Uranium in freshwater and marine water

​Toxicant default guideline values for protecting aquatic ecosystems

October 2000

Description of chemical

Speciation

Uranium may occur in natural waters in three oxidation states, U⁴⁺ uranium (IV), UO₂+ [uranium (V)] and UO₂²⁺ uranium (VI)] or uranyl ion]. In reducing surface waters, uranium occurs as U⁴⁺ and UO₂+. It is generally considered that UO₂²⁺ is the form of uranium (VI) primarily responsible for eliciting a toxic response in aquatic organisms (Markich et al. 1996). Typically, inorganic and organic complexes of uranium (VI) ameliorate the uptake and toxicity of uranium (VI) by reducing the concentration of UO₂²⁺.

Uranium (IV) has a strong tendency to precipitate and to remain immobile, whereas UO₂+ forms soluble, but relatively unstable, complexes (Langmuir 1978). In oxidised surface waters, uranium occurs as UO₂²⁺. and forms stable, readily soluble, cationic, anionic and/or neutral complexes which are highly mobile (Langmuir 1978, Osmond & Ivanovich 1992). The redox and complexation reactions of uranium in surface waters are strongly influenced by hydrolysis, since hydrolytic reactions may limit the solubility or influence sorption behaviour (Choppin & Stout 1989). The speciation of uranium is relatively complex in oxidised fresh surface waters (pH 5 to 9) (Grenthe et al. 1992, Palmer & Nguyen-Trung 1995).

In seawater, dissolved uranium exists predominantly as the uranyl-tricarbonate complex UO₂(CO₃)₃⁴¯(Djogic & Branica 1993). Uranyl-DOM complexes form a component (< 20%) of the dissolved uranium concentration (Mann & Wong 1993). This component is strongly dependent on the concentration of DOM.
A variety of methods are available for determining the speciation of uranium in water. These include:

• Analytical techniques, including physical separation (e.g. (ultra)filtration, dialysis, centrifugation), voltammetry (e.g. cathodic stripping voltammetry) ion exchange chromatography, spectroscopy (e.g. time-resolved laser-induced fluorescence) and ligand competition methods (e.g. solvent extraction) (Choppin & Stout 1989, de Beer & Coetzee 1992, Djogic & Branica 1993, Moulin et al. 1995, Meinrath et al. 1996)
• Theoretical techniques, including geochemical modelling (Langmuir 1978, van den Berg 1993, Moulin et al. 1995, Markich et al. 1996, Meinrath et al. 1996).

Bioassays are typically used to ascertain metal-organism interactions. These can be coupled with the measured and/or predicted speciation of uranium to determine the bioavailable uranium species. The current analytical practical quantitation limit (PQL) for uranium is 0.01 µg/L in fresh water and 0.3 µg/L in marine water (NSW EPA 2000).

Factors that affect the toxicity of uranium

Several studies have established that uranium toxicity is inversely related to water hardness and alkalinity (Tarzwell & Henderson 1960, Parkhurst et al. 1984, Poston et al. 1984). Parkhurst et al. (1984) reported that the 96-hour LC50 for brook trout (Salvelinus fontinalis) was 5.5 mg/L in soft water (hardness, 35 mg/L as CaCO₃; alkalinity, 11 mg/L as CaCO₃; pH, 6.7). In contrast it was 23 mg/L in hard water (hardness, 208 mg/L as CaCO₃; alkalinity, 53 mg/L as CaCO₃; pH, 7.5).

Probably the most important complexing agent for uranium in oxidised freshwaters is carbonate (Clark et al. 1995). Markich et al. (1996) showed that the toxicity of uranium to a freshwater bivalve (Velesunio angasi) was inversely proportional to alkalinity, where both pH and water hardness were held constant. Complexes of uranium with carbonate are less toxic than UO₂²⁺ (Nakajima et al. 1979, Poston et al. 1984, Greene et al. 1986). Phosphate is an important complexing agent when its concentration approaches 75 µg/L (Langmuir 1978).

Several studies have shown that the uptake and toxicity of uranium is inversely related to pH, over the range 2 to 7, where both water hardness and alkalinity were held constant (Nakajima et al. 1979, Greene et al. 1986, Markich et al. 1996). Markich et al. (1996) showed that the sublethal toxicity of uranium to V. angasi in a synthetic water was about five times greater at pH 5 (48-hour EC50 = 117 µg/L) than at pH 6 (48-hour EC50 = 634 µg/L). They concluded that changes in uranium speciation were responsible for the changes in toxicity of uranium.

