Scholarly article on topic 'Strontium sorption and precipitation behaviour during bioreduction in nitrate impacted sediments'

Strontium sorption and precipitation behaviour during bioreduction in nitrate impacted sediments Academic research paper on "Earth and related environmental sciences"

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{Strontium / Bioreduction / Incorporation / Sorption / Nitrate / Nuclear}

Abstract of research paper on Earth and related environmental sciences, author of scientific article — Clare L. Thorpe, Jonathan R. Lloyd, Gareth T.W. Law, Ian T. Burke, Samuel Shaw, et al.

Abstract The behaviour of strontium (Sr2+) during microbial reduction in nitrate impacted sediments was investigated in sediment microcosm experiments relevant to nuclear sites. Although Sr2+ is not expected to be influenced directly by redox state, bioreduction of nitrate caused reduced Sr2+ solubility due to an increase in pH during bioreduction and denitrification. Sr2+ removal was greatest in systems with the highest initial nitrate loading and consequently more alkaline conditions at the end of denitrification. After denitrification, a limited re-release of Sr2+ back into solution occurred coincident with the onset of metal (Mn(IV) and Fe(III)) reduction which caused minor pH changes in all microcosms with the exception of the bicarbonate buffered system with initial nitrate of 100mM and final pH>9. In this system ~95% of Sr2+ remained associated with the sediment throughout the progression of bioreduction. Analysis of this pH 9 system using X-ray absorption spectroscopy (XAS) and electron microscopy coupled to thermodynamic modelling showed that Sr2+ became partially incorporated within carbonate phases which were formed at higher pH. This is in contrast to all other systems where final pH was <9, here XAS analysis showed that outer sphere Sr2+ sorption predominated. These results provide novel insight into the likely environmental fate of the significant radioactive contaminant, 90Sr, during changes in sediment biogeochemistry induced by bioreduction in nitrate impacted nuclear contaminated environments.

Academic research paper on topic "Strontium sorption and precipitation behaviour during bioreduction in nitrate impacted sediments"


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Strontium sorption and precipitation behaviour during bioreduction in nitrate impacted sediments

Clare L. Thorpe a, Jonathan R. Lloyd a, Gareth T.W. Law a, Ian T. Burke b, Samuel Shaw b, Nicholas D. Bryan a, Katherine Morris a'*

a Research Centre for Radwaste and Decommissioning and Williamson Research Centre for Molecular Environmental Science, School of Earth, Atmospheric and Environmental Sciences, The University of Manchester, Manchester M13 9PL, UK

b Earth Surface Science Institute, School of Earth and Environment, The University of Leeds, Leeds LS2 9JT, UK



Article history: Received 20 December 2011 Received in revised form 1 March 2012 Accepted 2 March 2012 Available online 9 March 2012

Editor: J. Fein








The behaviour of strontium (Sr2+) during microbial reduction in nitrate impacted sediments was investigated in sediment microcosm experiments relevant to nuclear sites. Although Sr2+ is not expected to be influenced directly by redox state, bioreduction of nitrate caused reduced Sr2+ solubility due to an increase in pH during bioreduction and denitrification. Sr2+ removal was greatest in systems with the highest initial nitrate loading and consequently more alkaline conditions at the end of denitrification. After denitrification, a limited re-release of Sr2+ back into solution occurred coincident with the onset of metal (Mn(IV) and Fe(III)) reduction which caused minor pH changes in all microcosms with the exception of the bicarbonate buffered system with initial nitrate of 100 mM and final pH > 9. In this system ~95% of Sr2+ remained associated with the sediment throughout the progression of bioreduction. Analysis of this pH 9 system using X-ray absorption spectroscopy (XAS) and electron microscopy coupled to thermodynamic modelling showed that Sr2 + became partially incorporated within carbonate phases which were formed at higher pH. This is in contrast to all other systems where final pH was < 9, here XAS analysis showed that outer sphere Sr2+ sorption predominated. These results provide novel insight into the likely environmental fate of the significant radioactive contaminant, 90Sr, during changes in sediment biogeochemistry induced by bioreduction in nitrate

impacted nuclear contaminated environments.

© 2012 Elsevier B.V. All rights reserved.

1. Introduction

Strontium-90, a high yield fission product resulting from nuclear fuel cycle operations, is a significant radioactive contaminant at nuclear facilities worldwide (Jackson and Inch, 1989; Riley and Zachara, 1992; Mason et al., 2000; Dewiere et al., 2004; McKinley et al., 2007; Priest et al., 2008; McKenzie and Armstrong-Pope, 2010). The behaviour of 90Sr is of particular environmental concern in contaminated land due to both its ~29year half-life (meaning that it will persist over several hundred years), and it is relative mobility in the shallow sub-surface at some nuclear facilities (McKinley et al., 2007; McKenzie and Armstrong-Pope, 2010). The remediation of 90Sr and other radionuclides (e.g. U and Tc) from groundwaters at these sites is a key challenge for nuclear decommissioning. It is therefore important to explore remediation strategies which show promise of removing or immobilising a wide variety of problematic radionu-clides with differing biogeochemical behaviour (e.g. 90Sr, U and Tc) using a single methodology. Furthermore, the geochemical conditions

* Corresponding author. Tel.: +44 161 275 7541. E-mail address: (K. Morris).

