Escherichia Coli Response to Uranyl Exposure at Low Ph and Associated Protein Regulations

Escherichia coli Response to Uranyl Exposure at Low pH and Associated Protein Regulations

Arbia Khemiri1,2,3, Marie Carrie`re4, Nicolas Bremond1,2,3, Mohamed Amine Ben Mlouka5, Laurent Coquet5, Isabelle Llorens6,7, Virginie Chapon1,2,3, Thierry Jouenne5, Pascal Cosette5*, Catherine Berthomieu1,2,3* 1 CEA, DSV, IBEB, Commissariat a` l’Energie Atomique, Laboratoire des Interactions Proteíne-Me´tal, Saint-Paul-lez-Durance, France, 2 CNRS, UMR Biologie Ve´ge´tale et Microbiologie Environnementales 7265, Saint-Paul-lez-Durance, France, 3 Universite´ d’Aix-Marseille, Saint-Paul-lez-Durance, France, 4 UMR E3 CEA-Universite´ Joseph Fourier, Service de Chimie Inorganique et Biologique, Laboratoire Le´sions des Acides Nucleíques (LAN), Grenoble, France, 5 UMR 6270 CNRS, Plateforme Proteómique PISSARO, IRIB -Universite´ de Rouen, Mont Saint Aignan, France, 6 ESRF-CRG-FAME beamline, Polygone Scientifique Louis Neél, Grenoble, France, 7 Commissariat a` l’Energie Atomique CEA, DSM, INAC, Laboratoire Nanostructure et Rayonnement Synchrotron, Grenoble, France

Abstract Better understanding of uranyl toxicity in bacteria is necessary to optimize strains for bioremediation purposes or for using bacteria as biodetectors for bioavailable uranyl. In this study, after different steps of optimization, Escherichia colicells were exposed to uranyl at low pH to minimize uranyl precipitation and to increase its bioavailability. Bacteria were adapted to mid acidic pH before exposure to 50 or 80 mM uranyl acetate for two hours at pH<3. To evaluate the impact of uranium, growth in these conditions were compared and the same rates of cells survival were observed in control and uranyl exposed cultures. Additionally, this impact was analyzedby two-dimensional differential gel electrophoresis proteomics to discover protein actors specifically present or accumulated in contact with uranium.Exposure to uranium resulted in differential accumulation of proteins associated with oxidative stress and in the accumulation of the NADH/quinone oxidoreductase WrbA. This FMN dependent protein performs obligate two-electron reduction of quinones, and may be involved in cells response to oxidative stress. Interestingly, this WrbA protein presents similarities with the chromate reductase from E. coli, which was shown to reduce uranyl in vitro.

Citation: Khemiri A, Carrie`re M, Bremond N, Ben Mlouka MA, Coquet L, et al. (2014) Escherichia coli Response to Uranyl Exposure at Low pH and Associated Protein Regulations. PLoS ONE 9(2): e89863. doi:10.1371/journal.pone.0089863 Editor: Christophe Beloin, Institut Pasteur, France Received September 19, 2013; Accepted January 23, 2014; Published February 26, 2014 Copyright: ß 2014 Khemiri et al. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited. Funding: This work was funded in part by the CNRS program EC2CO (Cytrix and MicrobiEn), and the ‘‘Toxicologie’’ program of the Commissariat a` l’Energie Atomique CEA. The funders had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript. Competing Interests: The authors have declared that no competing interests exist. * E-mail: [email protected] (CB); [email protected] (PC)

Introduction Interactions between uranium and bacteria have been extensively studied to identify the potential role of soil bacteria to Uranium is a heavy radioactive metal naturally present on change the speciation of uranium and its bioavailability. This work earth, which has been widely exploited for industrial and military also led to evaluate their potential for bioremediation of applications. Mining activities, uranium processing and leaching of contaminated areas [5,6,7,8]. Besides uranium reduction in anoxic wastes are significant anthropogenic sources of uranium dissem- conditions, the most frequent phenomenon described for uranium- ination in the environment, in subsurface sediments or ground bacteria interaction is uranium biosorption or precipitation at cell water. Although metals are often met in biological systems and are surface. Uranium binding to cell envelopes (cell wall/membrane) essential for living organisms, uranium is universally recognized as involvescoordinationby negatively charged groups, such as car- a toxin, presenting both chemical and radiological toxicity to living boxylates and phosphoryl groups [7,9–13]. Precipitation as uranyl organisms [1,2]. phosphate [11,14,15] or calcium-uranium-phosphate were notably In the environment, uranium is predominantly observed in two documented [15,16,17,18]. These interactions largely depend on oxidation states, the oxidized U(VI) form, mainly represented as the properties of cell wallsand of the external milieu and may be 2+ the uranyl UO2 cation in the presence of oxygen, and the independent of cells viability [7]. reduced form U(IV), mainly occurring as insoluble species, stable There is far less information concerning intracellular uranium in anoxic conditions. In addition, uranyl may be associated to a accumulation, or consequences of uranium exposure in bacterial large range of organic or inorganic ligands as carboxylates, cells [7]. Uranyl penetration in cells is thought to be associated carbonates or phosphates [3], that determine its solubility and with increased membrane permeability, and intracellular uranium hence its bioavailability, i.e. its ability to bind to or traverse the cell sequestration considered as a passive process involving formation surface of an organism. These properties -called uranium of inorganic compounds as polyphosphate granules [19,20]. It is speciation- have been demonstrated to be both pH and of major interest to enlarge our knowledge on the cellular concentration dependant [3,4]. response of cells to uranium exposure to better understand toxicity mechanisms of uranyl in bacteria and, as a consequence, to

PLOS ONE | www.plosone.org 1 February 2014 | Volume 9 | Issue 2 | e89863 Uranium Bioavailability for Escherichia coli identify potential interesting mechanisms in terms of bioremedia- In this study, we took advantage of the acid resistance property tion. of E. coli to characterize the E. coli K-12 strain MG1655after The response to uranium exposure of environmental bacteria uranium 238 (238U) exposure in LB (1/10)-glucose medium at such as Caulobacter crescentus [16], Geobacter uraniireducens [21], and pH 2.7 (Figure 1). These experimental conditions of exposure Shewanella oneidensis [22] was previously analysed at the transcrip- were chosen mainly for two reasons. First, the pH was fixed below tion level. In the uranium tolerant species C. crescentus, response to pH 4 and the LB medium diluted tenfold to minimize the level of uranium did not overlap substantially with other heavy metal complexation between uranyl and phosphate orcarbonate groups. stresses [16] and this specific response was used subsequently to In addition, diluted LB supplemented with glucose was chosen for develop a whole cell uranyl biosensor [23]. A proteogenomic the presence of various amino acids and organic molecules since approach was also conducted to analyse the evolution of Geobacter this may reduce the level of stress engendered by the hostile community structure and physiology under stimulated uranium growth conditions (i.e. low pH anduranyl exposure). Second, E. coli reduction [24]. has been used in basic research as a model organism for a long While several studies have used two-dimensional gel electro- time. Physiology and genetics of this bacterium are among the best phoresis (2D–E) to investigate changes in protein expression of characterized of all species. Accordingly, the activity of its encoded many bacteria after exposure to a wide range of toxic or biological proteins is probably at the best level of understanding in bacteria. metals, e.g., silver [25], copper, lead, cobalt [26,27,28], or chromium [27,29,30], none focused on the effect of uranium. 2. Uranyl speciation in the exposure medium This may be due to difficulties associated with the control of uranyl The speciation of uranyl in the exposure medium was analyzed speciation in bacteria culture media. Only two papers described using X-ray absorption spectroscopy (XAS) (Figure 2). This the impact of uranium stress on lung and kidney cells proteomes medium, dedicated to bacterial growth is mainly composed of [31,32,33]. proteins or amino acids, sugar moieties and salts. It contains To help uncover the mechanisms of toxicity associated with phosphate residues, but the presence of strong uranyl ligands such exposure to uranyl, we developed an experimental model to as carbonate or citrate is unexpected. Consequently, uranyl may expose the bacterial model Escherichia coli to uranium in conditions be found as complexes with phosphate, glucose, or carboxylate allowing uranium bioavailability and cells survival. To reach this moieties of proteins and amino acids residues [37,38]. Given the objective, we challenged bacterial cells with uranium at low pH in complexity of this medium, uranium speciation is suspected to be a diluted LB medium supplemented with glucose.For these rather complex. experiments, we took advantage of the acid resistance of E. coli, To get a more precise picture of uranium complexation, a XAS one of the best scientifically analyzed organisms that represents an spectrum of the exposure medium (diluted LB medium with invaluable tool in molecular microbiology and biotechnology.The glucose at pH 2.7, see below) supplemented with 50 mM uranyl uranium speciation in the exposure medium was analyzed by acetate was recorded at the uranium LIII-edge, and compared to means of X-ray absorption spectroscopy (XAS), and we analyzed spectra of reference uranyl complexes (Figure 2). The exposure medium was analyzed in the same conditions as during bacterial the proteomic response of the bacteria in light of its genome exposure even if the uranyl concentration is close to the limit of annotation.The regulated proteins were identified by mass detection of the majority of synchrotron beamlines (including those spectrometry after two-dimensional gel electrophoresis (2D–E). dedicated to the analysis of diluted samples such as FAME: The data that we report here may be of interest to optimise strains BM30B, ESRF, France). However, changing the uranyl concen- for biodetection or bioremediation purposes, notably in frame of tration in the medium would have modified its speciation. optimizing in-cell sequestration of uranium. In these conditions, the acquired spectra are noisy and their analysis can only be partial. The XANES (X-ray Absorption Near Results and Discussion Edge Structure) region and the position of the absorption edge 1. Uranium challenge conditions (,17.177 keV, Figure 2-A) is however characteristic of the U(VI) oxidation state [39]. The shoulder at 17.190 keV after the Uranium precipitates as uranium phosphate in culture media at absorption edge, typical from the linear uranyl dioxocation pH above 3.5–4.5 [34], limiting its bioavailability. Exposure of UO 2+, is absent. This suggests that this linear structure may be bacterial cells to uranyl has often been performed in water or 2 distorted in the uranyl complexes formed in the medium. In water supplemented with 0.1 M NaCl or 0.1 M NaClO at 4 addition, EXAFS spectra (Extended X-ray absorption fine various pH, including pH as low as 2 or 3, [7,35] or in the 21 structure Figure 2-B) show a first peak massif (4.5–6 A˚ ) with presence of strong uranyl complexing agents as citrate [17].These 2+ its maximum aligned with that of spectra recorded with UO2 or exposure conditions were used to avoid uranyl precipitation with U-malate, U-acetate, or U-glucose complexes, but not with U- phosphate that could occur in rich growth medium, or in minimal phosphate or U-carbonate complexes. The remaining of the mineral medium containing phosphates.In addition, a large EXAFS spectrum, although very noisy, also differs from spectra number of bacteria precipitate uranium in the form of inorganic recorded with U-phosphate and U-carbonate complexes. The uranyl-phosphate complexes at the cell surface or in the medium second peak massif (6.5–8.8 A˚ 21), aligns with contributions at neutral to mid acidic pH [7]. In these conditions, low uranyl present in the U-malate spectrum, while the peak massif at 9– bioavailability impairs the analysis of the impact of moderate 11 A˚ 21 has a strong contribution in the U-glucose spectrum. Even concentrations of uranyl on bacterial cells physiology. if this interpretation is strictly qualitative, uranium speciation in With Microbacterium isolates, uranyl precipitates in the form of the LB-glucose medium may be a mixture of glucose and inorganic phosphate complexes (meta-autunite) at pH 4.5, while it carboxylate complexes of uranyl. Therefore uranyl is expected to interacts with organic phosphate groups at more acidic pH [35]. be bioavailable for E. coli cells in these experimental conditions. Interaction of uranyl with organic phosphate was also described at low pH for Bacillus subtilis [12], and with protonated phosphoryl 3. Exposure assay of E. coli cells groups and carboxylic sites for Pseudomonas fluorescens [36]. These Acid resistance is an inherent property of the enteric bacterium data are in line with a higher bioavailability of uranyl at low pH. E. coli, associated to its necessity to develop strategies to overcome

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Chemical Geology

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Iron adsorption onto soil and aquatic bacteria: XAS structural study

A.G. González a,b,⁎,O.S.Pokrovskya,c, F. Jiménez-Villacorta d,L.S.Shirokovaa,e,J.M.Santana-Casianob, M. González-Dávila b,E.E.Emnovaf a Géochimie et Biogéochimie Expérimentale, GET UMR 5563 CNRS, Université de Toulouse, 14 Avenue Edouard Belin, 31400 Toulouse, France b Universidad de Las Palmas de Gran Canaria, Departamento de Química, Campus de Taﬁra, 35017 Las Palmas, Spain c BIO-GEO-CLIM Laboratory, Tomsk State University, Lenina 36, Tomsk 643050, Russia d Department of Chemical Engineering, Northeastern University, 02115 Boston, MA, USA e Laboratory of Aquatic Ecosystems, Institute of Ecological Problems of the North, Ural Branch Russian Academy of Science, Arkhangelsk, Russia f Institute of Genetics and Plant Physiology, Moldovan Academy of Sciences, Moldova article info abstract

Article history: Although the interaction between Fe and microorganisms has been extensively studied, the main physico- Received 2 July 2013 chemical factors controlling the mechanisms of Fe adsorption and precipitation on bacterial cell walls remain Received in revised form 12 February 2014 poorly understood. In this study, we quantified thermodynamic parameters of the Fe adsorption reaction and Accepted 13 February 2014 characterized the speciation of Fe adsorbed on the surface of cyanobacteria and soil heterotrophic bacteria. For Available online 24 February 2014 this purpose, the molecular mechanisms of iron interaction with typical aquatic and soil bacteria were investigat- fi Editor: J. Fein ed by combining batch macroscopic adsorption experiments with atomic-level Fe K-edge X-ray absorption ne structure spectroscopy (XAFS). Three cyanobacteria species (Synechococcus sp., Planktothrix sp. and Gloeocapsa 3+ 2+ Keywords: sp.) and aerobic heterotrophic soil rhizobacterium (Pseudomonas aureofaciens) were used for Fe and Fe Iron adsorption experiments. These experiments were carried out for a wide range of initial iron concentration Adsorption (4.5–57.3 μM) and pH (2.0–6.5). Surface adsorption data were rationalized using a Linear Programming Model Cyanobacteria (LPM), which allowed quantification of the surface adsorption constants and the number of binding sites. XAS Heterotrophs (XANES and EXAFS) analysis of adsorbed iron demonstrated the predominance of O-coordinated Fe3+ species. Electrophoresis Moreover, XANES data treatment using a linear combination fit of reference compounds suggested that the atom- XAS ic environment of iron adsorbed onto soil bacterial surfaces was dominated by phosphoryl moieties with a lesser amount of carboxylates and some contribution of Fe(III)-oxy(hydr)oxide component. Complete oxidation of Fe(II) to Fe(III) was observed in the solid phase as determined by XANES analysis. Binding of Fe(III) to carboxylate groups was only significant for capsular cyanobacteria (Gloeocapsa sp.). The relative proportions of various Fe species at the cell surface determined by thermodynamic analysis of the macroscopic data and by XAS are in a good agreement. Our results suggest that, in the presence of surface organic ligands, the oxidation of divalent iron does occur, but the polymerization of formed Fe(III)oxy(hydr)oxides is partially inhibited and adsorbed iron in the form of both Fe–O–Fe polymers and individual Fe atoms attached to phosphoryl moieties. The presence of EPS reduces metal-cell binding capacity and enhances Fe polymerization at the bacterial surface. © 2014 Elsevier B.V. All rights reserved.

