International Journal of Environmental Research University of Tehran ISSN: 1735-6865 EISSN: 2008-2304 Vol. 3, Num. 1, 2009, pp. 61-68

International Journal of Environmental Research, Vol. 3, No. 1, Winter, 2009, pp. 61-68

Uranium in Municipal Solid Waste Landfill Leachate

Øygard, J. K.^* and Gjengedal, E.

Department of Plant and Environmental Sciences, University of Life Sciences, N-1432 Aas, Norway
*Corresponding author E-mail:joar.oygard@umb.no

Received 10 Aug. 2007; Revised 25 July 2008; Accepted 15 Aug. 2008

Code Number: er09007

ABSTRACT

Landfill leachates and leachate sediments were investigated in order to survey the discharge of uranium (U) from municipal solid waste (MSW) landfills and municipal solid waste incinerator (MSWI) bottom ash landfills. Concentrations of U in the leachates were as high as or higher than concentrations of Cd, a metal more often discussed when considering the environmental effects of landfill leachate. However, the U concentrations in leachate were no greater than levels occurring naturally in water. The U concentrationsin leachate sediments were alsothe same as or lower than those reported in lake sediments around the world. Size charge fractionation of U in freshly sampled leachate showed that the metal was present mainly as either non-labile dissolved complexes or free anions, which have intermediate to high mobility and eco-toxicity. This is in contrast to other heavy metals, such as Cd, which are generally particulate bound in leachate. However, based onall the results of theinvestigations, it wasconcludedthat leachingofU is notof major concern in MSW and MSWI bottom ash landfills.

Key words: Landfill leachate, Heavy metals, Uranium, Sediments, Speciation

INTRODUCTION

One of the main concerns associated with the deposition of waste in sanitary landfills is that of the release of potentially hazardous landfill leachate to the environment. Landfill leachate contains varying amounts of both organic pollutants and toxic metals (Christensen et al., 2001; Kjeldsen et al., 2002; Baig et al., 2007). When assessing the pollution potential of the leachate, the elements commonly analysed are Pb, Cd, Cu, Zn, Cr and Ni (Kjeldsen et al., 2002; Baun and Christensen, 2004; Rieber et al., 2005). Even though these metals are usually considered the most important from an eco-toxicological perspective, other elements, such as Tl, Sb, Be and U, which could also pose a hazard, are often ignored. Therefore, although there is broad knowledge of the concentrations and toxicity of the former elements in landfill leachate (Baun and Christensen, 2004; Kjeldsen et al., 2002; Øygard et al., 2004; Øygard et al., 2007), little information is available about the latter elements.

The toxicity of U, calculated on a molar basis, has been found to be approximately equal or slightly lower than that of Cu, depending on the pH of the solution (Franklin et al., 2000), and higher or equal to the toxicity of Pb for aquatic organisms (Labrot et al., 1999). For naturally occurring U, the toxicity is determined by chemical properties of the metal, rather than its radioactivity. The toxic potential of U in landfill leachate is unknown, as few studies have reported concentrations of U in leachate, and none have investigated its physico-chemical form. The physico-chemical form of an element is a critical factor, influencing its mobility, bioavailability, and toxicity (Øygard et al., 2007). This study aims, firstly, to determine total concentrations of U in leachate from MSW related landfills, and, secondly, to determine the physico-chemical forms of any U present. As such, this investigation will show whether the presence of U in the discharge from waste landfills should be an issue of greater concern.

MATERIALS & METHODS

Fourteen solid waste landfills situated in western Norway (Fig. 1). were studied. All the landfills were in operation at the time of the sampling (Spring 2005). Landfills 1 and 5 received construction waste and municipal solid waste (MSW) incinerator bottom ash, landfill 11 received oil well sludge (after extraction of the oil) in addition to MSW, while the remaining landfills received only MSW. At landfill 3, the leachate was collected in a central collection point in a narrow valley downhill of the landfill. The other landfills had leachate collection systems, with leachate draining through pipes to a collection point, where the volume was continuously measured, and from where the samples were acquired. Not all the landfills had landfill liners. However, sampling of the groundwater at all sites had not detected any leaching from the landfills to the groundwater.

Samples for study by size charge fractionation were taken from the flowing leachate at the point immediately prior to its discharge into the environmental recipient, and fractionated in the field within 5 minutes. These samples were discussed previously by Øygard et al. (2007). Samples for total uranium determination were collected from the same locations. Leachate sediments were sampled either from the bottom of sedimentation tanks connected directly to the leachate discharge pipe, or from the bottom of the leachate volume sampling point.