0009-2541/$ - see front matter © 2012 Elsevier B.V. All rights reserved. doi:10.1016/j.chemgeo.2012.03.001

at many nuclear facilities are mildly acidic to neutral, have high nitrate and also have significant naturally occurring iron(lll) oxyhydroxide phases that must be taken into account when investigating remediation scenarios (Fredrickson et al., 2004; Istok et al., 2004; Begg et al., 2007; Edwards et al., 2007; McBeth et al., 2007; Law et al., 2010; McKenzie and Armstrong-Pope, 2010).

Strontium-90 exists in the natural environment solely as the Sr2 + ion, has very similar geochemical behaviour to Ca2 + and is therefore not directly affected by changes in redox conditions. Strontium specia-tion is controlled primarily by adsorption to mineral surfaces and incorporation or at high concentrations precipitation into Ca2 + bearing mineral phases (e.g. CaCO3). Strontium mobility in the subsurface is influenced by the adsorption capacity of the minerals within the sediment as well as the pH and ionic strength of the groundwaters, temperature, organic matter concentration and the exchangeable Ca2+/Mg2 + content (Cowan et al., 1991 ; Chen and Hayes, 1999; Solecki, 2005; Hull and Schafer, 2008; Chiang et al., 2010). The Sr2+ ion typically forms outer sphere adsorption complexes which are electrostatically bound to negatively charged mineral surfaces. As expected, increasing adsorption is observed as pH increases above the point of zero charge (PZC) of the relevant mineral phases (Cowan et al., 1991; Ferris et al., 2000; Sahai et al., 2000; Hofmann et al., 2005; Bascetin and Atun, 2006;

Bellenger and Staunton, 2008; Chorover et al., 2008). In iron rich sediments, adsorption to both aluminosilicate clays and Fe(III)-oxyhydroxide minerals will have a significant control over Sr2+ mobility with pH important in controlling the mineral surface charge and therefore the extent of cation adsorption (Chiang et al., 2010). Clay minerals (eg illite, chlorite, kaolinite and montmorillonite) provide important adsorption surfaces for cations even at low pH due to their relatively low PZC (pH 4-6) and permanent structural charge, (Hussain et al., 1996; Dyer et al., 2000; Coppin et al., 2002; Zhuang and Yu, 2002; Alvarez-Silva et al., 2010) whilst Fe(III)-hydroxides tend to contribute significant adsop-tion sites at higher pH (PZC pH 7-8) (Small et al., 1999; Hofmann et al., 2005). With increasing pH and alkalinity, groundwater will become oversaturated with regard to carbonate phases and at high Sr2+ concentrations, this may allow the precipitation of Sr2+ as strontianite (SrCO3) or at lower Sr2+ concentrations, the incorporation of Sr2 + into CaCO3 phases such as calcite or aragonite (Zachara et al., 1991; Tesoriero and Pankow, 1996; Greegor et al., 1997; Parkman et al., 1998; Finch et al., 2003; Fujita et al., 2004; Mitchell and Ferris, 2005). Additionally, where phosphate is present in the contaminated environment at significant concentrations, the sequestration of Sr2+ by phosphate minerals such as apatite is also likely to be a significant control on its behaviour (Handley-Sidhu et al., 2011).

Microbial metabolism has the ability to affect the geochemistry and mineralogy of subsurface sediments, as a result "bioreduction" systems have been considered for the remediation of groundwaters containing the redox active radionuclides Tc and U (Lloyd and Renshaw, 2005). Tc and U have been shown to be immobilised by reduction from the more soluble Tc(VII) and U(VI) to poorly soluble Tc(IV) and U(IV) during Fe(III) reducing conditions (Lloyd, 2003; Law et al., 2010, 2011). As radioactive 90Sr is often found as a co-contaminant in Tc/U contaminated land (Riley and Zachara, 1992; Hartman et al., 2007; McKenzie and Armstrong-Pope, 2010), understanding the behaviour of Sr2+ during bioreduction is essential in predicting and managing the mobility of this problematic contaminant in both natural and engineered bioreduction scenarios. During bioreduction the solution pH will be affected by the reaction products which include OH- and HCO—, and metal reduction will affect sediment mineralogy (Law et al., 2010; Thorpe et al., 2012). Reductive dissolution of bioavailable Fe(III)/Mn oxides and formation of new Fe(II) mineral phases may result in Sr2+ that was sorbed to Fe(III) oxide surfaces being released due to mineral dissolution (Langley et al., 2009a, b). However, it has been shown that during Fe(III) oxide crystallisation adsorbed contaminant metals (e.g. Pb2+) can become incorporated into the newly formed phase, therefore, the effect of Fe(III) oxide recrystallisation has the potential to increase or decrease Sr2+ environmental mobility. At the same time, the increase in pH caused by bioreduction processes may lead to enhanced removal of Sr2+ through increased sorption to mineral surfaces and carbonate precipitation/substitution (Roden et al., 2002; Mitchell and Ferris, 2005; Chorover et al., 2008). Microbial metabolism can result in the production of CO2-/HCO- which promotes alkaline pH conditions and supersaturation with regard to carbonate mineral phases (SrCO3 or CaCO3) in which Sr2+ can be precipitated (Coleman et al., 1993; Fujita et al., 2004; Mitchell and Ferris, 2005). These processes can also lead to siderite (Fe(II)CO3) formation during microbial reduction of Fe(III), which may result in minor Sr2+ becoming incorporated into the newly formed mineral phase (Parmar et al., 2000; Roden et al., 2002).