1. Introduction (Gledhill and van den Berg, 1994; Donat and Bruland, 1995; Wu and Luther, 1995) as well as in the form of organo-ferric colloids (de Baar Iron is one of the most abundant and biologically important metals and de Jong, 2001; Hassellöv and von der Kammer, 2008). and micronutrients in terrestrial and aquatic environments (Sunda and Microorganisms play a key role in the geochemical cycling of iron Huntaman, 1997; Maldonado and Price, 2001). The geochemical cycling because they can serve as nucleation centres for iron adsorption and of iron is known to control the primary productivity in natural waters. precipitation (Daughney et al., 2004). The mobility, reactivity and bio- The major fraction of iron in aqueous solutions is represented by Fe(III), availability of iron in aqueous solutions are affected by the adsorption the most stable form of iron in oxygenated waters. Dissolved trivalent and uptake processes at the interface between the microorganism and iron is present in solution in the form of organic and hydroxocomplexes the aqueous solution (Geesey et al., 1977; Harvey et al., 1982; Mahmood and Rao, 1993; Corapcioglu and Kim, 1995). Microorganisms have a high capacity for the adsorption of metal onto their cell walls ⁎ Corresponding author at: Géochimie et Biogéochimie Expérimentale, GET UMR 5563 (Beveridge and Murray, 1980; Gonçalves et al., 1987; Beveridge, 1989; CNRS, Université de Toulouse, 14 Avenue Edouard Belin, 31400 Toulouse, France. Tel.:+33 561332649; fax: +33 561332650. Ledin et al., 1996; Barns and Nierzwicki-Bauer, 1997; Fein et al., 1997; E-mail address: [email protected] (A.G. González). Ledin et al., 1999; Wu et al., 2006) through a number of proton- and

http://dx.doi.org/10.1016/j.chemgeo.2014.02.013 0009-2541/© 2014 Elsevier B.V. All rights reserved. A.G. González et al. / Chemical Geology 372 (2014) 32–45 41 several samples (Pl-Fe2, Syn-Fe2 and EPS-rich-3), there was a strong cadmium, zinc or lead (Fein et al., 1997; Pokrovsky et al., 2008a; and broad intensity peak at around ~3 Å, similar to that in goethite González et al., 2010) and consistent with the well-established correla- which probably comes from the Fe–Fe and other neighbouring Fe–Ore- tion between the degree of hydrolysis and the stability constants of gions. Judging from the distances at which this broad peak is located metal–bacterial surface functional group reactions, (Fein et al., 2001). (2.7–3.3 Å), it is most likely originated from polymeric species with It can be assumed that the majority of Fe3+ ions present in solution edge and double corner-sharing Fe–O6 octahedral (Pokrovski et al., can be adsorbed onto bacterial surfaces (Wightman and Fein, 2005), 2003). Overall, the EXAFS results confirm the observations of the although a certain amount of hydrous ferric oxide could also be precip- XANES spectra. itated onto and within the cell surface layers (Warren and Ferris, 1998). The results of the pH-dependent adsorption edge studies show that 4. Discussion Fe3+ adsorption from the aqueous solution is stronger for cyanobacteria compared to soil bacteria. This difference may be partially linked to the 4.1. Surface charge and zeta potential key role of organic functional groups such as phosphoryl which are known to strongly bind Fe(III) in acidic solutions (Kelly et al., 2001; While the zeta-potential of all studied bacteria is essentially negative Boyanov et al., 2003; Tourney et al., 2008). Apparently, a high concen- at pH ≥ 2, the excess adsorption of protons occurs at pH below 4–6. tration of carboxylate and carboxyl-phosphodiester groups in the EPS Therefore, the results of zeta-potential measurements and surface titra- of soil bacteria is capable of protecting the main Fe binding sites of the tion are consistent with the dominance of negatively-charged carboxyl cell surface, the phosphoryl, thus leading to a decrease of the adsorption and phosphoryl functional groups at the outermost bacteria surfaces, yield for P. aureofaciens compared to cyanobacteria. whereas the positively charged amine groups are located at some distance from the surface towards the interior of the cell. Similar consider- 4.3. Molecular structure of adsorbed Fe ations have been put forward for the diatom (González-Dávila, 1995; Gélabert et al., 2004), heterotrophic bacteria (Martinez et al., 2002; The structural results of this study demonstrate that adsorbed Fe is Burnett et al., 2006; Ha et al., 2010) and cyanobacteria (Koelmans always present in the form of Fe(III) species with O local environment. et al., 1996; Pokrovsky et al., 2008a). In fact, a small amount of the Both XANES and EXAFS analyses reveal that Fe adsorption results in more negatively-charged surface functional groups (carboxyl and phos- the formation of O-coordinated Fe3+ products. Additionally, XANES lin- phoryl), even partially deprotonated, would be enough to produce neg- ear combination analysis suggests that the main functional groups re- ative ζ-potential in the full range of pH above 4. sponsible for Fe3+ adsorption are likely to be carboxylates and In the case of soil heterotrophic rhizobacterium P. aureofaciens,the phosphorylates. Even during adsorption from solutions highly supersat- main functional groups responsible for proton binding are most likely urated with respect to Fe-hydroxide, Fe(III) ions associated with carboxylate, phosphoryl and amine with carboxylate being the most Gloeocapsa sp. cyanobacteria are stabilized in the presence of surface abundant, especially in EPS-rich strains. The presence of phosphoryl − carboxylates (bR–COO ), forming spatially separated Fe(III) centres groups is particularly important given their key role in cation complex- on the cell surface alginate-like backbone instead of a mere coverage ation by bacterial surfaces at low pH (Kelly et al., 2001; Boyanov et al., of the surface with FeOOH precipitates. Fe adsorption on NR–PO4 (phos- 2003; Tourney et al., 2008). The total amount of proton-active phospho- phorylates) is the most likely mechanism of Fe-bacteria interaction at ryl binding sites potentially capable of complexing Fe3+ at the bacterial low pH and high Fe loading for the two other cyanobacteria, surface was 0.54 and 0.84 mmol g−1 for EPS-rich and EPS-poor cultures, wet Synechococcus sp., Planktothrix sp., as well as the soil heterotrophic respectively (Table 1). Tentative identifications of surface functional rhizobacterium P. aureofaciens. Regardless of the specificity of the stud- groups and their surface concentrations are comparable with those ied soil and aquatic bacteria, namely the type of cellular organization reported for other microorganisms such as Bacillus subtilis and and the chemical character of surface binding groups, adsorbed Fe(III) Escherichia coli (0.04–0.29 mmol g−1)(Martinez et al., 2002), is probably linked to phosphoryl and carboxylate surface moieties and Shewanella oneidensis (0.18–0.32 mmol g−1)(Ha et al., 2010), Bacillus is not subject to significant polymerization at pH b 5. licheniformis (0.14–1.79 mmol g−1)(Tourney et al., 2008), P. putida The EPS-rich bacteria exhibit a generally higher proportion of Fe poly- (0.04–0.21 mmol g−1)(Ueshima et al., 2008), and bacteria consortia mers at the surface, presumably because the EPS-covered bacterial cell (0.05–0.11 mmol g−1)(Borrok and Fein, 2004). The change of bacterial may serve as a centre of condensation during Fe(II) → Fe(III) oxidation surface speciation in the presence of adsorbed Fe3+ is evidenced from in aqueous suspension. This conclusion corroborates numerous previous both zeta-potential measurements and macroscopic adsorption experiments as further discussed below. 100

Cyanobacteria-Fe(II) 4.2. Effect of Fe(III) on zeta-potential and Fe(III) adsorption Cyanobacteria-Fe(III) 80 EPS-poor-Fe(III) The increase of zeta-potential with the addition of Fe3+ at pH 3 may EPS-rich-Fe(III) reflect 1) negative surface charge screenings due to adsorption of Fe(III) cation on carboxylate or phosphoryl surface groups and 2) formation of 60 surface Fe oxy(hydr)oxide polymers having a strongly positive charge at this acidic pH which thus contributes to overall positive ζ-potential 3+ N ζ 40 at [Fe ]aq 1mM(Fig. 2). The increase of -potential is much smaller % Goethite for EPS-rich cultures compared to EPS-poor strains. This is consistent with weaker adsorption of metal cations on EPS-rich bacterial strains 20 (González et al., 2010 and this study, see below) and, therefore, sup- portive of the first mechanism of ζ-potential increase in the presence of Fe3+ . It can be further hypothesized that Fe(III) polymers, detect- (aq) 0 able by XAFS, occur within the whole EPS layer of P. aureofaciens cells 1234567 and thus are not “visible” on the shear plane during ζ-potential mea- pH surements. The adsorption of Fe3+ on bacterial surfaces started at pH as low as 1.5 and further increased with the pH. The starting pH of Fig. 9. Percentage of goethite-like compound at the bacterial surface computed from the 3+ Fe adsorption is 1 to 3 units lower than that of metals such as copper, linear fitting of the Fe K-edge XANES spectra. 44 A.G. González et al. / Chemical Geology 372 (2014) 32–45

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Journal of Hazardous Materials 273 (2014) 231–238

Contents lists available at ScienceDirect

Journal of Hazardous Materials

j ournal homepage: www.elsevier.com/locate/jhazmat

Comparison of chemical washing and physical cell-disruption

approaches to assess the surface adsorption and internalization of

cadmium by Cupriavidus metallidurans CH34

∗

Aurélien Desaunay, Jean M.F. Martins

LTHE-CNRS-Univ. Grenoble I (UMR 5564), Domaine Universitaire BP 53, 38041 Grenoble Cedex 9, France

h i g h l i g h t s

•

Subcellular distribution of cadmium in Cupriavidus metallidurans CH34 cells.

•

Comparison of a chemical (EDTA washing) and a physical method (physical disruption).

•

EDTA washings strongly overestimated membrane-bound Cd concentrations.

•

The physical method revealed surprisingly over 80% of Cd internalization in the cells.

•

Metal biosorption by bacteria cannot be considered as a surface complexation process.

a r t i c l e i n f o a b s t r a c t

Article history: Bacterial biosorption of heavy metals is often considered as a surface complexation process, without

Received 26 September 2013

considering other retention compartments than cell walls. Although this approach gives a good descrip-

Received in revised form 19 February 2014

tion of the global biosorption process, it hardly permits the prediction of the fate of biosorbed metals in the

Accepted 4 March 2014

environment. This study examines the subcellular distribution of cadmium (Cd) in the metal-tolerant bac-

Available online 16 March 2014

terium Cupriavidus metallidurans CH34 through the comparison of an indirect chemical method (washing

cells with EDTA) and a direct physical method (physical disruption of cells). The chemical washing

Keywords:

approach presented strong experimental biases leading to the overestimation of washed amount of Cd,

Cd biosorption

Bacteria supposedly bound to cell membranes. On the contrary, the physical disruption approach gave repro-

ducible and robust results of Cd subcellular distribution. Unexpectedly, these results showed that over

Surface complexation

Subcellular distribution 80% of passively biosorbed Cd is internalized in the cytoplasm. In disagreement with the common con-

Internalization cept of surface complexation of metals onto bacteria the cell wall was poorly reactive to Cd. Our results

indicate that metal sorption onto bacterial surfaces is only a ﬁrst step in metal management by bacteria

and open new perspectives on metal biosorption by bacteria in the environment, with implications for

soil bioremediation or facilitated transport of metals by bacteria.

1. Introduction This high reactivity leads to metal bioaccumulation by bacteria.

This process is usually described as a unique mechanism of surface

The use of microorganisms to remediate heavy metal contam- metal sorption onto the bacteria cell, although it may be the sum of

ination in waters and soils is a challenging issue for sustainable various retention mechanisms such as metal bounding to the cell

development. Microorganisms, especially bacteria, are known to wall, the cell membrane, speciﬁc carrier proteins, or intracellular

mediate many geochemical processes such as transformation, pre- ligands.

cipitation, dissolution, sorption or accumulation of major or trace The interactions between bacterial cells and metals are gov-

elements [1–3]. They can mobilize, immobilize and/or transform erned by mechanisms of transfer through the membranes [4,5].

metals through speciﬁc mechanisms, which are strongly controlled Upon exposure to high metal concentrations, almost all unicellu-

by the high reactivity of these tiny cells with high speciﬁc area. lar organisms accumulate the metal because of leakage through

the external membrane(s), including, for example, Ca chan-

nels, exchangers and transporters. The bioaccumulation of a toxic

∗ metal is then often lethal to the organism. Heavy-metal resistant

Corresponding author. Tel.: +33 476635604.

␤

E-mail addresses: [email protected], [email protected] (J.M.F. Martins). bacteria such as the Gram-negative -proteobacterium Cupravidius

http://dx.doi.org/10.1016/j.jhazmat.2014.03.004

232 A. Desaunay, J.M.F. Martins / Journal of Hazardous Materials 273 (2014) 231–238

metallidurans CH34, harbour numerous metal resistance genes compartmentalization studies [22–24]. Dynamic subcellular dis-

enabling cell detoxiﬁcation and survival via a number of tribution of Cd was successfully modelled using the parameters

mechanisms. These include complexation, efﬂux, or reductive pre- determined with the physical method. Results were discussed

cipitation that confer tolerance to high concentrations of many in terms of metal bioavailability to bacteria in solution and the

metals, including Cd [6]. This bacterium (DSM2839) has been potential beneﬁts for soil bioremediation through heavy metal

described as highly reactive with metals and is currently used for bioleaching].

remediating polluted soils or to recover precious metals through

bioleaching processes [7].

2. Materials and methods

This bacterium is known to be able to survive to high heavy

metal concentrations such as zinc, cadmium, cobalt, nickel, copper,

2.1. Bacterial stocks and growth conditions

chrome, mercury and lead. This multi resistance to heavy metal is

principally due to the two plasmids pMOL28 (171 kb) et pMOL30

C. metallidurans CH34 is a metal-tolerant Gram-negative aerobic

(234 kb) [8,9]. ◦

bacterium [25]. Bacterial cells were stored in 15% glycerol at −80 C.