Samples for determination of total U were acid preserved (to 1 % HNO₃) and stored for two weeks prior to analysis. Fractionated samples were also acid preserved to 1 % HNO_3. The sediment samples were dried at 40 ºC, sieved to 2 mm, ground and homogenized in an electrical agate mortar (Retsch Mortar Grinder KM 100), and kept in a desiccator prior to microwave assisted digestion (Anton Paar Multiwave microwave) and analysis.Total concentrations are of limited use when considering the eco-toxicological effects of metals, since they indicate neither the metal mobility nor the availability to aquatic organisms (Campbell, 1995; Batley et al., 2005). Free/labile metals have a more harmful effect on aquatic organisms than metals in non-labile colloids (Campbell, 1995; Batley et al., 2004). It is therefore important to determine the physicochemical form of the elements when assessing metal toxicity in landfill leachate. One method of doing this, used here, is size charge fractionation. This is a two-stage process. Firstly, filtration using a specific pore size (0.45 μm) was used to separate metals into two size classes (referred to as ‘particulate’ and ‘dissolved’ fractions). Ion exchange is subsequently used to fractionate the dissolved fraction into dissolved non-labile colloids / free anions and free cations/labile colloids. Data for the size charge fractionation of arsenic, cadmium, copper, chromium, iron, lead, manganese, molybdenum, nickel, and zinc in some of the leachate samples described in this work have been reported by Øygard et al., (2007).High pure nitric acid (Merck suprapur), and deionised water (> 18 MΩcm^-1) were used throughout. Total metal concentrations in the leachate were determined using a Thermo Finnigan Element 2 High Resolution ICP-MS. The certified reference material NRCC SLRS-4 was analysed along with the samples, and gave results within ±10 % for all available certified values. The metals in the leachate sediments were analyzed with a Perkin Elmer Sciex 6000 ICP-MS.

RESULTS & DISCUSSION

Total concentrations of uranium in the sampled leachates are shown in (Table 1). All the heavy metal concentrations are in the lower part of the concentration ranges reported by Christensen et al. (2001). All the leachate samples had pH values between 6.5 and 7.5. In many of the samples, the molar U concentrations were considerably higher than those of Cd. In a few samples, the U concentrations came close to those of Pb, but were always considerably lower than the concentrations of the other heavy metals (Cu, Zn, Ni and Cr).

In general, the concentrations of uranium found in natural waters range from 0.1 – 7 μg/L, whilst concentrations in seawater range from 2 – 4 μg/L (Langmuir, 1997). A Japanese study found average background concentrations of 28 ± 33 μg /L in river waters and 24 ± 30 μg /L in lakes (Tsamura and Yamasaki, 1992), whilst another large study found median concentrations of 20 μg /L in Russian lake water and 41 μg /L in Norwegian lakes (Reimann et al., 1999) Compared to these values, the U concentrations in the leachate samples are quite high, although well within the naturally occurring concentration ranges reported by Langmuir (1997). Background water concentrations depend upon the geology of the rock in contact with the water (Bakaç and Kumru, 2000). With the exception of L4, all the landfills discharged their leachate into the ocean. Since background levels of U in seawater are the same or higher than the concentrations found in the leachates, the environmental strain on the marine recipient due to U can be assumed to be insignificant.

U concentrations in natural waters depend upon the geology of the rock in contact with the water (Bakaç and Kumru, 2000). Some of the landfills were constructed without bottom liners and had an inwards directed groundwater flow, suggesting that U concentrations in the leachate could have been influenced by U from the groundwater. To investigate this, groundwater samples were taken from the vicinity of selected landfills, from sample points known not to be influenced by the landfills themselves or by any other sources of contamination. The concentrations of U are shown in (Table 2).

With the exception of landfill L8, the U concentrations in the leachates were considerably higher than concentrations in the corresponding local groundwater, indicating that the primary source of U in MSW landfill leachate generally is the deposited MSW. Only in landfill L8 could groundwater infiltration be contributing significantly to the U in the leachate.For some of the landfills, samples were taken directly from the flowing leachate, and immediately size charge fractionated in the field. These samples represent the leachate discharged into the environment, and the various U fractions thus represent the actual environmental load on the leachate recipients. The percentages of the total U concentration found in the various leachate fractions are shown in (Fig. 2).