Here, we consider the effects ofmicrobial metabolism on the biogeo-chemistry and speciation of Sr2+ in conditions relevant to radioactively contaminated sites using stable Sr2+ as an analogue for 90Sr. Specifically, sub-surface nitrate concentrations are often elevated and have been reported in excess of 100 mM some nuclear facilities ( Riley and Zachara, 1992; Finneran et al., 2002; Fredrickson et al., 2004; Istok et al., 2004; Senko et al., 2005; McKenzie and Armstrong-Pope, 2010). In this study, we have examined the behaviour of Sr2+ during the

development of bioreducing conditions in sediments representative of the Sellafield nuclear facility that have been amended with between 0.3 and 100 mM nitrate. We tested the hypothesis that an increase in OH- and CO2-/HCO- during nitrate reduction may lead to increased adsorption of Sr2+ to mineral surfaces and, once over-saturation was reached, the precipitation and or incorporation of Sr2+ into carbonate phases at the high Sr/Ca ratio used in this study (1:1.5). Overall, our aim is to assess whether bioreduction approaches may be relevant to a range of problematic radionuclides including redox active U and Tc as well as 90Sr and thus provide a holistic remediation strategy where co-contamination of these radionuclides occurs.

2. Methods

2.1. Experimental section

2.1.1. Sample collection

Sediments representative of the Quaternary unconsolidated alluvial flood-plain deposits that underlie the UK Sellafield reprocessing site were collected from the Calder Valley, Cumbria, during December 2008 (Law et al., 2010). The sampling area was located ~2 km from the Sellafield site and sediments were extracted from the shallow sub-surface (Lat 54°26'30 N, Long 03°28'09W). Sediments were transferred directly into sterile containers, sealed, and stored at 4 °C prior to use.

2.1.2. Bioreduction microcosms

Sediment microcosms (10±0.1 g Sellafield sediment, 100±1 ml groundwater) were prepared using a synthetic groundwater representative of the Sellafield region (Wilkins et al., 2007) that was manipulated to produce a range of treatments (Table 1). Aerated systems were first established at variable pH (4.5, 5.5 and 7) to assess Sr2+ sorption in oxic systems. Following this a range of sealed microcosm systems were prepared. Unbuffered systems with an initial pH of ~5.5, and representative of the mildly acidic in situ pH at the sample site, were prepared with 0.3,10, and 25 mM nitrate amendments. Bicarbonate buffered systems with an initial pH of ~7 were prepared with 0.3, 10, 25, and 100 mM nitrate amendments. Sodium acetate was added as an electron donor in excess of available electron acceptors (10 mM for 0.3-10 mM nitrate treatments, 20 mM for 25 mM nitrate treatments and 70 mM for 100 mM nitrate treatments) and a deoxygenated NaNO3 solution was used for NO— amendment. Finally, Sr2+ (as stable SrCl2) was added to each microcosm to achieve 1.15 mM (100 ppm). Both the Sr concentration and the solid-solution ratio were chosen to allow full geochemical and spectroscopic characterisation and therefore are very much higher than values of 90Sr encountered at even the most contaminated sites (Riley and Zachara, 1992; McKenzie and Armstrong-Pope, 2010). It should be noted that 90Sr has a relatively

Table 1

Initial geochemical composition of microcosm systems.

System Initial Nitrate Acetate Ionic

pH (mM) (mM) strength

Bicarbonate unamended 5.5 0.3 10 0.024

0.3 mM nitrate

Bicarbonate unamended 5.5 10 10 0.034

10 mM nitrate

Bicarbonate unamended 5.5 25 20 0.059

25 mM nitrate

Bicarbonate amended 7 0.3 10 0.027

0.3 mM nitrate

Bicarbonate amended 7 10 10 0.037

10 mM nitrate

Bicarbonate amended 7 25 20 0.062

25 mM nitrate

Bicarbonate amended 7 100 70 0.190

100 mM nitrate

short half-life and thus high specific activity and even for the most impacted sites where groundwater concentrations of >1000Bql—1 have been reported the molar concentration of 90Sr will be very low (< 10—11 mol l-1) compared to our experimental concentrations (Riley and Zachara, 1992; McKinley et al., 2007; Mckenzie and Armstrong-Pope, 2010). Triplicate microcosms were then sealed with buytl rubber stoppers and incubated anaerobically at 20 °C in the dark for 110-250 days. At appropriate time points, sediment slurry was extracted under an O2 free Ar atmosphere using aseptic technique and centrifuged (15,000 g; 10min) to provide wet sediment pellets and porewater samples for analysis ofbioreduction products and strontium.