The characterization of bacterial surfaces and the knowledge

Bacterial biomass was produced by inoculating aliquots of these

of the chemistry at the interface cell/solution are necessary to

stock suspensions in Tris Salt Medium (TSM) (pH6) completed with

better understand these interactions. Several studies have shown −1 ◦

2 g L of gluconate. Bacteria were grown at 30 C in a rotary tum-

that numerous Gram negative and positive bacteria present sim-

bler at 250 rpm to the late exponential phase (OD600 = 2.2 ± 0.1) and

ilar metal sorption capacities [10–12,13]. Fein et al. [10] have

collected by centrifugation at 5000 × g for 10 min. To normalize the

proposed a “universal” surface complexation model to describe −1

results, the ﬁnal concentration of bacteria was set to 1 g L ± 0.2

the overall binding of metals by bacterial cells. This modelling

for each experimental condition.

approach involves three types of membrane functional groups:

an acid group (carboxyls and phosphodiesters; pK < 4.7), a neu-

2.2. Bacteria exposure to cadmium

tral group (phosphomonoesters; pK∼7) and a basic group (amines

and/or hydroxyls; pK > 8). Although this universal model strongly

Bacterial cells were ﬁrst grown in metal-free culture media until

improved the prediction of metal reactivity to biotic colloids, lit-

reaching the late exponential phase (about 48 h). Aliquots of the

tle is known about the bacteria–metal interactions occurring onto

bacteria suspensions were centrifuged 10 min at 5000 × g, rinsed

cell walls and in particular how these interactions control the fur-

three times with sterile distilled water and resuspended in Cd aque-

ther subcellular distribution of heavy metals in bacteria. Guiné

−5 −4

ous solutions at 0, 10 and 10 M, in order to obtain a ﬁnal 20 mL

et al. [13] have suggested that the binding of cadmium and zinc

−1

assay samples at a bacterial concentration of 1 g L . Assay sam-

on the bacterial membrane of Escherichia coli cells represents only ◦

ples were then incubated for 2 h at 4 C in order to enable only

a weak portion of the metals bioaccumulated by bacteria and that

passive metal uptake by bacterial cells, as described in Guiné et al.

other bacterial components or other mechanisms must be involved

[26]. Cells were then harvested by centrifugation at 5000 × g for

in metal retention by bacteria. Among these are internalization

10 min. The supernatant (referred to as the extracellular compart-

processes or sorption to highly reactive extracellular components. ◦

ment) was collected and stored at −20 C for further metal analysis.

Because of technical limitations, most of past experimental stud-

Bacterial pellets were then processed to obtain bacterial subcellular

ies carried on metal sorption onto bacteria have considered only

compartments as described below.

cell surface reactions, thus neglecting the further subcellular dis-

tribution of metals in bacterial. Indeed recent microscopic and

spectroscopic methodology, such as EXAFS, allow localization of 2.3. Subcellular distribution of Cd

the metal [12,14], or identify the functional groups involved in its

retention by bacterial whole cells [13,15]. However this method 2.3.1. Chemical fractionation with EDTA

cannot be used to quantify the distribution of metals within the The distribution of Cd in the three subcellular compartments

cells. For this, some authors have used indirect methods such as of the bacterial cells was determined chemically using EDTA, a

chemical washing of membrane-bound metals using strong organic strong metal ligand, following the procedure proposed by Hassler

ligands to quantify metal attachment to the cell membrane or to et al. [18] Although EDTA was described as the best suited wash-

follow the internalization rate of metals in bacteria, with variable ing agent for this type of measurements, no accepted experimental

success [16–20,21]. conditions were proposed. EDTA concentration and contact time

Since surface complexed and internalized metals will present with bacterial cells were described as important factors that can

neither the same toxicity nor the same desorption potential, the dif- modulate and bias experimental results [18]. Consequently a pro-

ferentiation between metal that has been internalized by the cells cedure was optimized for our speciﬁc objectives, which consisted

and that adsorbed indiscriminately to cell components appears in the use of a low EDTA concentration (10 mM EDTA in 10 mM

thus extremely relevant and necessary in order to better predict MES buffer) and increasing contact times between cells and the

metal toxicity as well as metal reactive transport by bacteria. washing agent: 6, 7, 10, 15, 20, 35 min (including 5 min of cells

The objectives of the present study were to characterize, at two harvesting by centrifugation). Bacterial cells initially exposed to Cd

relevant concentrations, the subcellular distribution of cadmium in were re-suspended in EDTA solutions for the different contact times

cells of the metal-tolerant bacterium C. metallidurans CH34, pas- with gentle shaking (100 rpm). These suspensions were centrifuged

◦

sively exposed to Cd at 4 C. The subcellular distribution of Cd at 5000 g for 5 min and the supernatants and the pellets were

◦

−

was evaluated by comparing an indirect chemical method with stored at 20 C until Cd analysis. Cadmium amount measured in

a direct physical cell-disruption method. The chemical method the supernatant was washed by EDTA and supposed to be initially

aimed at quantifying membrane-bound Cd speciﬁcally extracted associated with bacterial membranes and periplasm (referred to as

with ethylenediaminetetraacetic acid (EDTA), an organic ligand the membrane compartment). The amount of cadmium measured

described as the best suited for this purpose [18]. This method was in the pellet was assumed to be retained in the cell’s cytoplasm

compared to a physical method based on the direct measurement after internalization (intracellular compartment). It is important to

of Cd concentrations in the membrane and cytoplasmic compart- note that this contact time could not be shorter than 5 min which

ments obtained after cell disruption at high pressure with a French corresponds to the irreducible duration of bacterial cell centrifuga-

Press, a technique widely used in cell biology for macromolecules tion.

A. Desaunay, J.M.F. Martins / Journal of Hazardous Materials 273 (2014) 231–238 237

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pubs.acs.org/CR

Mineral−Water Interface Reactions of Actinides Horst Geckeis,* Johannes Lützenkirchen, Robert Polly, Thomas Rabung, and Moritz Schmidt Karlsruhe Institute of Technology (KIT), Institute for Nuclear Waste Disposal (INE), Karlsruhe, P.O.Box 3640, D-76021 Karlsruhe, Germany 3.3.7. Sorption of Uranyl at the Corundum (001) Surface 1040 3.4. Sorption of Trivalent Curium 1041 3.4.1. Sorption of Trivalent Curium Ions at the Corundum (001) Surface 1041 3.4.2. Sorption of Trivalent Curium Ions at the Corundum (110) Surface 1041 3.5. Classical Molecular Dynamics (MD) and Classical Monte Carlo (MC) Simulations 1042 3.5.1. Sorption of Uranyl Carbonate in Feld- CONTENTS spar Nanosized Fractures 1042 1. Introduction 1016 3.5.2. Sorption of Uranyl on Quartz 1042 2. Actinide Sorption Mechanisms 1018 3.5.3. Sorption of Uranyl on Goethite 1042 2.1. Outer-Sphere Sorption 1018 3.5.4. Sorption of Uranyl on Calcite 1042 2.1.1. Trivalent Actinide Ions 1019 3.5.5. Sorption of Uranyl on 2:1 Dioctahedral 2.1.2. Tetravalent Actinide Ions 1020 Clays 1042 2.1.3. Pentavalent Actinide Ions 1020 4. Geochemical Modeling of Actinide Sorption 1042 2.1.4. Hexavalent Actinide Ions 1020 4.1. Approaches to Ion-Exchange Modeling 1043 2.2. Inner-Sphere Surface Complexation 1020 4.2. Approaches to Surface Complexation Mod- 2.2.1. Trivalent Actinide Ions 1020 eling 1043 2.2.2. Tetravalent Actinide Ions 1023 4.2.1. Treatment of Heterogeneity and Deﬁ- 2.2.3. Pentavalent Actinide Ions 1024 nition of Surface Sites 1045 2.2.4. Hexavalent Actinide Ions 1024 4.2.2. Protonation Mechanism and Electro- 2.2.5. Competitive Sorption 1027 statics 1045 2.3. Actinide Incorporation Processes 1027 4.2.3. Actinide Adsorption Model 1047 2.3.1. Trivalent Actinide Ions 1027 4.2.4. Input Data for Model Development and 2.3.2. Tetravalent Actinide Ions 1030 Parameter Estimation 1049 2.3.3. Penta/Hexavalent Actinide Ions 1030 4.2.5. Generalization of Results 1049 2.4. Surface-Induced Actinide Redox Reactions 1031 4.3. Approaches to Modeling Incorporation of 2.5. Sorption of Intrinsic Actinide Colloids 1032 Actinides in Matrices 1051 3. Theoretical Studies of Actinide Sorption 1032 5. Summary 1052 3.1. Background of Actinide Quantum Chemistry 1033 Author Information 1053 3.2. Quantum Chemical Approaches to Describe Corresponding Author 1053 Sorption on Mineral Surfaces 1034 Notes 1053 3.2.1. Plane Wave DFT Employing Periodic Biographies 1053 Boundary Conditions (PBC) 1034 Acknowledgments 1054 3.2.2. Orbital-Based DFT Employing Finite References 1054 Cluster Models 1034 2+ 3.3. UO2 Sorption 1036 3.3.1. Sorption of Uranyl at Rutile (110) Sur- 1. INTRODUCTION face 1036 − 3.3.2. Sorption of Uranyl at Gibbsite (001) and Reactions of the light actinide ions (Th to Cm) at the mineral Edge Faces of Gibbsite 1037 aqueous solution interface have attracted much attention during 3.3.3. Sorption of the Uranyl Ion onto the recent years. Initially research was mainly driven by the need to Nickel (111) Surface 1038 obtain solid/liquid distribution constants (KD-values) as input 3.3.4. Sorption of the Uranyl Ion at the data for performance assessment modeling related to nuclear Kaolinite (001) Surface 1039 waste repository projects or to the development of remediation 3.3.5. Sorption of the Uranyl Ion onto the Kaolinite (010) Edge Surface 1039 Special Issue: 2013 Nuclear Chemistry 3.3.6. Sorption of the Uranyl Ion onto Received: September 3, 2012 Goethite 1040 Published: January 31, 2013

To quantify actinide sorption, batch sorption experiments are The present review deals comprehensively with the various usually carried out by varying the chemical conditions. Sorption mechanisms of actinide interaction with mineral surfaces, isotherms are determined by measuring the solid−liquid including surface sorption, actinide incorporation into the solid distribution of an actinide ion at variable metal ion matrix, surface induced redox reactions, and colloid adsorption. concentration but at fixed pH. The pH dependent actinide As quantum chemical approaches are increasingly applied to sorption is studied at constant metal concentration and elucidate actinide surface structures and tentative binding − provides the so-called “sorption edges”.22 24 Both sorption modes, we dedicate a separate section to this topic, to provide isotherms and sorption edges provide imperative input data for an overview on the methods used and their benefits. The development of mechanistic geochemical/thermodynamic quantitative assessment of actinide behavior in a given sorption models, able to account for variable geochemical environment, such as contaminated sites or nuclear waste conditions. Combination of results with complementary disposal systems requires geochemical models to describe, for fi spectroscopic and theoretical methods allows for elucidation example, sorption reactions. In the nal section, we give an and verification of sorption mechanisms and actinide surface outline on available modeling approaches, their underlying species. Even though this approach is elaborate, it is much more concepts and required parameters, and discuss exemplary appropriate for predicting contaminant mobility under natural applications to actinide interaction with minerals. ffi Actinide reactions with microbial surfaces are quite conditions than using a single distribution coe cient (KD value) determined for a specific geochemical environment comparable to interactions with mineral surfaces. Bacterial cell walls consist of exopolymers, proteins and lipids and, thus, (see also discussion in section 4). Such KD values are, in general, only applicable at trace concentration ranges, for contain various surface functional groups potentially capable of reversible sorption reactions and at invariable geochemical interacting with actinide cations. While hydroxyl groups conditions. While single distribution coefficients are still used in dominate mineral surfaces, bacteria contain in addition migration calculations, the development of accurate mecha- carboxylate, phosphate, thiol, and amino moieties, where carboxylate and phosphate groups generally exhibit high nistic models is obviously reaching limits when the complexity affinities for actinide cations (see, for example, ref 28). The of real-world surfaces becomes involved. range of possible surface sorption mechanisms, such as outer- The complexity of the light actinide aquatic chemistry has a sphere binding, inner-sphere surface complexation, surface strong impact on sorption. Under naturally relevant conditions precipitation, and surface induced redox reactions, have also actinide ions can exist in various redox states (An(III), An(IV), been observed for actinide−microbe interactions. Geochemical An(V), An(VI)). The actinide redox state and speciation of the modeling approaches for quantitative description are virtually actinides change due to surface reactions and can form a broad the same for metal cation sorption to mineral surfaces and variety of complexes with groundwater constituents, mainly − 2− biosorption. Under certain conditions, biomass can dominate OH and CO3 . While the tri- and tetravalent actinide species actinide solid−liquid interface reactions, for example, in soil form more or less spherical aquo-ions with 9−10 water fi systems. Unlike mineral surfaces, living cells can actively take molecules in the rst coordination sphere, penta- and up metal ions and transfer them to the cell interior and hexavalent cations form actinyl cations with covalently bound sequester them in the intracellular medium. Many of these axial oxygen atoms and 4- to 6-fold coordination in the processes have been investigated in recent years by various equatorial plane (e.g., ref 25). Steric constraints are thus to be types of spectroscopies. It is outside the scope of this review to expected for the interaction of actinyl cations with surfaces. treat this topic comprehensively, but a wealth of studies is Notably the tetravalent actinide but also the hexavalent actinyl available in the open literature; some can be found cited in cations tend to form oligomeric or even polymeric species more comprehensive reviews (e.g., refs 29−34). which may interact with surfaces by either electrostatic forces or chemical bonds. This chemical complexity has to date 2. ACTINIDE SORPTION MECHANISMS prohibited development of any unified model for dealing with “ ” actinide sorption under varying geochemical conditions. A The term sorption encompasses a variety of possible further inherent difficulty is related to the unique properties of processes and does not make any distinction with regard to different mineral surfaces. Even actinide sorption onto pure and the underlying mechanism. In this section, we deal with well-defined crystal planes is not entirely understood, actinide reactions at mineral surfaces by various surface demonstrating the need for further study. phenomena such as outer-sphere attachment and inner-sphere A recent review by Tan et al.4 summarizes available studies surface complexation of ionic and colloidal actinide species. on actinide/lanthanide interaction with mineral surfaces, Beyond pure surface attachment and complexation reactions, including detailed descriptions of spectroscopic methods. Tan actinides can react with minerals by incorporation (mineralization) and surface induced redox reactions (see Figure 1). et al. discuss examples of actinide/lanthanide interface reactions in order to elaborate mechanistic insight provided by individual 2.1. Outer-Sphere Sorption spectroscopic techniques, such as TRLFS and EXAFS. Another Cations can interact with negatively charged mineral surfaces by review on actinide speciation is ref 26, providing a nice purely electrostatic attraction. Such reactions are well-known overview of actinide reactions in environmental compartments for cation interactions with permanently charged clay mineral including complexation with major relevant natural ligands surfaces (i.e., the typical “ion-exchange” sites). The origin of the abundant in groundwater and naturally occurring biological relevant surface charge for this type of interaction in clay (microbial) and mineral surfaces. Another focus is placed on minerals is isomorphic substitution of, for example, Al in AlO6 actinide environmental behavior at contaminated sites. More octahedral layers by divalent cations (Mg/Fe(II)) and/or Si in detailed information on aquatic actinide chemistry and notably tetrahedral SiO4 layers by trivalent cations (Al/Fe(III)). the related thermodynamics is available from another Permanent charge densities can be experimentally determined contribution to this issue.27 by quantifying their cation exchange capacity (CEC), which

1018 dx.doi.org/10.1021/cr300370h | Chem. Rev. 2013, 113, 1016−1062 Chemical Reviews Review

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1055 dx.doi.org/10.1021/cr300370h | Chem. Rev. 2013, 113, 1016−1062

Forum Article

pubs.acs.org/IC

Environmental Speciation of Actinides † ‡ † ‡ § Kate Maher,*, John R. Bargar, and Gordon E. Brown, Jr. , , † Department of Geological & Environmental Sciences, Stanford University, Stanford, California 94305-2115, United States ‡ § Stanford Synchrotron Radiation Lightsource and Department of Photon Science, SLAC National Accelerator Laboratory, 2575 Sand Hill Road, MS 69, Menlo Park, California 94025, United States

ABSTRACT: Although minor in abundance in Earth’s crust (U, 2−4 ppm; Th, 10−15 ppm) and in seawater (U, 0.003 ppm; Th, 0.0007 ppm), light actinides (Th, Pa, U, Np, Pu, Am, and Cm) are important environmental contaminants associated with anthropogenic activities such as the mining and milling of uranium ores, generation of nuclear energy, and storage of legacy waste resulting from the manufacturing and testing of nuclear weapons. In this review, we discuss the abundance, production, and environmental sources of naturally occurring and some man-made light actinides. As is the case with other environmental contaminants, the solubility, transport properties, bioavailability, and toxicity of actinides are dependent on their speciation (composition, oxidation state, molecular-level structure, and nature of the phase in which the contaminant element or molecule occurs). We review the aqueous speciation of U, Np, and Pu as a function of pH and Eh, their interaction with common inorganic and organic ligands in natural waters, and some of the common U-containing minerals. We also discuss the interaction of U, Np, Pu, and Am solution complexes with common Earth materials, including minerals, colloids, gels, natural organic matter (NOM), and microbial organisms, based on simplified model system studies. These surface interactions can inhibit (e.g., sorption to mineral surfaces, formation of insoluble biominerals) or enhance (e.g., colloid-facilitated transport) the dispersal of light actinides in the biosphere and in some cases (e.g., interaction with dissimilatory metal-reducing bacteria, NOM, or Mn- and Fe-containing minerals) can modify the oxidation states and, consequently, the behavior of redox-sensitive light actinides (U, Np, and Pu). Finally, we review the speciation of U and Pu, their chemical transformations, and cleanup histories at several U.S. Department of Energy field sites that have been used to mill U ores, produce fissile materials for reactors and weapons, and store high-level nuclear waste from both civilian and defense operations, including Hanford, WA; Rifle, CO; Oak Ridge, TN; Fernald, OH; Fry Canyon, UT; and Rocky Flats, CO.