Natural DOM is also a very effective complexing agent of uranium in natural waters (Choppin & Stout 1989, Moulin et al. 1992). Organic matter may act as a sink for uranium, if the uranyl-DOM complex is insoluble, or may serve as a mobile phase, if the uranyl-DOM complex is soluble (Livens et al. 1996). In soft, low-alkaline, organic-rich, fresh surface waters (pH 5 to 7), uranyl-DOM complexes are the dominant species of dissolved uranium; complexation increases with increasing pH (Choppin & Stout 1989). Uranyl carbonate and hydroxide species become more important than uranyl-DOM complexes as the hardness, alkalinity and pH of the water increase (usually pH > 7 to 8) (Moulin et al. 1992).

Sorption plays a dominant role in determining the fate of uranium in freshwater systems. Sorption to clay minerals below pH 5, and iron and aluminium (oxy)hydroxides, silica and micro-organisms at higher pH, reduces the mobility of uranium in oxic waters (Prikryl et al. 1994, Waite et al. 1994, Kohler et al. 1996, Turner et al. 1996). Sorption of uranium to insoluble organic matter, or organic matter attached to particles also reduces the mobility of uranium (Pompe et al. 1996). It is generally established that sorption of uranium to particles increases with increasing pH until a threshold point is reached around pH 6–8 (Dzombak & Morel 1990, Choppin 1992, Willett & Bond 1995). The bioavailability and toxicity of sorbed uranium has not been studied. Sorption of uranium to particles and organic matter decreases with increasing salinity (van den Berg 1993).

No studies have reported the effect of salinity on the uptake and toxicity of uranium to estuarine and marine organisms.

Guideline

Both chronic and acute data were screened for acceptability, giving 14 chronic data points covering four taxonomic groups and around 40 acute data points covering four taxonomic groups (but excluding algae). Unfortunately the chronic data covered a wide range of pH and hardness values, particularly for algae, and it was necessary to use the acute data. Data were as follows (the wide ranges reflect different water conditions; pH range was 6.0 to 8.5):

Fish acute: seven species, 96-hour LC50, 1390 µg/L (Melanotaenia splendida splendida; Australian species) to 135,000 µg/L (P. promelas). Australian fish tested in tropical waters appeared more sensitive, as Mogurnda mogurnda had LC50 between 1570 to 3290 µg/L, lower than most other results.

Fish chronic: two species, 7-day NOEC (mortality) of 810 µg/L for both M. splendida splendida and M. mogurnda. The 14-day NOEC for M. mogurnda was 400 µg/L.

Crustacean acute: one species, D. magna, 48-hour EC50, immob, 5340 to 74,340 µg/L. The 48-hour acute LOEC for Moinodaphnia macleayi was 200 µg/L.

Crustacean chronic: two species, 10 µg/L (M. macleayi; 5-d NOEC, mortality, Australian data) to 200 µg/L (D. magna; 21-day LOEC, reproduction). Several repetitions of the M. macleayi chronic test with reproduction as an end-point have given similar low chronic figures (van Dam, pers. comm. 2000).

Hydra: one species, 48-hour LC50: H. viridissima, 150 µg/L, 48-hour LC50; 3-day LOEC, population growth, 250 µg/L (Australian data). The pH was 6.0–6.7 but, as it is typical of northern Australian tropical waters the figure was included.

Annelid: one species, Tubifex tubifex, 48 to 96-hour LC50, 2050 to 7890 µg/L (no pH figures were reported and the data could not be used).

Algae: one species, Chlorella vulgaris, 2000 µg/L, NOEC, population growth (these were on a wide pH range down to 2.2).

References

ANZECC & ARMCANZ 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality, Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand, Canberra.

Choppin GR 1992. The role of natural organics in radionuclide migration in natural aquifer systems. Radiochimica Acta 58/59, 113–120.

Choppin GR & Stout BE 1989. Actinide behaviour in natural waters. Science of the Total Environment 83, 203–216.

Clark DL, Hobart DE & Neu MP 1995. Actinide carbonate complexes and their importance in actinide environmental chemistry. Chemical Reviews 95, 25–48.

de Beer H & Coetzee PP 1992. Ion chromatographic separation and spectrophotometric determination of U(IV) and U(VI). Radiochimica Acta 57, 113–117.

Djogic R & Branica M 1993. Uranyl mixed-ligand complex formation in artificial seawater. Chemical Speciation and Bioavailability 5, 101–105.

Dzombak DA & Morel FMM 1990. Surface complexation modeling: Hydrous ferric oxide. John Wiley & Sons, New York.