2.1.3. Geochemical analyses and imaging

During microcosm sampling, total dissolved Fe, Mn(II), and NO2 concentrations were measured with standard UV-vis spectroscopy methods on a Jenway 6715 UV-vis spectrophotometer (Goto et al., 1977; Viollier et al., 2000; Harris and Mortimer, 2002). Aqueous NO-, SO4—, HCO-/COf — and acetate were measured by ion chromatography (Dionex 4000i liquid chromatography). Aqueous Sr2+ and Ca2+ were measured by ICP-AES (Perkin-Elmer Optima 5300). Total bioavailable Fe(III) and the proportion of extractable Fe(II) in the sediment was estimated by digestion of ~0.1 g of sediment in 5 ml of 0.5 N HCl for 60 min followed by the ferrozine assay (Stookey, 1970; Lovley and Phillips, 1986). The pH and Eh were measured with a pH/Eh metre (Denver Instruments, UB10) and probes calibrated to pH 4,7 and 10. Standards were routinely used to check the reliability of all methods and calibration regressions typically had R2 > 0.99. The elemental composition and bulk mineralogy of the sediment were determined by X-ray fluorescence (Thermo ARL 9400 XRF) and X-ray diffraction (Philips PW 1050 XRD). Selected end point samples were imaged using Environmental Scanning Electron Microscope (ESEM) in combination with Backscattering Electron Detection (BSE) and Energy Dispersive X-ray Analysis (EDAX) (Philips XL30 ESEM-FG).

2.1.4. X-ray absorption spectroscopy

Selected samples from the bicarbonate buffered pH 7 systems with 10,25 and 100 mM nitrate amendments were chosen to examine Sr2 + speciation in: (1) oxic sterile control pH 7 sediment; (2) Fe(III)/SO|— reducing end point pH 7.2 sediments, (3) Fe(III)/SO|— reducing end point pH 8.1 sediments; and (4) Fe(III) reducing end point pH 9.3 sediments. Typical concentrations of Sr2+ in these samples were in the range 600-1000 ppm. Standards: (1) SrCl2 (aq), 3000 ppm (Fisher Scientific), (2) SrCO3 (s) (Fisher Scientific) and (3) natural Sr2+ substituted aragonite from crushed aragonite mineral sample (Sr2+ concentration-1000 ppm), were prepared and diluted with boron nitride where necessary. Samples were transferred to XAS cells under anaerobic conditions, cooled to — 80 K with a liquid nitrogen cryostat (see Nikitenko et al., 2008), and Sr K-edge XAS spectra were collected on beamline BM26A at the European Synchrotron Radiation Facility (ESRF). For sediment samples, Sr K-edge spectra (16115.26 keV) were collected in fluorescence mode using a 9 element solid state Ge detector. Multiple scans were averaged in Athena version 0.8.061 (Ravel and Newville, 2005) and normalised XANES data plotted. Background subtraction for EXAFS analysis was performed using PySpline v1.1 (Tenderholt et al., 2007). EXAFS data were fitted using DLexcurv v1.0 (Tomic et al., 2005) using full curve wave theory (Gurman et al., 1984) by defining a theoretical model which was informed by the relevant literature (e.g. O'Day et al., 2000; Finch et al., 2003) and comparing the model to the experimental data. Shells of backscatterers were added around the Sr2+ and by refining an energy correction Ef (the Fermi Energy; which for final fits typically varied between — 3.8 and — 2.6), the absorber-scatterer distance, and the Debye-Waller factor for each shell. Model iterations were performed until a least squares residual was minimised. Shells were only included in the model fit if the overall least square residual (the R-factor; Binsted et al., 1992) was improved by >5%.

3. Results and discussion

3.1. Sediment characteristics

Sediment composition was measured by X-ray fluorescence and was found to comprise Si (31.57%), Al (7.63%), Fe (3.64%), K (2.79%), Na (0.99%), C (0.96%), Mn (0.87%), Ti (0.45%), Ca (0.23%) and P, S and Cl (<0.1%). We note that XRF analyses show that phosphate is present in our systems at very low concentrations (<0.008%) and is not likely to be a significant control on Sr2+ behaviour in this system.

Trace metal analysis showed natural Sr2+ to be present in sediments at 62.8 ±0.2 ppm and natural aqueous Sr2+ was < 1 ppm. Strontium was added to groundwater media in significant excess to the natural background at 100 ppm (1.15 mM Sr2+) and Ca2+ was present at 67 ppm (1.67 mM) thus a Sr/Ca ratio of-1:1.5 was present in the synthetic groundwater media. The concentration of 0.5 N HCl extractable Fe(III) in the sediment was 5.6 ±0.5 mmol kg—1 prior to incubation and the sediment pH was -5.5.