■ INTRODUCTION transport of uranium at several U.S. DOE sites (e.g., Hanford, fl Over the last century, the mining and processing of uranium WA; Oak Ridge, TN; and Ri e, CO) and of plutonium at the ores and the manufacturing and testing of nuclear weapons Rocky Flats Environmental Technology Site near Denver, CO, have created a legacy of contamination of uranium and other have been studied extensively. This work has led to an actinides in soils and groundwater, particularly at U.S. enhanced understanding of uranium and plutonium speciation ff Department of Energy (U.S. DOE) sites that have manufac- and biogeochemical dynamics under di erent environmental tured nuclear weapons or store high-level radioactive waste. In conditions and to the elucidation of key scientific challenges, addition, driven by the demand for more electricity and the particularly in the selection and optimization of remediation need for alternative energy sources that emit less CO2 than strategies. Further insight into the environmental behavior of fossil fuels, there has been a renewed interest in nuclear power actinides is gained from studies of the oxidative weathering of − and advanced nuclear fuel cycles over the past decade.1,2 The uranium ore deposits, such as Nopal I, Mexico,5 10 which projected acceleration in the development of uranium and provide examples of long-term processes that may govern the thorium resources worldwide suggests that actinides will evolution of storage repositories.7 continue to pose a threat to the environment and water This review discusses the speciation of some of the most resources long into the future, unless improved methods for common actinides (Th, U, Np, and Pu) in aqueous solutions extraction, processing, and storage are developed and and in solids common in the environment, as well as their 3,4 deployed. interactions with mineral surfaces, organic matter, and In the United States, the long history of uranium mining, processing, and disposal provides a platform for evaluating the Special Issue: Inorganic Chemistry Related to Nuclear Energy effects of improper handling of actinide-bearing waste solutions and solids and the efficacy of remediation strategies, particularly Received: July 31, 2012 for uranium. For example, the speciation, mobility, and Published: November 8, 2012

© 2012 American Chemical Society 3510 dx.doi.org/10.1021/ic301686d | Inorg. Chem. 2013, 52, 3510−3532 Inorganic Chemistry Forum Article conditions studied. Current surface complexation models of ThIV,NpV, and AmIII have also been found to form inner- uranyl sorption on clay minerals should be reevaluated to sphere complexes on various mineral surfaces using EXAFS and − account for the possible increased importance of cation- X-ray photoelectron spectroscopies.80,115,117,126 131 exchange reactions at low ionic strengths, the presence of Interaction of U6+ with Iron (Oxyhydr)oxides. In reactive octahedral iron surface sites, and the formation of addition to forming sorption complexes on iron (oxyhydr)oxide 2+ uranylcarbonato ternary surface complexes. Considering the surfaces, actinide ions such as UO2 can also be sequestered adsorption mechanisms observed in the study by Catalano and through the incorporation into or physical association with iron 38 Brown, future studies of UVI transport in the environment (oxyhydr)oxides such as ferrihydrite and their transformation should consider how uranium retardation will be affected by products.132,133 However, this alternative sequestration mech- changes in key solution parameters, such as the pH, ionic anism is not fully understood. For example, we do not know strength, exchangeable cation composition, and presence or how the incorporation of common impurities associated with fi absence of CO2. The ndings of this study are consistent with natural ferrihydrites, such as Al3+ or Si4+, impact (promote or the generalization that actinides may be adsorbed via an ion- retard) the additional incorporation of trace impurities such as exchange mechanism under acidic solution conditions where 2+ UO2 . A prerequisite to achieving this understanding is uptake of positively charged actinide species is generally low. detailed knowledge of natural nanominerals such as ferrihydrite, However, as the pH increases to the near-neutral range and which is poorly crystalline and typically shows two diffuse hydrolysis of actinide solution complexes is more favored, diffraction lines for the common ferrihydrite nanoparticle size uptake is maximized. At higher pH values, actinide uptake is range (1−2 nm). In a baseline study of the ferrihydrite generally reduced because of the formation of anionic hydroxo structure, a high-energy (90 keV) total X-ray scattering or carbonato actinide complexes in solution. experiment on synthetic two-line ferrihydrite was carried out The carbonate oxoanion is an important component of soil as a function of aging in the presence of a citrate solution at 175 and aquatic systems and forms strong complexes with actinyl °C.134 While aging under these conditions results in the ions such as UO 2+ at circum-neutral and higher pH (Figure 3). 2 formation of hematite in ∼14 h, analysis of the atomic pair Consequently, as shown in Figure 6, at pH values >8, uranyl distribution functions and complementary physiochemical and magnetic data indicate the formation of intermediate ferrihydrite phases at aging times of ≤8 h with larger particle size (up to 12 nm), fewer defects (i.e., iron vacancies), less structural water, less lattice strain, and electron-spin ordering, which results in pronounced ferrimagnetism relative to its disordered antiferromagnetic ferrihydrite precursor. This study also showed that the two-line ferrihydrite structure determined by Michel et al.,135 although highly defective, is topologically very similar to the ordered ferrimagnetic ferrihydrite structure andthusprovidesabasisforunderstandingstructural constraints on the incorporation of impurity ions into the two-line ferrihydrite structure. The proposed structure of two- line ferrihydrite is shown in Figure 7 (together with a ff 2+ Figure 6. (A) E ect of the pH and CO2 on the sorption of UO2 on photograph of natural ferrihydrite) and consists of layers of 103,104 105 goethite and kaolinite (after Thompson et al. ). edge-shared FeO6 octahedra (Fe1 sites) in the xy plane adsorption complexes on goethite and kaolinite surfaces can be desorbed in favor of the formation of this solution − complex.36,103 105 In an earlier EXAFS study, Chisholm-Brause et al.106 also found that uranyl binds to several different sites on montmorillonite. More recently, Bargar et al.35,107 showed that stable uranium(VI) carbonato complexes can form on hematite surfaces. Subsequent studies have shown that Ca2+ can competetively inhibit the adsorption of uranylcarbonato complexes on quartz and ferrihydrite surfaces because of the 2‑ 0 formation of stable CaUO2(CO3)3 and Ca2UO2(CO3)3 108,109 aqueous complexes. Figure 7. (A) Photograph of massive ferrihydrite in a stream bed in an There have been many EXAFS spectroscopic studies of the acid mine drainage system associated with a mercury mine in central compositions, geometries, and binding modes of UVI sorption California. (B) Average hexagonal unit cell structure of ordered complexes at mineral/water interfaces over the past 20 years ferrimagnetic ferrihydrite (after aging for 8 h in the presence of a 110 88 citrate solution at 175 °C) determined from high-energy total X-ray (see Brown and Sturchio, Geckeis and Rabung, Antonio fi and Soderholm,111 and Tan et al.112 for reviews of some of scattering and pair distribution analysis. Re nement of iron site VI occupancies in two-line ferrihydrite indicates that all of the Fe1 sites these studies). More recent studies include U /montmor- − 113−115 VI 116 VI 117 VI are occupied but only 50 55% of the Fe2 and Fe3 sites are occupied illonite, U /kaolinite, U /feldspar, U /ferric by Fe3+, potentially allowing the precursor for U6+/5+ incorporation. 118−120 VI γ 121,122 VI 123,124 VI oxides, U / -Al2O3, U /calcite, and U / The arrows represent magnetic moments. The composition of 125 VI · MnO2. These studies have generally shown that U forms disordered two-line ferrihydrite is Fe8.2O8.5(OH)7.2 3H2O, whereas dominantly inner-sphere complexes on oxygen-based mineral the composition of ordered ferrimagnetic ferrihydrite (after aging) is · − 134 surfaces, typically with bidentate linkages to surface oxo groups. Fe10O14(OH)2 0.5 1.2H2O (from Michel et al. ).

3516 dx.doi.org/10.1021/ic301686d | Inorg. Chem. 2013, 52, 3510−3532 Inorganic Chemistry Forum Article

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Implementation of microbial processes in the performance assessment of spent nuclear fuel repositories ⇑ Thilo Behrends a, , Evelyn Krawczyk-Bärsch b, Thuro Arnold b a Department of Earth Sciences, Faculty of Geosciences, Utrecht University, PO Box 80021, NL-3508 TA Utrecht, The Netherlands b Institute of Radiochemistry, Helmholtz-Zentrum Dresden-Rossendorf e.V., PO Box 510119, D-01314 Dresden, Germany article info abstract

Article history: Present strategies for the long-term disposal of high-level nuclear wastes are based on the construction of Available online 19 September 2011 repositories hundreds of meters below the earth surface. Although the surrounding host-rocks are relatively isolated from the light at the earth surface they are by no means lifeless. Microorganisms rule the deep part of the biosphere and it is well established that their activity can alter chemical and physical properties of these environments. Microbial processes can directly and indirectly affect radionuclide migration in multiple ways. Within 6th FP IP FUNMIG the interplay between microbial biofilms and radionuclides and the effect of microbially induced redox transformations of Fe on radionuclide mobility have been investigated. For the first time, formation of U(V) as a consequence of microbial U(VI) reduction in a multi-species biofilm was detected in vivo by combining laser fluorescence spectroscopy and confocal laser scanning microscopy. Furthermore, it was demonstrated that addition of U(VI) can lead to increased respiratory activity in a biofilm. Increased respiration in a biofilm can create microenvironments with lower redox potential, and hence induce reduction of radionuclides. Transient mobilization of

U was observed in experiments with Fe oxides containing adsorbed U(VI) in which the activity of SO4- reducing organisms was mimicked by sulfide addition. Faster reaction of sulfide with Fe oxides compared to U(VI) reduction, and decreasing U(VI) adsorption due to the transformation of Fe oxides into FeS can explain the observed intermittent U mobilization. The presented research on microbe-radionuclide interactions performed within FUNMIG addresses only a few aspects of the potential role of microorganisms in the performance assessment of nuclear waste repositories. For this reason, additionally, this article provides a cursory overview of microbial processes which were not studied within the FUNMIG project but are relevant in the context of performance assessment. The following aspects are presented: (a) the occurrence and metabolic activity of microorganisms of several proposed types of host-rocks, (b) the potential importance of microorganisms in the near-field of nuclear waste repositories, (c) indirect effects of microbial processes on radionuclide mobility in the repository far-field, (d) binding of radionuclides to microbial biomass, (e) microbial redox transformations of radionuclides, and (f) the implementation of microbial processes in reactive transport models for radionuclide migration. Ó 2011 Elsevier Ltd. All rights reserved.

1. Introduction brief overview on the various aspects of microbial activity and their relevance in performance assessment. This overview is, in Within FUNMIG, a large variety of processes relevant for the particular, intended for readers who are unfamiliar with microbial performance assessment of radionuclide migration from the processes, in order to illustrate the motivation and context of stud- near-ﬁeld of radioactive waste repositories to the biosphere/ ies performed within FUNMIG. Providing a comprehensive review hydrosphere have been investigated. Summarizing the results on the effect of microbial processes on radionuclide mobility lies achieved within FUNMIG in most cases inherently provides a com- outside the scope of this article and more detailed information prehensive overview about the current state of knowledge of the can be found in the various review articles which are referred in related process and its relevance for radionuclide migration. The the different sections. role of microorganisms in radionuclide migration, however, is so diverse that the activities within FUNMIG only cover a small part of the whole picture. For this reason, the article also includes a 2. Microbial activity in different host-rocks

Microorganisms are an essential component of the earth’s biota ⇑ Corresponding author. Tel.: +31 30 253 5008; fax: +31 30 253 5302. representing a signiﬁcant part of the living biomass on earth E-mail address: [email protected] (T. Behrends). (Whitman et al., 1998). Their occurrence extends from the earth’s

induced formation of clay minerals, which can reduce the perme- can increase the solubility of UO2 and enhance the rates of UO2 dis- ability of igneous rocks (Hama et al., 2001; Tuck et al., 2006). solution (Frazier et al., 2005). Consequently, bacterial production of Another relevant group of indirect processes are the production siderophores can counteract U immobilization by microbial reduc- or consumption of reactive oxidants or reductants by microorgan- tion of U(VI) to U(IV). isms or, in other words, the microbially induced shift of redox con- Finally, precipitation of dissolved radionuclides can be caused ditions. Aerobic respiration can effectively scavenge O2 from by the microbial production of ligands, which, in turn, form highly aerated groundwater in granite pores, a process which has been insoluble precipitates with radionuclides. One example is the liber- investigated in detail at the Aspö HRL (Kotelnikova and Pedersen, ation of phosphate from organic molecules by bacterial phospha-

1998; Puigdomenech et al., 2001). Effective removal of O2 from tase leading to the precipitation of uranyl phosphates (Jroundi the percolating solution is a prerequisite for the initiation of anaer- et al., 2007; Martinez et al., 2007). Release of PO4 from an organic obic respiration and microbial reduction of radionuclides and re- phosphate source such as glycerol-3-phosphate has been proposed duces the exposure time of waste containers to O2 and possible as an alternative bioremediation approach for U contaminated corrosion damage. sites at which the presence of NO3 does impede bioreduction of Microbial Fe(III) and SO4 respiration leads to the formation of U(VI). The U containing precipitates formed upon the degradation reduced Fe and S species which can act as efficient reductants for of glycerol-3-phosphate by Rhanella species have been identified as radionuclides. Of special interest hereby are mixed valence Fe oxi- chernikovite. (Beazley et al., 2007, 2009). Indirect effects might be des. Microbial formation of mixed valence Fe oxides and their of greater significance for the migration of radionuclides in the far- capability to reduce radionuclides has been reported in numerous field of repositories than direct microbe/radionuclide interactions. studies (Lloyd et al., 2000; Nakata et al., 2004; Behrends and Van Adequate implementation of indirect effects into reactive transport Cappellen, 2005; Behrends et al., 2005; Scott et al., 2005). Other modelling of radionuclides requires a comprehensive description examples for potential radionuclide reduction by microbial respi- of the biogeochemical reaction network and its external forcing. ration products include the reduction of U(VI) by dissolved S(-II) Examples for the incorporation of microbial dynamics into reactive