Uranium was found to be present mainly as dissolved free anions or non-labile colloids (on average 81 ± 26 %), with only a small amount as free cations or labile colloids (7 ± 11 %). The proportion of U in the > 0.45 μm fraction was, on average, 12 ± 15 %. The U distribution in the leachate from landfill L3 deviates somewhat from the pattern shown by the other landfills. The reason for this is unknown to the authors. Waite and Payne (1993) reported that uranium is present mainly as an anionic hydroxides complex at neutral pH conditions. Alternatively, when carbonate is present in the solutions, uranium may form carbonate containing anionic complexes (Katsoyiannis, 2007). The uranyl ion and its complexes are considered to be the U species most bioavailabile and toxic to freshwater biota (Markich, 2002; Unsworth et al., 2005). Complexes with humic acids are readily formed, however (Markich, 2002; Unsworth et al., 2005). Unsworth et al., (2005) found that naturally occurring U in river water was largely (80 % of total U), bound to organic matter, and by computer simulation (WHAM) also predicted U to be bound to organic matter. The formation of organic complexes was proportional to the humic acid concentration in the water samples. It is thus likely that the dissolved non-labile U complexes found in this experiment consisted of U bound to organic matter. However, since landfill leachate generally has a high carbonate content, U could also be present as anionic carbonate complexes. Further work needs to be carried out in order to determine which of these is dominant.

The U concentrations in leachate sediments from some of the investigated landfills are presented in Table 3. Heavy metals, such as Pb, Cu, Cd, and Zn, are generally found to be present to a large degree as particulate or colloidal matter in leachate, and thus accumulate in leachate sediments (Øygard et al., 2007). In aquatic systems, sediments are known to act as sinks for uranium (Cadee, 1985; Klinkhammer and Palmer, 1991; Ribera et al., 1996). Naturally occurring Japanese lake sediments had a concentration range of 0.5 – 3.9 mg/kg (Nagao et al., 2002), while marine sediments (from the North Pacific Ocean) had a concentration range from 1.3 – 5.8 mg/kg (Klinkhammer and Palmer, 1991). The U concentrations in the landfill leachate sediments are in the lower part of these natural concentration ranges, suggesting that the leachate sediments do not pose a particular U pollution risk. Uranium occurs in the earth’s crust at concentrations of 34 mg/kg (Kabata and Kembias, 1984; Bosshard et al., 1992), and the sampled sediments are all below the average soil concentrations.

It can be seen from (Table 3) that Zn, Cu, Pb and Cd were present proportionally in much higher concentrations in the sediments compared to the leachates from the same landfills than were Ni, Cr, and U. Considering that U is present in leachate mainly in the < 0.45 μm fraction, it can be assumed that U will sediment out of the leachate more slowly than Zn, Cu, Pb and Cd, as these metals are present to a much greater degree as particulate matter > 0.45 μm (Øygard et al., 2007).

CONCLUSION

The U concentrations in sampled landfill leachates were low, and in accordance with naturally occurring water in other regions of the world. Sampling and analysis of groundwater at six of the landfills suggested that at only one of these landfills could groundwater infiltration potentially be a significant source of U. The U concentrations in leachate sediments were similar to those reported in lake sediments around the world. Size charge fractionation of U in freshly sampled leachate showed that U was present mainly as either non-labile dissolved complexes or free anions. Since these species could not be separated in this work, it is possible that U is present either as non-labile organic matter complexes, which have intermediate mobility and toxicity, or anionic carbonate complexes, which are highly mobile and ecotoxic. Based on the results of the investigations, it can be concluded that the presence of U is not a major concern in MSW and MSWI bottom ash landfills. Though it is likely that the U in the leachates is present as species which may be highly mobile and eco-toxic (worst case scenario), the concentrations are low both in the leachate and in the sediments, making the presence of the metal a minor environmental problem.