3.2. Sorption to oxic sediment

In sterile control microcosms, increased Sr2+ sorption was observed in microcosms with a high pH and a low ionic strength. For example, for a constant ionic strength system (I = 0.027 mol dm— 3) run at pH 4.5, 5.5 and 7, the Sr2+ removal from solution was 35.6 ± 1.9%, 47.4 ±5.3% and 63.2 ±2.1% respectively (equating to Kd values of 5.5, 9.0 and 17.1 ml g— 1). Differences in strontium behaviour in the sterile microcosms were attributed to pH dependent differences in sorption to mineral surfaces present in the sediment. Sorption to both clays and Fe(III)-oxyhydroxide surfaces is possible although clay minerals is likely to predominate in unbuffered microcosms in which the pH of 4.5-5.5 is above the PZC for many clay minerals (Coppin et al., 2002; Zhuang and Yu, 2002; Alvarez-Silva et al., 2010) whilst Fe-oxyhydroxides become more significant as pH approaches their PZC at pH ~7 (Dyer et al., 2000; Hofmann et al., 2005). Additionally, in control experiments at pH 7 and with increasing ionic strength (0.027, 0.037, 0.062 and 0.190 mol dm— 3) resulting from sodium nitrate and sodium acetate additions, Sr2+ sorption was 68, 65, 55 and 28% respectively (equating to Kd values of 21.2, 18.5, 12.2 and 3.8ml/g), presumably reflecting increased competition for Sr2+ sorption sites at higher ionic strengths due to cation exchange processes (Hull and Schafer, 2008).

3.3. Biogeochemistry in sediment microcosms

The unbuffered (initial pH 5.5; nitrate range 0.3-25 mM) and bicarbonate buffered (initial pH 7.0; nitrate range 0.3-100 mM) experiments all underwent progressive anoxia and electron acceptors were utilised in the order NO— >NO— >Mn/Fe(III)>SO|— (Figs. 1 and 2). Microbially mediated nitrate reduction caused a decrease in pore-water nitrate and transient accumulation of nitrite in all systems. The onset of Fe(III) reduction was indicated by an increase in sediment extractable Fe(II) and this was then followed by a decrease in porewater SO4 — indicating sulfate reduction. No geochemical changes were observed in sterile control microcosms. In unbuffered systems (initial pH 5.5), as expected, microbial activity was inhibited at low pH and terminal electron accepting processes proceeded more slowly than in the parallel bicarbonate buffered microcosms (initial pH 7.0) (Figs. 1 and 2) (e.g. Law et al., 2010; Thorpe etal., 2012). However, in the unbuffered systems, nitrate reduction led to the release of OH— and HCO—, amending the pH such that a pH increase from 5.5 to 6.8, 7.5 and 8.3 occurred during the reduction of 0.3,10 and 25 mM nitrate respectively (Fig. 1; Table 2; Thorpe et al., 2012). Metal reduction commenced once nitrate reduction had occurred and mid-point 0.5 N HCl extractable Fe(III) reduction was observed at approximately 25, 35 and 45 days for systems with 0.3,10 and 25 mM nitrate (Fig. 1). Nitrate reduction and the associated pH increase in all microcosms coincided with removal of Sr2 +

Fig. 1. Unbuffered, microcosm incubation time-series data (days 0 —160). (A) pH (B) NO—, (C) 0.5 N HC1 % extractable sedimentary Fe as Fe(II), (D) SO:!-, (E) porewater Sr and (F) porewater Ca. • = 0.3 mM nitrate system; A = 10 mM nitrate system; ■ = 25 mM nitrate system. Initial pH in all microcosms was ~5.5. Error bars represent 1ct experimental uncertainty from triplicate microcosm experiments (where not visible error bars are within symbol size).

and Ca2+ from solution (Fig. 1). Interestingly, as bioreduction progressed through Fe(III) and SO4 - reduction in these dynamic systems, a small amount of both Sr2+ and Ca2+ (< 10% of that sorbed after nitrate reduction) was remobilised to solution. This re-release coincided with a slight decrease in pH (<0.5 pH units) presumably due to re-equilibration of the microcosm system following nitrate reduction. The re-release of sorbed Sr2+ and Ca2+ may be due solely to pH dependent sorption/desorption to mineral surfaces or in some systems (for example above pH 7) there may be release of Sr2+ and Ca2+ sorbed to Fe(III)-oxyhydroxides as reductive dissolution of the Fe(III) phases occurred (Small et al., 1999; Roden et al., 2002; Langley et al., 2009a, b).

In bicarbonate buffered systems with an initial pH of 7, the final pH following the reduction of 0.3, 10, 25 and 100 mM nitrate was 7.5, 8.0, 8.5 and 9.4 (Fig. 2: Table 2). Terminal electron accepting processes proceeded faster than in unbuffered microcosms with midpoint 0.5 N HCl extractable Fe(III) reduction occurring at <20 days for 0.3 and 10 mM nitrate systems and around 40 and 160 days for systems with 25 and 100 mM nitrate (Fig. 2). As with the unbuffered systems, Sr2+ and Ca2+ were removed from solution during nitrate reduction with increasing pH and a small amount (< 10% of that sorbed after nitrate reduction) of Sr2+ and Ca2+ was re-released into solution in all systems apart from the bicarbonate buffered,

Fig. 2. Buffered, microcosm incubation time-series data (days 0-160/260). (A) pH, (B) NO-, (C) 0.5 N HCl % extractable sedimentary Fe as Fe(II), (D) SO4(E) porewater Sr and (F) porewater Ca. • = bicarbonate buffered 0.3 mM nitrate system; A = bicarbonate buffered 10 mM nitrate system; ■ = bicarbonate buffered 25 mM nitrate system; ◊ = bicarbonate buffered100 mM nitrate system. Initial pH in all microcosms was -7. Error bars represent 1ct experimental uncertainty from triplicate microcosm experiments (where not visible error bars are within symbol size).