(Hua et al., 2006) or pyrite (FeS2)(Wersin et al., 1994) which are transport models are given in a later section of this article. both produced due to the activity of SO4-reducing bacteria. Boonchayaanant et al. (2010) concluded that the rates of U(VI) 5.2. Direct effects in the far-field reduction in experiments with the SO4-reducing organism Desulf- ovibrio aerotolerans can be ascribed to the abiotic reaction between Here, direct effects are defined as interactions between radio- U(VI) and the microbially produced S(-II). Hence the indirect path- nuclides occurring at the surface or interior of microbial cells. This way appeared to be more relevant than the direct enzymatic includes biosorption and bioaccumulation as well as enzymatically reduction of U(VI) by the organisms. Qafoku et al. (2009) showed catalyzed redox transformations of radionuclides. These processes that U can be sequestered by framboidal pyrite, which formed are the subjects of several reviews on interactions between micro- upon microbial reduction of Fe(III) and SO4 in sediments. Uranium organisms and radionuclides (Keith-Roach and Livens, 2002; associated with this pyrite was not completely reduced but was Landa, 2005; Lloyd and Renshaw, 2005). Direct interactions be- detected in both redox states, U(VI) and U(IV). In contrast, U(IV) tween microorganisms and radionuclides are relevant for the was not associated with framboidal pyrite or FeS of presumably safety assessment of HLNWR but they also have been investigated microbial origin in sediments from an abandoned U mine (Suzuki extensively in view of their importance for the risk assessment and et al., 2005). The U(IV) was found to be concentrated at bacterial remediation of sites contaminated with radionuclides. Recently, for cell walls indicating that direct enzymatic reduction was responsi- example, in situ experiments performed within the framework of ble for U immobilization and burial in these sediments . the Integrated Field Research Challenge initiative at Rifle, CO, The third group of indirect effects is the microbial production or USA, (Anderson et al., 2003; N’Guessan et al., 2008), at the Field Re- consumption of radionuclide binding ligands which change the search Site at Oak Ridge, TN, USA (Wu et al., 2006a,b; Cardenas radionuclide speciation. Degradation of soluble complexes of or- et al., 2008) or other locations (Senko et al., 2002) have provided ganic ligands and radionuclides can lead to adsorption or precipi- new insight into the influence of microorganisms on U mobility tation of the released free metal ion. For example, metabolisation in complex field settings. Banaszak et al. (1999) reviewed interac- of citrate caused polymerization of Pu when Pseudomonas fluores- tions between microorganisms and radionuclides regarding the cens was added to Pu(IV) in the presence of excess citric acid (Fran- remediation of sites contaminated with actinides and organic acti- cis et al., 2006). In contrast, bioformation of chelating agents nide complexing compounds. forming soluble complexes with radionuclides can facilitate the transport of actinides and fission products from disposal sites. 5.2.1. Biosorption Johnsson et al. (2006) investigated the partitioning of Fe, Pm, Th Biosorption of metals is considered an important process affect- and Am between quartz sand and aqueous solution in the absence ing the transport of trace metals in subsurface environments (Bor- and presence of exudates from three different bacterial strains of rok et al., 2004) and sorption of metals to biomolecules has several the genera Pseudomonas and Shewanella. In comparison with con- technical applications. In the group of actinides U is the most fre- trols, the concentration of these elements in the dissolved phase quently studied element and an overview of metabolism indepen- increased by more than 40% in the presence of the exudates. This dent uptake of U has been given by Suzuki and Banfield (1999). The increase was most likely due to the binding of the metals to com- cell walls of microorganisms are important sorbents for radionuc- plexing ligands excreted by the bacteria. A special group or organic lides (McLean et al., 1996) but also extracellular polymers such as ligands with high affinity for metals are siderophores. Siderophores S-layers or polysaccharides (Pollmann et al., 2006) can have a high are synthesized by organisms to satisfy their demand for Fe by affinity for radionuclides (Johnsson et al., 2006). X-ray absorption mobilizing Fe3+ from solid phases. These siderophores can also spectroscopy revealed that U(VI) forms inner sphere complexes have a high affinity for radionuclides and siderophore production with carboxyl and phosphoryl groups at bacterial cell walls and can lead to radionuclide mobilization. The stability constants of S-layers. (Kelly et al., 2001, 2002; Panak et al., 2002b; Merroun complexes with pyoverdins, produced by the organism Pseudomo- et al., 2005). Similarly, Pu(VI) has a high affinity for phosphoryl nas fluorescens, and Cm(III), U(VI), and Np(V) are greater than those groups at bacterial surfaces (Panak et al., 2002a). When biosorption of EDTA by factors of 1.05, 1.3, and 1.83, respectively (Moll et al., of Pu(VI) was followed by reduction of Pu(VI) to Pu(V) most of the 2009). Presence of the bacterial siderophore desferrioxamine-B sorbed Pu was released back into solution due to the weaker 460 T. Behrends et al. / Applied Geochemistry 27 (2012) 453–462

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Geochimica et Cosmochimica Acta 75 (2011) 2684–2695 www.elsevier.com/locate/gca

Evidence for multiple modes of uranium immobilization by an anaerobic bacterium

Allison E. Ray a,b,⇑, John R. Bargar c, Vaideeswaran Sivaswamy d,1, Alice C. Dohnalkova e, Yoshiko Fujita b, Brent M. Peyton d,2, Timothy S. Magnuson a

a Department of Biological Sciences, Idaho State University, Pocatello, ID 83209, USA b Biological Systems Department, Idaho National Laboratory, Idaho Falls, ID 83415, USA c Stanford Synchrotron Radiation Lightsource, 2575 Sand Hill Rd., Menlo Park, CA 94025, USA d Center for Multiphase Environmental Research and Department of Chemical Engineering, Washington State University, P.O. Box 642719, Pullman, WA 99164-2719, USA e Environmental Molecular Sciences Laboratory, Paciﬁc Northwest National Laboratory, Richland, WA 99352, USA

Received 19 March 2010; accepted in revised form 9 February 2011; available online 2 March 2011

Abstract

Microbial reduction of hexavalent uranium has been studied widely for its potential role in bioremediation and immobilization of soluble U(VI) in contaminated groundwater. More recently, some microorganisms have been examined for their role in immobilization of U(VI) via precipitation of uranyl phosphate minerals mediated by microbial phosphate release, alleviating the requirement for long-term redox control. Here, we investigated the mechanism of U(VI) removal mediated by an environmental isolate, strain UFO1, that is indigenous to the Field Research Center (FRC) in Oak Ridge, TN and has been detected in U(VI)- contaminated sediments. Changes in U(VI) speciation were examined in the presence and absence of the electron-shuttling moiety, anthraquinone-2,6-disulfonate (AQDS). Cell suspensions were capable of nearly complete removal of 100 lM U(VI) from solution within 48 h; U(VI) removal was not dependent on the presence of an exogenous electron donor or AQDS, although AQDS increased the rate of U(VI) removal. X-ray Absorption Near Edge Structure (XANES) and Extended X-ray Absorption Fine Structure (EXAFS) spectroscopic measurements indicated that U(IV) was the predominant oxidation state of uranium in cell suspensions in both the absence and presence of 100 lM AQDS. Interestingly, 17% of the cell-associated precipitates in a U(VI)-treated suspension that lacked AQDS had spectral characteristics consistent with a uranyl phosphate solid phase. The potential involvement of phosphate was consistent with observed increases in soluble phosphate concentrations over time in UFO1 cell suspensions, which suggested phosphate liberation from the cells. TEM-EDS conﬁrmed the presence of uranyl phosphate with a U:P ratio consistent with autunite (1:1). EXAFS analyses further suggested that U(IV) was bound to low-Z neighbors such as C or P, inferred to be present as functional groups on biomass. These results suggest that strain UFO1 has the ability to facilitate U(VI) removal from solution via reductive and phosphate precipitation mechanisms. Both mechanisms oﬀer potential for the remediation of U-contaminated sediments at the FRC or elsewhere. Ó 2011 Elsevier Ltd. All rights reserved.

⇑ Corresponding author. Present address: Biofuels and Renew- able Energy Technologies, Idaho National Laboratory, P.O. 1. INTRODUCTION Box 1625, Idaho Falls, ID 83415, USA. Tel.: +1 208 526 4554; fax: +1 208 526 0828. Uranium is a major soil and groundwater contaminant E-mail address: [email protected] (A.E. Ray). 1 at 12 of the 18 major Department of Energy (DOE) facili- Present address: NLC Nalco India Limited, 20A Park Street, Kolkata 700016, India. ties due to nuclear fuels and weapons production and waste 2 Present address: Department of Chemical and Biological reprocessing (Riley et al., 1992). At Area 3 of the former Engineering, Montana State University, Bozeman, MT 59717, DOE Field Research Center (FRC; currently the site is USA. known as the Oak Ridge Integrated Field-Research

0016-7037/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.gca.2011.02.040 Multiple modes of uranium immobilization by an anaerobic bacterium 2689 linear combination of spectra for two reference compounds, ity of the data to the fit, linear combination fits were also autunite and UO2.00 (Fig. 3). In the sample without AQDS, applied using schoepite and rutherfordine in place of autun- 17 ± 10% of the uranium was found to be U(VI), and ite. As shown in Fig. EA-1, both fits were significantly 83 ± 10% was determined to be in the U(IV) oxidation worse than obtained with autunite. EXAFS spectra of state. In the sample treated with 100 lM AQDS, the ura- U(VI) phosphates are generally similar to each other (Fuller nium was found to be 100 ± 10% U(IV). These results sug- et al., 2002); hence, this result does not lead to unambigu- gest that, regardless of the presence of an electron-shuttling ous identification of autunite, but does provide evidence moiety, uranium species after incubation with UFO1 were for the presence of a member of the autunite mineral fam- primarily in the reduced form (U(IV)) and reduction was ily. These results suggest that a partial explanation for the the predominant mechanism for U(VI) removal under the removal of soluble U(VI) from cell suspensions in the ab- conditions tested. sence of an exogenous electron donor could be the precipi- ULIII-edge EXAFS spectra for UFO1 treated with tation of U(VI)-containing minerals; Sivaswamy previously 100 lM U(VI) in the presence and absence of AQDS are reported the precipitation of a U(VI) phosphate phase by shown in Fig. 4. The EXAFS spectra for these samples dif- Cellulomonas strain ES6 (Sivaswamy et al., 2011). Linear fer significantly in the region between 6 and 9 A˚ 1, where combination fits were also attempted using reference spec- the positive antinode of the no AQDS spectrum has sub- tra for U(VI) surface complexes on cells of P. fluorescens stantially larger amplitude than that of the 100 lM AQDS (Bencheikh-Latmani et al., 2003); these spectra were used sample. Differences between these two samples are expected because attempts to prepare analogous samples of U(VI) from the results from XANES analyses. To test the hypoth- sorbed onto aerobic live and heat-killed cell suspensions esis that the inventory of U(VI) in the no-AQDS sample of strain UFO1 for EXAFS analysis were unsuccessful. was present as a U(VI) phosphate precipitate, its EXAFS This component however did not appear to contribute to spectrum was fit using a linear combination of EXAFS the no-AQDS EXAFS spectrum, suggesting that biomass- from autunite and the AQDS sample. The latter spectrum sorbed U(VI) was below the detection limit in the samples. provides a representation of a fully reduced sample, i.e., The chemical and physical nature of the U(IV) in the no uranyl phosphate present. A good reproduction of the AQDS-reacted sample (Fig. 4, spectrum a) remains unclear. EXAFS of the no-AQDS sample was obtained. The fit indi- Qualitative inspection of the Fourier Transform of the EX- cated that 17% of U in the sample was consistent with a AFS spectrum from the 100 lM AQDS sample (Fig. 5) uranyl phosphate (autunite). This result agrees well with shows the absence of the 3.8 A˚ peak (R + dR) that is char- the XANES-derived conclusion that 17% of U was present acteristic of UO2 (Schofield et al., 2008), suggesting that as U(VI) in this sample. To qualitatively assess the sensitiv- uraninite is not present above the detection limit (ca 10– 20% of total U). If UO2 was not present, then it must be concluded that U(IV) is present as a molecular complex sorbed to biomass (Senko et al., 2007; Kelly et al., 2008; Bernier-Latmani et al., 2010; Fletcher et al., 2010) or present as a different mineral (Bernier-Latmani et al., 2010). Biomass on which sorption could occur is abundant, making this a reasonable conclusion. To test this hypothesis, shell-by-shell EXAFS fits were performed on the AQDS-reacted sample. The EXAFS are dominated by an O shell with a corresponding FT peak at ca 1.75 A˚ (R + dR) (FT peak values are phase-shifted relative to the physical arrangement of atoms and are thus denoted as R + dR). A secondary, smaller FT peak occurs at ca 2.9 A˚ (R + dR). EXAFS were fit up to 9.5 A˚ 1. The spline was ˚ 1 truncated at 10.2 A to mitigate the impact of an LIII-N multi-electron excitation in the data at this location (Hen- nig, 2007). The O shell could be fit as a single shell of oxygens at 2.29 A˚ (Table 1), which is generally consistent with 8-coordinated U(IV). Shells of neighboring atoms at distances beyond the O shells are expected to arise from strong metal binding functional groups present on biomass, i.e., carboxylate and phosphoryl (Kelly et al., 2001; Kelly et al., 2002; Senko et al., 2007), or neighboring shells of P or U atoms that may occur if solid phases such as uraninite or phosphates precipitated. Uraninite can be neglected because the characteristic FT peak at ca 3.8 A˚ (Schofield et al., 2008) is not present. Moreover, when the 2.9 A˚

Fig. 3. U LIII-edge XANES spectra for uranium precipitates formed (R + dR) FT peak was ﬁt with a U shell, then a distance in cell suspensions of strain UFO, and for reference materials of ca 3.2 A˚ was obtained, which is not physically realistic uraninite and meta-autunite. Dotted points show ﬁt results. (Burns et al., 1997; Catalano and Brown, 2004). Uranyl 2694 A.E. Ray et al. / Geochimica et Cosmochimica Acta 75 (2011) 2684–2695

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Submitted in Partial Fulfillment of the Requirements for the Degree of

The University of New Mexico Albuquerque, New Mexico

were sorbed. This study indicated that the presence of calcium in solution with carbonate will increase uranium sorption.

The effects of calcium in solution with carbonate on uranium biosorption were also observed in results obtained by N’Guessan et al. [68]. In their unpublished study, the group obtained an isotherm for uranium sorption in the Old Rifle site groundwater that has an average carbonate concentration of greater than 3 mM and an average calcium -1 concentration of 6.5 mM [70]. This isotherm resulted in a KD of approximately 57 L kg of protein for G. When compared to the partitioning coefficient obtained for G. uraniireducens in a bicarbonate solution void of calcium, 230 L kg-1 of protein for G, there is an approximate four time decrease in estimated KD values [68]. This indicates that uranium biosorption is decreased in the presence of calcium and carbonate in solution.