REFERENCES

• Bacaç, M., and Kumru, M.N. (2000). Uranium, radium and field measurements in the water of Gediz river. Turk. J. Eng. Env. Sci., 24, 229-236.
• Baig, A. S. A.; Rana, M. S.; Zaki, M. T. & Khan, K. (2007). Helminth Parasitic Load in Soil of Northern Lahore. Int. J. Environ. Res. 1, 3, 248-251.
• Batley, G.E., Apte, S.C. and Stauber, J.L. (2004). Speciation and bioavailability of trace metals in water: Progress since 1982. Aust. J. Chem., 57, 10, 903-919.
• Batley, G.E., Bowles, K.C., Apte, S.C. and Rogers, N.J. (2005). A rapid chelex column method for metal speciation in natural waters. 8th International Conference on the Biogeochemistry of Trace Elements (ICOBTE), Adelaide, Australia.
• Baun, D.L. and Christensen, T.H. (2004). Speciation of heavy metals in landfill leachate: a review. Waste Manage. Res. 22, 3-23.
• Bosshard, E., Zimmerli, B. and Schlatter, C. (1992). Uranium in the diet: risk assessment of its nephro-and radiotoxicity. Chemosphere, 24, 3, 309-321
• Cadee, G.C. (1984). Biological activity and sediments. (In: Sediments and pollution in waterways. General considerations. IAEA, Vienna, 111-126.)
• Campbell, P.G.C. (1995). (In: Metal Speciation and Bioavailability in Aquatic Systems.
• Tessier, A., Turner, D. R., Eds. John Wiley & Sons. Chichester, U.K., 45-102.)
• Christensen, T.H., Kjeldsten, P, Bjerge, P.L., Jensen, D.L.,Christensen, J.B., Baun,A.,Albrechtsen, H.J. and Heron, G., (2001). Biogeochemistry of landfill leachate plumes. Appl. Geochem., 16, 7-8, 659-718.
• Franklin, M.F., Stauber, J.L., Markich, S.J. and Lim, RP. (2000). pH-dependant toxicity of copper and uranium to a tropical freshwater alga (Chlorella sp.). Aquat. Toxicol., 48, 275-289.
• Kabata-Pendias, A. and Pendias, H. (1984). Trace elements in soil and plants. CRC press, London, 315.
• Katsoyiannis, I.A. (2007). Carbonate effects and pH-dependence of uranium sorption onto bacteriogenic iron oxides: kinetic and equilibrium studies. J. Haz. Mater., B139, 31-37.
• Klinkhammer, G.P. and Palmer, M.R. (1991). Uranium in the oceans: Where it goes and why. Geochim. Cosmochim. Ac., 55, 7, 1799-1806.
• Kjeldsen, P., Barlaz, M.A., Rooker, A.P., Baun, A., Ledin, A. and Christensen, T.H. (2002). Present and long-term compositionof MSWlandfill leachate: A review. Critical Rev. Env. Sci. Technol., 32, 4, 297-336.
• Labrot, F., Narbonne, J.F., Ville, P., Saint Denis, M. and Ribera, D. (1999). Acute toxicity, toxicokinetics, and tissue target of lead and uranium in the clam Corbicula fluminea and the worm Eisenia fetida: comparison with the fish Brachydanio reteri. Arch. Environ. Con. Tox., 36, 2, 167-178.
• Langmuir, D. (1997). Aqueous environmental geochemistry. Prentice Hall, Upper Saddle river, NJ, USA.
• Markich, S.J. (2002). Uranium speciation and bioavailability in aquatic systems: an overview. Scientific World J., 2, 707-729.
• Nagao, S.,Yanase, N.,Yamamoto, M., Kofuji, H., Sohrin, Y. and Amano, H. (2002). The geochemistryof uranium in pore waters from lake sediments. J. Radioanal. Nucl. Chem., 252, 2, 225-232.
• Reimann, C., Banks, D., Bogatyrev, I., de Caritat, P., Kashulina G. and Niskavaara, H. (1999). Lake water geochemistry on the western Kola Peninsula, northwest Russia. Appl. Geochem., 14, 787-805.
• Ribera, D., Labrot, F., Tisnerat, G. and Narbonne JF. (1996). Uranium in the environment: occurrence, transfer and biological effects. Rev. Environ. Contam. T., 146, 53-89.
• Rieber, C. Fredriksen, G.S. and Christensen, T.H. (2005). Heavy metal content of combustible municipal solid waste in Denmark. Waste Manage. Res. 23, 126-132.
• Tsamura, A. and Yamasaki, S. (1992). Direct determination of rare earth elements and actinoids in fresh water by double-focusing and high resolution ICP-MS. Radioisotopes, 41, 4, 185-192.
• Unsworth, E.R,. Jones, P., Cook, J.M. and Hill, S.J. (2005). Uranium speciation in moorland river water samples: A comparison of experimental results and computer model predictions. J. Env. Mon., 7, 6, 559-567.
• Waite, D.T. and Payne, T.E. (1993). Uranium transport in the sub-surface environment Koongarra – a case study. (In: Metals in groundwater. Eds. Allen, HE., Perdue, ME., Brown, DS. Lewis publishers, London.)
• Øygard, J.K., Gjengedal, E. and Røyset, O. (2007). Size charge fractionation of metals in municipal solid waste landfill leachate. Water Res., 41, 1, 47-54.
• Øygard, J.K., Måge, A. and Gjengedal, E. (2004). Estimation of the mass-balance of selected metals in four sanitary landfills in Western Norway, with emphasis on the heavy metal content of the deposited waste and the leachate, Water Res. 38, 2851-2858.

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