Table 2

Percentage strontium sorbed to sediments during nitrate and metal reduction compared to an oxic control.

System Sterile oxic 90% nitrate reduction 90% Fe(III)/SO4 reduction End point

% Sr2 + on sedimenta pH % Sr2 + on sediment pH % Sr on sediment pH Net decrease (-) or increase (+) of Sr2 + on sediments

Bicarbonate unamended 0.3 mM nitrate 47.0 5.5 54 ±2.2 6.6 50 ±2.3 6.7 +-2%

Bicarbonate unamended 10 mM nitrate 47.3 5.5 75 ±1.8 8.0 55 ±3.6 7.5 +-9%

Bicarbonate unamended 25 mM nitrate 45.5 5.5 82 ±2.1 8.2 60 ±0.5 7.8 +-18%

Bicarbonate amended 0.3 mM nitrate 67.9 7.0 63 ±1.4 7.0 50 ±1.8 7.0 —16%

Bicarbonate amended 10 mM nitrate 64.2 7.0 78 ±0.7 8.0 57 ±0.3 7.5 --6%

Bicarbonate unamended 25 mM nitrate 54.8 7.0 84 ±0.4 8.5 62 ±0.7 8.1 +-7%

Bicarbonate unamended 100 mM nitrate 32.6 7.0 93 ±1.4 9.3 94 ±0.8 9.3 +-61%

a Differences in Sr2+ sorption to sterile controls occur due to varying pH and ionic strength due to the addition of NaHCO3, Na-acetate and NaNO3.

100 mM nitrate system (Fig. 2). Here, the final pH was 9.3 and interestingly, Sr2+ remained associated with the sediment throughout Fe(lll) and sulfate reduction. In this high nitrate loaded system, the utilisation of 70 mM acetate resulted in the accumulation of 207 ± 4.9 mM of dissolved inorganic carbon and amended the pH to alkaline conditions.

For comparison with other studies it is useful to examine distribution coefficients for Sr2+ (Kd = (solid in gg- 1/aqueous in gml-1)). Distribution coefficients are only relevant to the specific geochemical conditions of each system of study and give an indication of the extent of Sr2+ partitioning onto the solid phase in different systems. Here Kd values ranged from < 10mlg-1 in systems with a high ionic strength (0.190 moldm-3) or a low pH (5.5) and increased to >50mlg-1 with increasing pH. The distribution coefficient in systems with a final pH>9 was calculated to be 133 ml g-1. These Kd values compare well with literature values where surface sediment Kd values are typically between 10 and 200 ml g-1 (Deldebbio, 1991; Liszewski et al., 1998; Fernandez et al., 2006) and deeper more quartz rich sediments have Kd values of < 10 ml g-1 (Stephens et al., 1998; Dewiere et al., 2004).

Modelling of the solution chemistry in bioreduced systems (PHREEQC-2: LLNL database) suggested that for all unbuffered biore-duced system end-points, Fe(ll)CO3 and SrCO3 were oversaturated in all the different nitrate amendments whilst the CaCO3 phases were undersaturated in the bioreduced 0.3 mM nitrate amended systems and oversaturated in all other treatments (Table 3). As expected for these carbonate phases, the degree of oversaturation increased as alkalinity increased. In the bicarbonate amended bioreduced system end points, all nitrate amendments showed oversaturation of Fe(ll) CO3, CaCO3, and SrCO3 and with increasing oversaturation with increasing alkalinity (Table 3). Clearly, although unable to resolve the detail of the dynamic bioreduction experiments, these modelling data suggest an increased tendency to oversaturation of carbonate mineral phases with increased nitrate reduction and microbially produced alkalinity.

ESEM (Environmental Scanning Electron Microcoscopy) was used to assess the distribution of Sr2+ in end point pH 7 (bicarbonate buffered 10 mM nitrate) and pH 9.3 (bicarbonate buffered 100 mM nitrate) sediments (Fig. 3). In backscattering mode image brightness is related to the average atomic mass present (Z contrast). In the pH 9.3 system, secondary electron and backscatter images in combination with EDAX analysis show a number localised bright spots of ~20 |jm diameter enriched in Sr2+ (Fig. 3) whilst there were no observed localised bright spots in the pH 7 system. Semi-quantitive analysis of EDAX spectra of the localised bright spots showed a significant concentration of Ca2+ and Sr2+ in agreement with predicted SrCO3 and CaCO3 oversaturation.

In order to further understand Sr2+ speciation during bioreduc-tion in these complex systems, samples from an oxic pH 7 control with Sr2+ sorption, and selected bicarbonate buffered, nitrate amended (0.3, 25 and 100 mM) bioreduced end points with a final pH of 7.2, 8.1 and 9.3 were analysed using X-ray absorption spectros-copy. XANES spectra for all samples show a single peak indicative of 9