A difference in KD values was not the only observation noted during experiments focused on uranium sorption to biomass. The locations of sorbed uranium were noticed to be variable. It has been observed that uranium can sorb to the surface of bacteria [69, 71] or inside the cell membrane of the bacteria [72, 73]. While uranium has been seen in the membrane of bacteria that are capable of enzymatic reduction, it has also been seen within the membrane of fungi [58, 74, 75], which do not perform reduction. Uranium sorption within the cell, as well as on the cell surface, may make it difficult to accurately determine partitioning coefficients and to directly compare the surface sorption affinity of the different types of bacteria.

As summarized above, a significant amount of research has been performed, investigating uranium sorption to biomass. While this work provides supporting data for the occurrence of uranium sorption to the surface of biomass, the results obtained from these studies cannot be directly correlated to predict uranium sorption to biomass under the condition of the Old Rifle site aquifer. Many of the studies were done at a pH lower than that found in the aquifer at the Old Rifle site [64-67]. Of the studies that were performed at the proper pH range, only those done by N’Guessan et al. [68] and Gorman- Lewis et al. [69] captured the effects of calcium and DIC in solution; although, the levels of DIC were lower than that found at the Old Rifle site. Also, the study by N’Guessan et

71. Kelly, S.D., Boyanov, M.I., Bunker, B.A., Fein, J.B., Fowle, D.A., Yee, N. & Kemner, K.M. (2001). XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. Journal of Synchrotron Radiation, 8, pp. 946-948. 72. Senko, J.M., Kelly, S.D., Dohnalkova, A.C., McDonough, J.T., Kemner, K.M., & Burgos, W.D. (2007). The effect of U(VI) bioreduction kinetics on subsequent reoxidation of biogenic U(IV). Geochimica Et Cosmochimica Acta, 71 (19), pp. 4644-4654. 73. Shelobolina, E.S., Maddalena, V.C., Korenevsky, A.A., DiDonato, L.N., Sullivan, S.A., Konishi, H., Hu, H., Leang,C., Butler, J.E., Kim, B., & Lovley, D.R. (2007). Importance of c-Type cytochromes for U(VI) reduction by Geobacter sulfurreducens. BMC Microbiology, 7 (16). 74. Liu, M.X., Dong, F.Q., Yan, X.Y., Zeng, W.M., Hou, L.Y., & Pang, X.F. (2010). Biosorption of uranium by Saccharomyces cerevisiae and surface interactions under culture conditions. Bioresource Technology, 101 (22), pp. 8573-8580. 75. Sakamoto, F., Ohnuki, T., Kozai, N., Fujii, T., Lefuji, H., & Fancis, A.J. (2005). Effect of Uranium (VI) on the Growth of Yeast and Influence of Metabolism of Yeast on Adsorption of U(VI). Journal of Nuclear and Radiochemical Sciences, 6 (1), pp. 99-101. 76. Ranville, J.F. and Stucker, V (2009). Ion concentrations of Well B1 at Old Rifle. Personal communication. 77. Guillaumont, R., Fanghanel, T., Neck, V., Fuger, J., Palmer, D.A., Grenthe, I., & Rand, M.H. (2003). Chemical thermodynamics 5, Update on the chemical thermodynamics of uranium, neptunium, plutonium, americium and technetium. Amsterdam: Elsevier. 78. National Institute of Standards and Technology (NIST) (2004). Critically selected stability constant of metal complex database (Ver. 8.0). Gaithersburg, MD: U.S. Department of Commerce. 79. Turner, D.R. & Sassman, S.A. (1996). Approaches to sorption modeling for high- level waste performance assessment. Journal of Contaminate Hydrology, 21 (1- 4), pp. 311-332. 80. Stumm, W. & Morgan, J.J. (1996). Aquatic chemistry. Hoboken, NJ: John Wiley & Sons. 81. Snoeyink, V.L. & Jenkins, D.J. (1980). Water chemistry. Hoboken, NJ: John Wiley & Sons. 82. Kohler, M., Curtis, G.P., Meece, D.E., & Davis, J.A. (2004). Methods for estimating adsorbed uranium(VI) and distribution coefficients of contaminated sediments. Environmental Science and Technology, 38 (1), pp. 240-247 83. Yabusaki, S.B. (2009). Conversation about sorption density. Personal communication.

108

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Chemical Geology

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Complexation of carboxyl groups in bacterial lipopolysaccharides: Interactions of H+, 2+ 2+ 2+ 2+ Mg ,Ca ,Cd , and UO2 with Kdo and galacturonate molecules via quantum mechanical calculations and NMR spectroscopy

P. Selvarengan a, J.D. Kubicki b,⁎, J.-P. Guégan c, X. Châtellier a a Université de Rennes 1, CNRS, UMR 6118, Géosciences Rennes, Campus de Beaulieu, F-35042 Rennes Cedex, France b Department of Geosciences and the Earth & Environmental Systems Institute, The Pennsylvania State University, University Park, Pennsylvania 16802, USA c Ecole Nationale Supérieure de Chimie de Rennes, CNRS, UMR 6226, Campus de Beaulieu, F-35042 Rennes Cedex, France article info abstract

Article history: Quantum chemical calculations were combined with acid/base titrations and 13C NMR experiments to Received 15 September 2009 determine the structures, Gibbs free energy changes, vibrational frequencies and NMR chemical shifts Received in revised form 5 February 2010 + 2+ 2+ 2+ 2+ characterizing the interactions of H ,Mg ,Ca ,Cd and UO2 with the carboxyl group of the 2-Keto-3- Accepted 9 February 2010 deoxyoctanoate (Kdo) and galacturonate molecules as models for reactive sites on bacterial surfaces. Editor: J.D. Blum Monodentate, bidentate and outer-sphere complexation modes were considered. Results are also consistent with previous experimental and theoretical work on metal–carboxyl complexation, in particular with EXAFS 2+ Keywords: studies. Among the four metals considered here, the theoretical results suggest that Mg complexation is 2+ Kdo most thermodynamically favorable and UO2 complexation is least favorable. On the other hand, our NMR Galacturonic acid experimental results suggested that Cd2+ complexation is the most favorable, whereas Mg2+ and Ca2+ Metal complexation complexation are weak processes, hindered by competition with H+. Apart from this discrepancy, there was Quantum chemical calculations a reasonable agreement between the theoretical and experimental results, which, taken together, also Gibbs free energy indicated that outer-sphere complexation was the dominant mode of complexation in our systems. 13C NMR © 2010 Elsevier B.V. All rights reserved.

1. Introduction tively evaluate and describe metal sorption onto bacterial cell walls (Fein et al., 1997; Daughney and Fein, 1998; Langley and Beveridge, Bacterial surfaces affect the fate of metals in the near-surface 1999; Suzuki and Banfield, 1999; Lovley, 2000; Texier et al., 2000; geologic systems through adsorption reactions (Suzuki and Banfield, Kelly et al., 2001; Kelly et al., 2002; Martinez et al., 2002; Panak et al., 1999; Lovley, 2000; Dynes et al., 2006). The extent of adsorption of 2002; Boyanov et al., 2003; Ngwenya et al., 2003; Borrok et al., 2004; aqueous metals onto bacterial cell walls can vary markedly with Chatellier and Fortin, 2004; Yee et al., 2004; Gorman-Lewis et al., changing conditions such as pH, ionic strength and fluid composition 2005; Takahashi et al., 2005; Burnett et al., 2006; Dynes et al., 2006; (Fein et al., 1997; Daughney and Fein, 1998; Martinez et al., 2002; Gorman-Lewis et al., 2006; Guine et al., 2006). Ngwenya et al., 2003; Borrok et al., 2004; Chatellier and Fortin, 2004). Bacterial lipopolysaccharides (LPS) are the major constituent of These surface characteristics are controlled by the proton-active the outer membrane of the cell wall of Gram-negative bacteria (Raetz, functional groups found on the cell walls. The adsorption of aqueous 1990; Mozes et al., 1991; Madigan et al., 2003). LPS are believed to be metal cations onto bacterial surfaces has been extensively studied highly reactive with aqueous metal cations, and to contribute to the using various laboratory approaches. In the early 1980s, Beveridge sorption capacity of the bacteria towards metals, which affects metal and Murray (1980) and Beveridge et al. (1982) examined the mobility in the subsurface (Suzuki and Banfield, 1999; Lovley, 2000; complexation of several metals onto Gram-positive bacterial cell Dynes et al., 2006). LPS are polysaccharide chains composed of a lipid walls having undergone different types of chemical modifications to A, a core and in general an O-antigen, with variations depending on neutralize selectively the different kinds of reactive groups. They the kind of bacterial species. The surface of a LPS molecule carries a found that phosphate and carboxyl groups associated with the negative charge owing to phosphate, pyrophosphate and carboxylate molecules of the cell walls played a crucial role in the complexation groups located predominantly in the lipid A and the inner part of the of most metals. Later on, many studies were performed to quantita- core. The metal-binding properties of LPS are related to those reactive groups; however, the molecular details of the relevant interactions have not been elucidated. The structural complexity, intrinsic ⁎ Corresponding author. Tel.: +1 814 865 3951; fax: +1 814 863 7823. heterogeneity and different solubility properties of LPS have contrib- E-mail address: [email protected] (J.D. Kubicki). uted to slow progress in LPS structural determination by traditional

0009-2541/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.chemgeo.2010.02.012 56 P. Selvarengan et al. / Chemical Geology 273 (2010) 55–75 chemical methods (Schnaitman and Klena, 1993; Muller-Loennies et Both experimental (mass spectrometry, FT-IR, and NMR) and al., 2003; Kim et al., 2005). theoretical (i.e., quantum mechanics and classical Monte Carlo and Earlier studies have shown that the Kdo acid (3-deoxy-D-manno- molecular dynamics) techniques have been employed to deter- octulosonic acid, C8O8H14, Fig. 1A) and galacturonic acid (C6H10O7, mine the structure of LPS and of their sugar and fatty acid Fig. 1B) are important reactive components of LPS (Strain et al., 1983; components ( Kohlbrenner and Fesik, 1985; Jung et al., 1996; Lins and Straatsma, 2001; Bystrova et al., 2002; Nikaido, 2003). Both Frecer et al., 2000; Lins and Straatsma, 2001; Bystrova et al., 2002; are acidic sugars containing a carboxylic functional group. Kdo is Caroff and Karibian, 2003; Muller-Loennies et al., 2003; Kim et al., usually present in the core, an oligosaccharide, which serves as the 2005). However, limited attention has been given to determine the bridging unit between the complex O-antigenic polysaccharide found metal-binding properties of their monomeric residues (Malovikova in wild-type bacteria and the hydrophobic lipid A moiety that anchors et al., 1994; Bazin et al., 1995; Kondoh and Oi, 1997; Quiles and LPS in the bacterial outer membrane. Galacturonic acid is sometimes Burneau, 1998; Kelly et al., 2001; Saladini et al., 2001; Saladini et found in the O-antigen polysaccharide. The anionic character of Kdo al., 2002a; Saladini et al., 2002b; Inomata et al., 2004; Lu and Guo, and galacturonic acid residues in the LPS is thought to be important in 2006; Mani et al., 2006). The major difficulty in understanding the sequestering the divalent metal ions that are needed to stabilize the metal-binding of LPS, results from their polyelectrolytic properties, outer membrane. polyfunctionality and capacity to change their 3-D conformation. In the last decade, many studies have been performed using The relative interplay between each of these effects depends surface complexation models (SCM), which aim at predicting mainly on the physicochemical properties of the external medium quantitatively bulk-partitioning processes through a description of including the pH, the ionic strength, and the metal-to-acidic site the properties of each type of reactive site (Fein et al., 1997; Daughney ratio (Fein et al., 1997; Daughney and Fein, 1998; Chatellier and and Fein, 1998; Martinez et al., 2002; Ngwenya et al., 2003; Borrok et Fortin, 2004). al., 2004; Chatellier and Fortin, 2004). Generally, acid/base titrations Strain et al. (1983) used NMR spectroscopy to demonstrate that are used to determine the concentrations and acidity constants of the phosphate moieties in LPS molecules are reactive sites for the different kinds of sites. Metal sorption experiments are then used to complexation of several metals. They also found that the complexa- assess which sites are involved in metal complexation and to quantify tion of the metals with the LPS contributes to modify their the related complexation constants. However, it is now known that conformation. Similar results were observed via attenuated total experimental metal sorption isotherms are usually insufficient to reflectance FTIR spectroscopy (Parikh and Chorover, 2007). However, constrain properly the SCMs. For instance, the work of Chatellier and Strain et al. (1983) could not conclusively determine whether Fortin (2004) suggested that sorption isotherms typically define a carboxyl groups were involved as well in metal complexation lower limit for the pKa of the reactive groups, but in general they processes. Malovikova et al. (1994) compared the interactions of the 2+ 2+ cannot easily be used to define a precise value for this pKa value. Mg and Ca cations with polygalacturonic acid. They found that, Hence, sorption isotherms can usually be described by assuming that while Mg2+ complexation was electrostatic, Ca2+ cations were able to the reactive groups are sufficiently different and that compensating coordinate with two carboxyl groups and to induce inter-chain for the different pKa values can be done by adjusting the complexation associations. More recently, Synytsya et al. (2004) studied the constants. complexation of Co2+,Cu2+,Ni2+, and VO2+ with monomers or To solve this problem, new approaches have been used to dimers of D-galacturonic acid in the solid state, using by NMR and investigate metal sorption onto bacterial cells (Langley and Beveridge, vibrational spectroscopy techniques, and they were able to describe 1999; Texier et al., 2000; Kelly et al., 2001; Kelly et al., 2002; Panak et the complexation modes in some detail. However, the complexation al., 2002; Boyanov et al., 2003; Yee et al., 2004; Gorman-Lewis et al., of metal cations with either galacturonic or Kdo acids in solution has 2005; Takahashi et al., 2005; Burnett et al., 2006; Gorman-Lewis et al., not been investigated. 2006; Guine et al., 2006). For instance, EXAFS spectroscopy has As discussed above, spectroscopic and other experimental methods proved to be a powerful tool to investigate the nature of the reactive are useful to investigate metal sorption onto bacterial cells to identify groups involved in metal complexation at different pH values (Kelly et site preferences. However, spectra may be difficult to interpret due to al., 2001; Kelly et al., 2002; Panak et al., 2002; Boyanov et al., 2003; the lack of constraints on the molecular mechanisms involved in the Burnett et al., 2006; Guine et al., 2006). EXAFS studies have shown complexation of the metals with the different types of reactive groups that carboxyl and phosphate/phosphoester moieties are usually the present on the cell surface. With this in mind, computational chemistry dominant reactive groups in the cell walls of different bacteria. can be a powerful tool in helping to interpret spectra. However, in order to describe precisely the speciation of metals in Classical molecular modeling studies have been reported for LPS contact with bacterial surfaces, there is still a strong need to constrain molecules, Gram-positive or Gram-negative cell walls (Jung et al., further the experimental data available, using various complementary 1996; Bhattacharjya et al., 1997; Obst et al., 1997; Frecer et al., 2000; approaches. For instance, as bacterial cell walls have a complex Lins and Straatsma, 2001; Johnson et al., 2006). They have suggested macromolecular architecture, it is interesting to examine the role of that Ca2+ cations were essential to stabilize the outer membrane of smaller yet significant components of their general structure. Gram-negative bacterial cells through interactions with phosphate

Fig. 1. Structures of (A) Kdo acid, (B) galacturonic acid, (C) Kdo with methyl groups. 70 P. Selvarengan et al. / Chemical Geology 273 (2010) 55–75

Table 14 Experimental and theoretical metal–oxygen and metal–carbon distances (in Å) in various previous studies examining metal complexation with carboxyl groups.