Greene B, Henzl MT, Hosea JM & Darnall DW 1986. Elimination of bicarbonate interference in the binding of U(VI) in mill-waters to freeze-dried Chlorella vulgaris. Biotechnology and Bioengineering 28, 764–767.

Grenthe I, Fuger J, Konings RJM, Lemire RJ, Muller AB, Nguyen–Trung C & Wanner H 1992. Chemical thermodynamics of uranium. eds H Wanner & I Forest, North-Holland, Amsterdam.

Kohler M, Curtis GP, Kent DB & Davis JA 1996. Experimental investigation and modeling of uranium(VI) transport under variable chemical conditions. Water Resources Research 32, 3539–3551.

Langmuir D 1978. Uranium solution-mineral equilibria at low temperatures with applications to sedimentary ore deposits. Geochimica et Cosmochimica Acta 42, 547–569.

Livens FR, Morris K, Parkman R & Moyes L 1996. Actinide chemistry in the far field. Nuclear Energy 35, 331–337.

Mann DK & Wong GTF 1993. ‘Strongly bound’ uranium in marine waters: Occurrence and analytical problems. Marine Chemistry 42, 25–37.

Markich SJ, Brown PL & Jeffree RA 1996. The use of geochemical speciation modelling to predict the impact of uranium to freshwater biota. Radiochimica Acta 74, 321–326.

Meinrath G, Klenze R & Kim JI 1996. Direct spectroscopic speciation of uranium(VI) in carbonate solutions. Radiochimica Acta 74, 81–86.

Moulin C, Decambox P, Moulin V & Decaillon JG 1995. Uranium speciation in solution by time-resolved laser-induced fluorescence. Analytical Chemistry 67, 348–353.

Moulin V, Tits J & Ouzounian G 1992. Actinide speciation in the presence of humic substances in natural water conditions. Radiochimica Acta 58/59, 179–190.

Nakajima A, Horikoshi T & Sakaguchi T 1979. Ion effects on the uptake of uranium by Chlorella regularis. Agricultural Biology and Chemistry 43, 625–629.

NSW EPA 2000. Analytical Chemistry Section, Table of Trigger Values 20 March 2000, LD33/11, Lidcombe, NSW.

Osmond JK & Ivanovich M 1992. Uranium-series mobilisation and surface hydrology. In Uranium-series disequilibrium: Applications to earth, marine and environmental sciences, 2nd edn, eds M Ivanovich & RS Harmon, Clarendon Press, Oxford, 259–289.

Palmer DA & Nguyen-Trung C 1995. Aqueous uranyl complexes. 3. Potentiometric measurements of the hydrolysis of uranyl(VI) ion at 25oC. Journal of Solution Chemistry 24, 1281–1291.

Parkhurst BR, Elder RG, Meyer JS, Sanchez DA, Pennak RW & Waller WT 1984. An environmental hazard evaluation of uranium in a rocky mountains stream. Environmental Toxicology and Chemistry 3, 113–124.

Pompe S, Bubner M, Denecke MA, Reich T, Brachmann A, Geipel G, Nicolai R, Heise KH & Nitcsche H 1996. A comparison of natural humic acids with synthetic humic acid model substances: Characterisation and interaction with uranium (VI). Radiochimica Acta 74, 135–140.

Poston TM, Hanf RW & Simmons MA 1984. Toxicity of uranium to Daphnia magna. Water Air and Soil Pollution 22, 289–298.

Prikryl JD, Pabalan RT, Turner DR & Leslie BW 1994. Uranium sorption on a-alumina: Effects of pH and surface-area/solution-volume ratio. Radiochimica Acta 66/67, 291–296.

Tarzwell CM & Henderson C 1960. Toxicity of less common metals to fishes. Industrial Wastes 5, 12.

Turner GD, Zachara JM, McKinley JP & Smith SC 1996. Surface-charge properties and UO22+ adsorption of a subsurface smectite. Geochimica et Cosmochimica Acta 60, 3399-2414.

van den Berg CMG 1993. Complex formation and the chemistry of selected trace elements in estuaries. Estuaries 16, 512–520.

Waite TD, Davis JA, Payne TE, Waychunas GA & Xu N 1994. Uranium(VI) adsorption to ferrihydrite: Application of a surface complexation model. Geochimica et Cosmochimica Acta 58, 5465–5478.

Willett IR & Bond WJ 1995. Sorption of manganese, uranium, and radium by highly weathered soils. Journal of Environmental Quality 24, 834–845.