fold coordination and there was no evidence for 6 fold coordination (as in the calcite standard which has a clear doublet). Thus our experiments show no evidence for Sr2+ substituted calcite formation (Fig. 4; Parkman et al., 1998). XANES spectra for the oxic, bioreduced pH 7.2, and bioreduced pH 8.1 samples all comprised a single peak and compared well with a SrCl2 aqueous standard, implying that in these sediments, after bioreduction, Sr2+ was present primarily as adsorbed Sr2+ (Fig. 4). By contrast, the XANES spectra for the pH 9.3 bioreduced sample showed some evidence for peak flattening and thus some similarity to the model Sr-carbonate phases (e.g. strontianite and Sr-substituted aragonite) (Fig. 4). Modelling of the EXAFS spectra for the oxic, biore-duced pH 7.2 and bioreduced pH 8.1 samples showed an approximate 9-fold coordination environment at 2.60 A, indicative of outer sphere Sr2+ adsorption to mineral surfaces (Parkman et al., 1998; Chen and Hayes, 1999; Carroll et al., 2008; Fig. 5; Table 4). EXAFS for the pH 9.3 bioreduced sample could also be modelled with 9 fold "outer sphere" co-ordination. However, EXAFS model fits for this system were significantly improved by addition of shells of carbon and strontium backscat-ters at 3.03, 4.18 and 4.87 A (Table 4; Fig. 5) respectively. This clearly indicates a contribution from an additional Sr2+ species in this spectrum with bond distances indicative of SrCO3 (Parkman et al., 1998;

Table 3

Saturation index for key carbonate minerals in microcosm systems. Modelled using PHREEQC-2 (Lawrence Livermore National Laboratory database— llnl.dat).

Saturation index (PHREEQC-2)a

Sr2+ Final Siderite Calcite Aragonite Strontianite _(ppm) pH

Oxic sediment 100 5.5

Bioreduced 100


Unbuffered 100 6.7

0.3 mM nitrate

Unbuffered 100 7.5

10 mM nitrate

Unbuffered 100 7.8

25 mM nitrate

Bicarbonate buffered 100 7.0

0.3 mM nitrate

Bicarbonate buffered 100 7.2

10 mM nitrate

Bicarbonate buffered 100 8.1

25 mM nitrate

Bicarbonate buffered 100 9.3

100 mM nitrate

Bicarbonate buffered 10 9.3

100 mM nitrate

Bicarbonate buffered 1 9.3

100 mM nitrate

Bicarbonate buffered 0.1 9.3

100 mM nitrate

Bicarbonate buffered 0.01 9.3 100 mM nitrate

a Temperature 21 °C, concentration of ions in solution from Table 1, pH as measured.

1.97 2.76 3.07 2.31 2.49 3.28 3.25 3.25 3.25 3.25 3.25

- 3.52

- 0.17 0.70 1.18 0.22 0.42 1.48 2.36 2.36 2.36 2.36 2.36

- 3.67

- 0.32 0.55 1.03 0.08 0.28 1.33 2.21 2.21 2.21 2.21 2.21

- 2.87

0.47 1.35 1.83 0.86 1.07

2.14 3.65 2.25

1.15 0.25 -0.75

0 2 4 6 8 10 12 14 16 18 20 0 2 4 6 8 10 12 14 16 18 20

keV keV

Fig. 3. ESEM images of the Fe(lll) reducing bicarbonate buffered 100 mM nitrate sample at final pH 9 containing Sr-and Ca-rich crystalline structures and corresponding EDAX spectra. Images show: (A) ESEM backscatted detection mode image of sediment indicating heavier elements (Sr) as bright patches in the field of view 300 |jm; (B) Secondary electron image showing the structure of Sr/Ca rich area at a field of view 3 |m; (C) the energy dispersive electron analysis (EDAX) spectra for the entire sample; and (D) a spot EDAX analysis on the Sr/Ca rich structure.

O'Day et al., 2000). Further analysis showed that a model EXAFS fit was possible with additional shells of 2.6 carbon atoms at 3.03 A, 2.8 strontium atoms at 4.18 A and 2.5 strontium atoms at 4.87 A; these values

Sr K edge

. SrCl2 (Sr7*)

Oxic sediment

Bioreduced pH 7.2

Bioreduced pH 8.1 Bioreduced pH 9-3 Strontianite (SrCO,) Sr substituted aragonite (CaC03)

Sr substituted calcite (CaCOJ


16100 16120 16140 16160 Energy (eV)

Fig. 4. Normalised Sr K-edge XANES spectra for selected standards and microcosm systems. From top to bottom: SrCl2 aqueous standard, oxic sediment sample, bioreduced pH 7.2 sample, bioreduced pH 8.1 sample, bioreduced pH 9.3 sample, Sr substituted aragonite standard and strontianite standard.

are approximately 50% of what is expected for pure SrCO3, which is consistent with a model where approximately half of Sr2+ is present in a SrCO3 like environment (Table 4; Fig. 5). Indeed, this model, which is geochemically sensible, resulted in a better fit to the spectrum and a 27% reduction in the least square residual when compared to the data modelled as 100% adsorbed Sr2+ suggesting that both adsorption and incorporation occurred in this system (Table 4).