Mg–OCa–OCd–OU–OCd–C1 U–C1 (M) (B) (M) (B) (M) (B) (M) (B) (M) (B) (M) (B)

2.08a 2.18b 2.38±0.15c 2.53±0.15c 2.45±0.15g 2.45±0.15g 2.22m 2.39m 3.13±0.06j 2.70j,k,l 3.40n 2.89j 2.04b – 2.41±0.01d – 2.3h,i 2.4i 2.35±0.03n 2.46±0.01j 3.19±0.10k – 3.50o 2.88±0.03n ––2.39e – 2.28j 2.28k 2.39±0.01o 2.47±0.03n –––2.77o ––2.32f – 2.29±0.03k 2.29–2.30l – 2.41±0.02o ––––

a Gao et al. (2004). b Davies et al. (2007). c Einspahr and Bugg (1981). d Thanomkul et al. (1976). e Alagna et al. (1986). f Sepulcre et al. (2004). g Glusker (1995). h Inomata et al. (2004). i Stamatatos et al. (2003). j Kelly et al. (2001, 2002). k Burnett et al. (2006). l Boyanov et al. (2003). m Schlosser et al. (2006). n Jiang et al. (2002). o Moll et al. (2003).

Strathmann and Myneni, 2004; Guan et al., 2006; Schlosser et al., Schlosser et al. (2006) studied the properties of uranyl mono- 2006; Davies et al., 2007; Kalinichev and Kirkpatrick, 2007). For Ca2+– carboxylates, using density functional theory. They calculated the U–O 2+ + carboxylate complexation, it has been discussed that Ca has a distance for the mono- and bidentate HCOO–UO2 complexes as 2.22 strong tendency to lie near the plane of the carboxyl group at a and 2.39 Å, respectively (Table 14). In our study, the monodentate distance from the O in the 2.3–2.6 Å range, with different possible U–O bond distances were found to be equal 2.35, 2.24, and 2.17 Å, for geometries (Einspahr and Bugg, 1981; Glusker, 1995). In general, the monodentate Kdo, galacturonic acid and carbonates respectively monodentate complexation was found to lead to Ca–O distances on (Tables 1–3). The bidentate U–O distances varied between 2.40 and + the order of 2.38±0.15 Å whereas bidentate complexation led to 2.47 Å for all the complexes, except the U–OII distance of the KdoUO2 larger Ca–O distances, on the order of 2.53±0.15 Å (Einspahr and complex, which was equal to 2.60 Å. Hence, our U–O distances for the + Bugg, 1981). Distances of about 2.40 and 2.42 Å have been reported as monodentate and bidentate Gal–UO2 and CO3UO2 complexes are the shortest Ca–O distances in a CaNa–galacturonate hexahydrate within 0.1 Å of those predicted by Schlosser et al. (2006) for the + crystal (Thanomkul et al., 1976). Similarly Alagna et al. (1986) HCOO–UO2 monodentate and bidentate complexes, but our U–O + performed an EXAFS study on Ca-poly(α-D-galacturonate) solid. The distances for the Kdo–UO2 complexes are slightly higher. shortest Ca–O distance was equal to 2.39 Å. More recently, Sepulcre et Our bidentate U–O bond distances are close to those obtained in al. (2004) examined the complexation in solution of Ca2+ with the three earlier EXAFS studies on uranyl–carboxyl complexes (Jiang et carboxyl groups present in the bacteriorhodopsin protein using al., 2002; Kelly et al., 2002; Moll et al., 2003). Kelly et al. (2002) XANES. They measured a Ca–O distance equal to 2.32 Å. All these examined the bidentate complexation in solution of uranyl with results are comparable to our predictions for the complexation of Ca2+ acetate and with Bacillus subtilis bacterial cells. Jiang et al. (2002) to the Kdo and galacturonate molecules, i.e., Ca–O distances equal to studied the complexation of uranyl with mono, bi and triacetates at 2.30–2.34 Å for monodentate complexation in both molecules and to various temperatures. Moll et al. (2003) examined uranium(VI)– 2.40–2.45 Å for Kdo bidentate complexation (Tables 1, 2 and 14). glycolate, –α-hydoxyisobutyrate, –α-aminoisobutyrate complexes. In AMg–O distance of 2.08 Å (Gao et al., 2004) has been reported for all three studies, bidentate U–O distances of 2.45±0.05 Å were found. a tetraaquabis(4-carboxyphenoxyacetato)-magnesium(II) crystal. Slightly smaller distances of 2.36±0.05 Å were obtained if the Similarly, Davies et al. (2007) recently performed crystallographic complexes were monodentate (Jiang et al., 2002; Moll et al., 2003). structural studies on magnesium carboxylate building units. They In addition to M–O distances, M–C distances have also been obtained average Mg–O bond distances of 2.04 and 2.18 Å for mono examined. In several EXAFS studies, including a few studies on metal and bidentate complexation, respectively. These distances are sorption onto bacterial cells, the M–C distance has been used as an comparable to our results, which range from 2.04 to 2.12 Å and indicator of the coordination mode. Cd2+ sorption was examined onto from 2.11 to 2.29 Å for monodentate and bidentate complexation two Gram-positive bacterial species, i.e., B. subtilis (Kelly et al., 2001; respectively (Tables 1–3 and 14). Boyanov et al., 2003) and Anoxybacillus flavithermus (Burnett et al., With respect to Cd2+, Glusker (1995) discussed that the metal–O 2006). In both cases, acetate was used as a standard. Cd2+ formed a distance in a Cd–carboxyl complex should fall in the 2.3–2.6 Å range, neutral complex with two acetate molecules with a Cd–C distance as for Ca2+. For the complexation of Cd2+ onto piperidine, Inomata et equal to 2.70 Å (Kelly et al., 2001; Boyanov et al., 2003; Burnett et al., al. (2004) obtained a Cd–O distance of 2.3 Å. Similarly Stamatatos et 2006). According to our results (Tables 1 and 2), this suggests that the al. (2003) studied the structure of Cd2+–carboxylate crystals. They Cd2+ was coordinated in a bidentate mode to each acetate molecule. found that the Cd–O mono- and bidentate distances were equal to 2.3 However, the complexation of Cd2+ onto the bacterial surfaces and 2.4 Å (average of two Cd–O bond distances) respectively, which involved a single carboxyl group and was characterized by a Cd2+–C is comparable to our results (Tables 1, 2 and 14). EXAFS studies on distance in the 3.07–3.29 Å range (Kelly et al., 2001; Burnett et al., Cd2+–acetate complexes and on Cd2+ sorption onto Gram-positive 2006). According to Tables 1 and 2, this is consistent with a mono- bacteria were also performed that resulted in Cd–Odistancesof dentate complexation where only one O atom of the carboxyl group 2.27–2.30 Å for the O atoms of the carboxyl groups coordinated with was coordinated with the Cd2+ cation. the Cd2+ cations (Kelly et al., 2001; Boyanov et al., 2003; Burnett et The complexation of uranyl with B. subtilis and Bacillus sphaericus al., 2006). has also been studied by EXAFS (Kelly et al., 2001; Kelly et al., 2002; 72 P. Selvarengan et al. / Chemical Geology 273 (2010) 55–75 hydration reactions of Mg2+,Ca2+,Sr2+ and Ba2+. They discussed the free energies of the free ligand and free metal were constant, and only influence of the inner- and outer-shell hydration spheres of H2O the Gibbs free energy of the metal–ligand complex changed. In molecules in the hydration reactions. They found that the thermo- addition, systematic errors inherent to the approximations of our chemical properties of the cations were significantly changed by the approach probably largely affected to the same extent the free presence of H2O molecules surrounding them. energies of such closely related complexes. For example, Kubicki 2+ 2+ Cd and UO2 cations have been observed to associate with (2001) analyzed the accuracy of the supermolecule/continuum 3+ carboxyl groups on Gram-positive and Gram-negative bacterial approach for the pKa calculations of aqueous metals such as Al , surfaces (Fein et al., 1997; Fowle et al., 2000; Haas et al., 2001; Kelly Fe3+ and Si4+ cations and found that the correlations of deprotona- et al., 2001; Boyanov et al., 2003; Yee and Fein, 2003; Borrok and Fein, tion energies with observed ln(Ka) values were good for individual 2+ 2005; Burnett et al., 2006). For UO2 , this seems to represent a cations. However, our results were still too imprecise to predict 2+ discrepancy with our theoretical results: according to Table 8,UO2 accurately the order of favorability of the complexation modes or of should indeed not form complexes with carboxylate groups because the metals. For instance, we obtained a lower ΔG value for the of the positive predicted Gibbs free energy changes. However, this monodentate than for the outer-sphere Cd2+–galacturonate complex. needs to be confirmed by the study of larger systems. For instance, the We also predicted that complexation was more favorable with Mg2+ 2+ 2+ possibility of complexes involving a UO2 cation and two carboxyl than with Cd , whereas our NMR experiments indicate the opposite. 2+ groups, or a UO2 cation, one carboxyl group, and another type of This indicates that relative errors between the ΔG values given in ligand, such as a phosphate moiety, should also be investigated. Table 9 are equal to about ±40 kJ/mol. Most experimental studies of metal association with carboxyl One problem with the current approach to calculating the ΔG groups as well as the theoretical study of Johnson et al. (2006) values is that the positions of the H2O molecules in the system do not discussed above were investigating metal sorption onto Gram- adequately explore configuration space. For instance, we tested the 2+ positive bacteria. The architecture of Gram-positive bacteria signifi- effect of moving one H2O molecule in the Cd –galacturonate outer- cantly differs from that of Gram-negative bacteria. Some Gram- sphere complex from near the galacturonate towards the Cd2+ end of positive bacteria possess teichuronic acids. Those polymers contain the model and this decreased the calculated ΔG by −13 kJ/mol. uronic acids, which are closely related to galacturonic acid (e.g. Mozes Hence, energy minimizations such as those that were done in this et al. (1991)). However, in general, most of the carboxyl groups study should be combined with molecular dynamics simulations of present in the cell walls of Gram-positive bacteria are located on the the models in order to obtain a better sampling of configuration space amino acids oligomers of the peptidoglycan. In addition, the cell walls and more accurate thermodynamics. The level of explicit H2O of Gram-positive as well as Gram-negative bacteria harbor various molecules solvating the molecules could also be increased in order kinds of proteins, which are also made of amino acids and possibly to create a full second solvation sphere. provide reactive carboxyl groups different from those investigated here. 5. Conclusion With respect to the other cations (H+,Mg2+,Ca2+ and Cd2+), the NMR experiments performed in the presence of increasing concen- In conclusion, this paper presents quantum mechanical calcula- trations of cations provided interesting constraints on the ΔG values of tions of the structural, thermodynamical and vibrational parameters the reactions of complexation, that can be compared with our of metal–carboxyl complexes involving two molecules commonly theoretical calculations. With respect to the metals, the NMR found in the cell walls of Gram-negative bacteria, i.e. the Kdo and experiments suggested for instance that inner-sphere complexation galacturonate molecules. We have investigated the complexation of + 2+ 2+ 2+ 2+ was not occurring and that, at least for the galacturonate molecule, ΔG five different cations (H ,Mg ,Ca ,Cd , and UO2 ) of biological values were lowest for the Cd2+ cations and largest for the Ca2+ and environmental importance, and we have considered the three cations. main modes of possible complexation (monodentate, bidentate, and An estimate of the ΔG value of the Cd2+–galacturonate complex can outer-sphere). Experimental acid–base and NMR titrations were also be obtained from our NMR measurements. As shown on Fig. 6B, the δ13C performed to complement the theoretical calculations. Our results are values of the protonated and complexed carboxyl groups were equal to comparable to previous studies on metal–carboxyl complexation. about 172.5 and 175.5 ppm, respectively. For a Cd2+/carboxyl ratio of However, they will be particularly useful to future work involving 4, a δ13Cvalueof≈174 ppm was obtained, corresponding to the average metal complexation onto Gram-negative bacterial cells, using for of 172.5 ppm and 175 ppm and indicating that the concentrations of instance spectroscopic techniques such as EXAFS. protonated and metal complexed carboxyl groups were approximately We have shown that significant differences can be expected for the equal. It follows that, in this system where the pH was equal to 2.2 and same complexation mode of the same metal onto either the Kdo or the total Cd2+ concentration was about 35 mM, the complexation galacturonate molecules, even though both of those molecules are 2+ + −1.4 − 3.3 + 2.2 −2.5 constant was equal to K≈[Cd ]Ka /[H ]≈10 ≈10 . relatively similar, i.e., they are both sugars harboring a carboxyl group Using the analogue of Eq. (3) for metal complexation, one obtains attached at the same position on the sugar ring. This indicates that ΔG≈−14 kJ/mol. A similar calculation for the Cd2+–galacturonate future EXAFS experiments might be able to recognize not only the complex leads to ΔG≈−11 kJ/mol. Our NMR results thus suggest that mode of complexation but also which type of carboxyl groups is the the ΔG values of formation of the Cd2+–Kdo and Cd2+–galacturonate most reactive towards a given metal on a bacterial surface. complexes are similar and equal to about −10 to −15 kJ/mol. Quantum This study is also a basis for further calculations using molecular calculations result in ΔG values of +3 to −38 kJ/mol for Cd2+–Kdo and dynamics methods to build up a better understanding of metal +17 to −16 kJ/mol for Cd2+–galacturonate. The discrepancy with the complexation onto larger molecules such as lipopolysaccharides. The theoretical results are thus equal to at most about 25 kJ/mol for the Cd2 complexation of a metal with two different ligands will also have to be +–Kdo and about 10 kJ/mol for the Cd2+–galacturonate complexes considered in future work. which is within the expected accuracy of these types of calculations. The most important conclusion of this study is that based on Hence these comparisons suggest that the imprecision of the observed 13C NMR spectra and calculations of 13C chemical shifts in approach used here on the values of the Gibbs free energy changes various complexes outer-sphere complexes dominant in this system. was on the order of ±25 kJ/mol. However, the relative differences Although accuracy is limited, the ΔG values computed from quantum between the ΔG values of related complexes should be predicted with mechanics suggest that outer-sphere complexation is competitive better accuracy. For instance, when comparing the three different with and sometimes more thermodynamically favorable than inner- modes of complexation of a given metal with a given ligand, the Gibbs sphere complexation. A significant caveat is that we have only 74 P. Selvarengan et al. / Chemical Geology 273 (2010) 55–75

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Biodegradation (2010) 21:81–95 DOI 10.1007/s10532-009-9283-x

ORIGINAL PAPER

Can microbially-generated hydrogen sulﬁde account for the rates of U(VI) reduction by a sulfate-reducing bacterium?

