi"cance of groundwaters with respect to U exploration, in which Ra/Ra ratios greater than approximately 0.6 would generally result in a rating of &&poor''. The data in Table 6 are in broad agreement with this value, although the mean Ra/Ra ratio of 0.69 for OB29 is surprisingly high, despite its proximity to orebody 3.
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285
5. Conclusions This study demonstrates the usefulness of isotopes of Ra, and of their Th and Ac parents, in studying adsorption}desorption interactions and SO\ mineral formation in groundwater systems. The short-lived isotopes Ra and Ra are particularly useful in studying adsorption}desorption interactions because they generally travel only short distances in an aquifer and are quickly replenished following disturbance of a steady-state condition. In the Ranger case, increasing major cation concentrations lead to increasing Ra and Ra concentrations. However, a reduction in the Ra/Ra ratio with time in two bores appears to have been caused by removal of Ra to a solid phase, with subsequent lack of support for Ra in the water. The Ra/Ra ratio proved to be a good indicator of the recent (compared with the half-life of Ra) formation of a Ra solid solution. For groundwater to the north of the Ranger tailings dam, we conclude that the mineral phase involved is barite, on the basis of saturation state calculations and the time-series concentration behaviour of Ba and Sr. Increases in the Ra/Ra ratio above the expected value of 0.046 were generally observed when the calculated barite saturation states were greater than about 1}2. Time-series increases in Ac concentrations were observed in seepagea!ected bores, but these were insu$cient to account for the high Ra/Ra ratios. Calculations based on the isotope ratio imply that up to 95% of the &&available'' Ra has been removed by solid solution in barite. Ratios lower than the theoretical parent decay supply values for Ra/Ra and Ra/Ra observed for samples from a few bores indicates a possible limitation on this method of detection of Ra solid solution. The presence of high Fe and Mn concentrations in water from two of these bores may mean that dissolution of aquifer minerals or the presence of colloids play a role in this case. In the Ranger case, the above in#uences are at present great enough to mask any Ra concentration increases which may be due to direct transport from the tailings dam.
Acknowledgements pH and major ion water quality data used in this paper were obtained from ERARanger environmental monitoring reports. We also thank Mr P. Cusbert, Mr G. Hancock, Dr C. leGras, Mr R. Marten, Ms T. Mitchell, Ms S. Paulka, Mr J. P"tzner and Mr B. Ryan for their assistance with sample collection and analysis.
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