In natural and engineered environments concentrations of Sr2 + and 90Sr are generally much lower than in these experiments (eg 0.1 ppm natural Sellafield groundwater) (Wilson, 1996). Under Sellafield conditions, model simulations predicted that bioreduced system even at pH 9 would be undersaturated with regard to SrCO3 below -0.1 ppm strontium (Table 3); nonetheless, at a pH>7, systems would remain supersaturated with respect to CaCO3. It is therefore feasible that substitution of Sr2+ into CaCO3 rather than precipitation as SrCO3 will be important in controlling the mobility of both natural Sr2+ and artificial 90Sr in such systems. Indeed, it is well documented that Sr2+ can substitute for Ca2+ within the calcium carbonate lattice (Pingitore et al., 1992; Tesoriero and Pankow, 1996; Greegor et al., 1997; Warren et al., 2001; Finch et al., 2003). Recent studies, focused on bacterial urolysis, have found that Sr2 + incorporation into the CaCO3 lattice was enhanced by the rapid precipitation rates resulting from HCO- production and the pH rise associated with microbial respiration (Fujita et al., 2004; Mitchell and Ferris, 2005). Both a pH rise and dissolved organic carbon production were observed during bioreduction by indigenous microorganisms in this study suggesting that nitrate reduction might also result in enhanced Sr2+ uptake into calcite compared to those observed under slower precipitation rates.

Fig. 5. Experimental (solid) and theoretical best fit (dashed) EXAFS spectra and corresponding Fourier transforms obtained for (from top to bottom): SrCl2 aqueous standard, oxic sediment sample, bioreduced pH 7.2 sample, bioreduced pH 8.1 sample, bioreduced pH 9.3 sample, Sr substituted aragonite standard and strontianite standard. Solid lines are the data and dashed lines are the fits to the data.

3.4. Summary and environmental relevance

Overall, our experiments showed that there is increased Sr2+ removal from solution during bioreduction in nitrate impacted sediments compared to sterile control systems. In systems with an initially low pH (5.5), removal of Sr2+ from solution after bioreduc-tion was particularly enhanced, presumably due to the increased sorption onto deprotonated mineral surfaces as the pH increased above 6. After nitrate reduction, system re-equilibration and an associated decrease (<0.5 pH units) in pH resulted in modest (< 10%) rerelease of Sr2+ into solution highlighting the vulnerability of adsorbed Sr2+ to re-release due to changing geochemical conditions.

Table 4

Parameters obtained from EXAFS data fitting of Sr K-edge spectra from Sr2 with sediment at various sediment conditions.


Sample Shell no Bond C.N. R(A) 2CT2 (A2) R-factor

SrCl2 1 Sr- O 8.75 2.60 0.029 23.0

Oxic sample pH 7 1 Sr- O 8.83 2.61 0.019 22.1

Bioreduced pH 7.2 1 Sr- O 8.68 2.60 0.020 18.0

Bioreduced pH 8.1 1 Sr- O 8.89 2.61 0.021 19.6

Bioreduced pH 9.3 (1) 1 Sr- O 8.03 2.60 0.024 27.9

Bioreduced pH 9.3 (2) 1 Sr- O 8.19 2.61 0.021 20.2

2 Sr- C 2.68 3.03 0.015

3 Sr- Sr 2.78 4.18 0.029

4 Sr- Sr 2.47 4.88 0.024

Strontianite 1 Sr- O 9a 2.64 0.027 20.3

2 Sr- C 6a 3.04 0.032

3 Sr- Sr 6a 4.22 0.029

4 Sr- Sr 4a 4.97 0.033

Sr substituted aragonite 1 Sr- O 9a 2.59 0.015 29.1

2 Sr- C 6a 2.98 0.031

3 Sr- Ca 6a 4.02 0.020

4 Sr- Ca 4a 4.87 0.013

N is the occupancy (±~25%), R(A) (is the interatomic distance (±~0.02 A), 2a2 is the Debye-Waller factor (A2) and R (least squares residual) is a measure of the overall goodness of fit. a Fixed.

In extreme environments with very high (100 mM) nitrate concentrations, bioreduction led to a final pH of > 9 and enhanced removal of Sr2+ from solution occurred throughout the bioreduction cascade. This study has shown that in very high nitrate systems an increase in pH and dissolved inorganic carbon associated with microbial reduction and particularly denitrification can promote the precipitation and incorporation of Sr2+ into carbonate phases although the engineering aspects of this process are as yet unexplored. Clearly, radio-strontium incorporation into carbonate phases is desirable in remediation scenarios as they are redox insensitive phases and are potentially more resistant to remobilization than sorbed Sr2+. It is also clear that bioreduction scenarios have the potential to impact Sr2+ mobility in the subsurface and that understanding the bioreduction behaviour of redox inactive radioactive contaminants can be of significance in assessing the efficacy of bioreduction schemes at nuclear facilities. We further suggest that under constrained conditions, bioreduction may have the potential to co-treat redox active radionuclides and 90Sr increasing the range of applications for this clean-up technology across the global nuclear waste legacy.


We thank Jon Fellowes, Alastair Bewsher and Paul Lythgoe for help in data acquisition. We also thank the ESRF beamline (B-26) scientists and Sarah Wallace (University of Leeds) for their help in XAS analysis. This work was supported by a studentship to CLT from the Engineering and Physical Science Research Council (EPSRC) as part of the Decommissioning, Immobilisation and Management of Nuclear Waste for Disposal (DIAMOND) consortium: grant EP/F055412/1. We also acknowledge the support of the Natural Environment Research Council grant NE/H007768/1.


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