Benjaporn Boonchayaanant Æ Baohua Gu Æ Wei Wang Æ Monica E. Ortiz Æ Craig S. Criddle

Received: 12 February 2009 / Accepted: 25 June 2009 / Published online: 14 July 2009 Springer Science+Business Media B.V. 2009

Abstract In situ remediation of uranium contami- had no clear effect at 15 mM. During the hydrogen nated soil and groundwater is attractive because a sulfide-mediated reduction of U(VI), a floc formed diverse range of microbial and abiotic processes containing uranium and sulfur. U(VI) sequestered in reduce soluble and mobile U(VI) to sparingly soluble the floc was not available for further reduction. and immobile U(IV). Often these processes are linked. Sulfate-reducing bacteria (SRB), for example, Keywords Uranium reduction enzymatically reduce U(VI) to U(IV), but they also Sulfate-reducing bacteria Hydrogen sulfide produce hydrogen sulfide that can itself reduce Growth kinetics Ferrous iron U(VI). This study evaluated the relative importance of these processes for Desulfovibrio aerotolerans,a SRB isolated from a U(VI)-contaminated site. For the List of symbols conditions evaluated, the observed rate of SRB- A Acetate concentration predicted from mediated U(VI) reduction can be explained by the predicted the stoichiometric ratio of acetate abiotic reaction of U(VI) with the microbially- production to ethanol consumption generated H S. The presence of trace ferrous iron 2 (mM) appeared to enhance the extent of hydrogen sulfide- A Acetate concentration at the initial time mediated U(VI) reduction at 5 mM bicarbonate, but 0 (mM) C Covariance matrix

faq Fraction of mole of hydrogen sulfide in B. Boonchayaanant C. S. Criddle (&) the aqueous phase Department of Civil and Environmental Engineering, faq,i-1 Fraction of mole of hydrogen sulfide in Stanford University, 473 Via Ortega, Yang and Yamazaki the aqueous phase on day i - 1 Environment and Energy Building 151, Stanford, f Fraction of mole of hydrogen sulfide in CA 94305, USA g e-mail: [email protected] the gas phase

fH2S;aq Fraction of mole of H2S in the total B. Gu W. Wang mole of hydrogen sulﬁde in the Environmental Sciences Division, Oak Ridge National H2SðaqÞ aqueous phase = Laboratory, Oak Ridge, TN, USA H2SðaqÞþHS aq J Jacobian matrix ð Þ M. E. Ortiz k Pseudo second-order rate constant for Department of Bioengineering, Stanford University, ethanol utilization (l/mg protein/day) Stanford, CA, USA 123 Biodegradation (2010) 21:81–95 93

At 5 mM of bicarbonate, the extent of U(VI) elemental sulfur formed during the reaction. This issue reduction was greater in the presence of Fe(II) or merits further investigation given the importance of trace metals. The pattern kinetics of uranium reduc- sulfate reduction in bioreduced systems used for tion in the presence of FeCl2 at 5 mM bicarbonate uranium remediation (Wu et al. 2006b). was similar to the pattern observed in the presence of the complete trace metal solution, suggesting that Fe(II) was responsible for enhanced U(VI) reduction Conclusion at 5 mM bicarbonate. At 15 mM bicarbonate, however, there was little effect of FeCl and trace metals 2 Hydrogen sulfide-mediated abiotic U(VI) reduction is addition. A possible explanation is that uranyl species the dominant reduction mechanism for the conditions at 5 and 15 mM bicarbonate differ in their suscep- evaluated. At 5 mM bicarbonate, addition of trace tibility to reduction by H S and/or Fe(II). Moreover, 2 levels of ferrous iron increased the extent of U(VI) given that the redox couples for Fe(II)/(III) and reduction in the presence of hydrogen sulfide, but at U(IV)/U(VI) are similar, it can be expected that the 15 mM bicarbonate, the effect of ferrous iron addition direction of the reaction is highly dependent on was unclear. At 5 mM bicarbonate, a large fraction of relative concentrations and chemical forms. the U(VI) removed from the aqueous phase became Previous studies have shown that Fe(II) and Fe(II) associated with a sulfur-containing white/gray floc minerals, such as green rusts and amorphous iron preventing further U(VI) reduction. The interactions sulfide (FeS), can reduce U(VI) at neutral pH though among U(VI), hydrogen sulfide, Fe(II), biomass, and surface catalysts were required under certain condi- solid phases that were observed in this study are also tions (Liger et al. 1999; O’Loughlin et al. 2003; likely to exist in actual environments where typical Behrends and Van Cappellen 2005; Hua and Deng bicarbonate concentrations are low (\5 mM). Future 2008). In our study, trace Fe(II) enhanced U(VI) studies should, therefore, take into consideration the reduction in the presence of hydrogen sulfide, but potential for such reactions, and their effects on the trace Fe(II) by itself did not. The formation of fate and transport of uranium in the subsurface. dissolved FeS may have occurred (saturation levels were not exceeded), and it may have reduced U(VI) Acknowledgments This work was funded by the Environ- at a faster rate than hydrogen sulfide itself. Alterna- mental Remediation Science Program (ERSP), U.S. Department tively, ferrous iron formed by sulfide reduction of of Energy, under grant number DOEAC05-00OR22725. We Fe(III) may have reduced U(VI). Hydrogen sulfide thank two anonymous reviewers for thoughtful reviews and recommendations that significantly improved the manuscript. might regenerate Fe(II) after its oxidation to Fe(III). Tests of the floc that formed after 4–5 h at 5 mM bicarbonate revealed a large fraction of U(VI) associated with the white/gray floc. This U(VI) was References not accessible for reduction. The observation that U(VI) reduction was faster in the absence of biomass Anderson RT, Lovley DR (2002) Microbial redox interactions than in its presence was also unexpected and at first with uranium: an environmental perspective. In: Keith- appears counter-intuitive, but can be explained by a Roach MJ, Livens FR (eds) Interactions of microorgan- similar mechanism: partitioning into an inaccessible isms with radionuclides. Elsevier, Oxford, pp 205–223 Anderson RT, Vrionis HA, Ortiz-Bernad I, Resch CT, Long phase which may include the biomass itself. Sorption PE, Dayvault R, Karp K, Marutzky S, Metzler DR, Pea- of U(VI) to biomass has been reported in previous cock A, White DC, Lowe M, Lovley DR (2003) Stimu- studies (Haas et al. 2001; Bencheikh-Latmani and lating the in situ activity of geobacter species to remove Leckie 2003; Kelly et al. 2001). At 15 mM, less floc uranium from the groundwater of a uranium-contaminated aquifer. Appl Environ Microbiol 69:5884–5891 formed, and the extent of transformation was some- Behrends T, Van Cappellen P (2005) Competition between what greater. enzymatic and abiotic reduction of uranium (VI) under Given that floc formation and concomitant cessation iron reducing conditions. Chem Geol 220:315–327 of U(VI) reduction occurred in the absence of phos- Bencheikh-Latmani R, Leckie JO (2003) Association of uranyl with the cell wall of Pseudomonas fluorescens inhibits phate, we conclude that U(VI) reduction is likely self- metabolism. Geochim Cosmochim Acta 67(21):4057– limiting and due to the formation of floc containing 4066 123 94 Biodegradation (2010) 21:81–95

Benjamin M (2002) Water chemistry. McGraw-Hill, New bioremediation of U(VI): spatial similarities along con- York, p 668 trolled flow paths. ISME J. Published on-line September 4 Boonchayaanant B, Kitanidis PK, Criddle CS (2008) Growth Kelly SD, Boyanov MI, Bunker BA, Fein JB, Fowle DA, Yee and cometabolic reduction kinetics of a uranium- and N, Kemner KM (2001) XAFS determination of the bac- sulfate-reducing Desulfovibrio/Clostridia mixed culture: terial cell wall functional groups responsible for com- temperature effects. Biotechnol Bioeng 99(5):1107–1119 plexation of Cd and U as a function of pH. J Synchrotron Bradford MM (1976) A rapid and sensitive method for quan- Rad 8:946–948 titation of microgram quantities of protein utilizing the Kumar S, Tamura K, Nei M (2004) MEGA3: integrated soft- principle of protein-dye-binding. Anal Biochem 72:248– ware for molecular evolutionary genetics analysis and 454 sequence alignment. Brief Bioinform 5:150–163 Cardenas E, Wu W-M, Leigh MB, Carley J, Carroll S, Gentry Liger E, Charlet L, Van Cappellen P (1999) Surface catalysis T, Luo J, Watson D, Gu B, Ginder-Vogel M, Kitanidis of uranium(VI) reduction by iron(II). Geochim Cosmo- PK, Jardine PM, Zhou J, Criddle CS, Marsh TL, Tiedje chim Acta 63:2939–2955 JM (2008) Microbial communities in contaminated sedi- Lovley DR, Phillips EJP (1992a) Reduction of uranium by ments associated with bioremediation of uranium to sub- Desulfovibrio desulfuricans. Appl Environ Microbiol micromolar levels. Appl Environ Microbiol 74(12):3718– 58:850–856 3729 Lovley DR, Phillips EJP (1992b) Bioremediation of uranium Elias DA, Suflita JM, McInerney MJ, Krumholz LR (2004) contamination with enzymatic uranium reduction. Envi- Periplasmic cytochrome c3 of Desulfovibrio vulgaris is ron Sci Technol 26:2228–2234 directly involved in H2-mediated metal but not sulfate Lovley DR, Roden EE, Phillips EJP, Woodward JC (1993a) reduction. Appl Environ Microbiol 70(1):413–420 Enzymatic iron and uranium reduction by sulfate-reducing Francis AJ, Dodge CJ, Lu F, Halada GP, Clayton CR (1994) bacteria. Mar Geol 113:41–53 XPS and XANES studies of uranium reduction by Clos- Lovley DR, Giovannoni SJ, White DC, Champine JE, Phillips tridium sp. Environ Sci Technol 28:636–639 EJP, Gorby YA, Goodwin S (1993b) Geobacter metalli- Ganesh R, Robinson KG, Reed GD, Sayler GS (1997) reducens gen. nov. sp. nov., a microorganism capable of Reduction of hexavalent uranium from organic complexes coupling the complete oxidation of organic compounds to by sulfate- and iron-reducing bacteria. Appl Environ the reduction of iron and other metals. Arch Microbiol Microbiol 63(11):4385–4391 159:336–344 Gorby YA, Lovley YAG (1992) Enzymatic uranium precipi- McCullough J, Hazen TC, Benson SM, Metting FB, Palmisano tation. Environ Sci Technol 26:206–207 AC (1999) Bioremediation of metals and radionuclides: Gu B, Brooks SC, Roh Y, Jardine PM (2003) Geochemical what it is and how it works. Lawrence Berkeley National reactions and dynamics during titration of a contaminated Laboratory, California groundwater with high uranium, aluminum, and calcium. Nyman JL (2006) Community structure and kinetics of Geochim Cosmochim Acta 67:2749–2761 microbial uranium reduction for field-scale bioremedia- Gu BH, Wu WM, Ginder-Vogel MA, Yan H, Fields MW, tion. Dissertation, Stanford University, Stanford Fendorf ZJ, Criddle CS, Jardine PM (2005) Bioreduction Nyman JL, Marsh TL, Ginder-Vogel MA, Gentile M, Fendorf of uranium in a contaminated soil column. Environ Sci S, Criddle CS (2006) Heterogeneous response to biosti- Technol 39:4841–4847 mulation for U(VI) reduction in replicated sediment Guillaumont R, Fanghanel T, Neck V, Fuger J, Palmer DA, microcosms. Biodegradation 17:303–316 Grenthe I, Rand MH (2003) Update on the chemical Nyman J, Wu H-I, Gentile M, Criddle CS (2007) Inhibition of thermodynamics of uranium, neptunium, plutonium, a U(VI)- and sulfate-reducing consortia by U(VI). Environ americium, and technetium. OECD Nuclear Energy Sci Technol 41:6528–6533 Agency (ed) Elsevier, Amsterdam, The Netherlands, O’Loughlin EJ, Kelly SD, Cook RE, Csencsits R, Kemner KM p 919 (2003) Reduction of uranium(VI) by mixed iron(II)/ Haas JR, Dichristina TJ, Wade R Jr (2001) Thermodynamics of iron(III) hydroxide (green rust): formation of UO2 nano- U(VI) sorption onto Shewanella putrefaciens. Chem Geol particles. Environ Sci Technol 37:721–727 180:33–54 Phillips EJP, Landa ER, Lovley DR (1995) Remediation of He Z, Gentry TJ, Schadt CW, Wu L, Liebich J, Chong SC, uranium contaminated soils with bicarbonate extraction Huang Z, Wu W, Gu B, Jardine P, Criddle C, Zhou J and microbial U(VI) reduction. J Ind Microbiol 14:203–207 (2007) GeoChip: a comprehensive microarray for inves- Pietzsch K, Babel W (2003) A sulfate-reducing bacterium that tigating biogeochemical, ecological, and environmental can detoxify U(VI) and obtain energy via nitrate reduc- processes. ISME J 1:67–77 tion. J Basic Microbiol 43:348–361 Hua B, Deng B (2008) Reductive immobilization of ura- Pietzsch K, Hard BC, Babel W (1999) A Desulfovibrio sp. nium(VI) by amorphous iron sulfide. Environ Sci Technol capable of growing by reducing U(VI). J Basic Microbiol 42(23):8703–8708 39:365–372 Hua B, Xu H, Terry J, Deng B (2006) Kinetics of uranium (VI) Sani R, Peyton B, Smith W, Apel W, Peterson J (2002) Dis- reduction by hydrogen sulfide in anoxic aqueous system. similatory reduction of Cr(VI), Fe(III), and U(VI) by Environ Sci Technol 40:4666–4671 Cellulomonas isolates. Appl Microbiol Biotechnol Hwang C, Wu W-M, Gentry TJ, Carley J, Carroll SL, Watson 60:192–199 D, Jardine PM, Zhou J, Criddle CS, Fields MW (2008) Spear JR, Figueroa LA, Honeyman BD (1999) Modeling the Bacterial community succession during in situ removal of uranium U(VI) from aqueous solutions in the 123 ANRV374-EA37-16 ARI 27 March 2009 15:20

Microbial Transformations of Minerals and Metals: Recent Advances in Geomicrobiology Derived from Synchrotron- Based X-Ray Spectroscopy and X-Ray Microscopy

Alexis Templeton and Emily Knowles

Department of Geological Sciences, University of Colorado, Boulder, Colorado 80309-0399; email: [email protected], [email protected]

Annu. Rev. Earth Planet. Sci. 2009. 37:367–91 Key Words First published online as a Review in Advance on bioﬁlm, XANES, EXAFS, STXM, X-ray microprobe

by University of Notre Dame on 08/23/14. For personal use only. December 4, 2008

The Annual Review of Earth and Planetary Sciences is Abstract online at earth.annualreviews.org The field of geomicrobiology has embraced efforts to define and quantify This article’s doi: Annu. Rev. Earth Planet. Sci. 2009.37:367-391. Downloaded from www.annualreviews.org the role of microbial organisms in low-temperature geochemical processes. 10.1146/annurev.earth.36.031207.124346 However, any mechanistic understanding of the interactions between micro- Copyright c 2009 by Annual Reviews. bial organisms and minerals or metals requires analytical tools with the ap- All rights reserved propriate chemical sensitivity and spatial resolution. In the past several years, 0084-6597/09/0530-0367$20.00 increasing application of nanometer-, micron-, and bulk-scale synchrotron- based X-ray techniques has provided new insights regarding the feedbacks among microbial growth, mineral dissolution, redox transformations, and biomineralization processes. In this review, recent findings derived from X-ray absorption spectroscopy, X-ray microprobe mapping, and X-ray microscopy studies are integrated to provide a new view of the dynamic biogeochemistry occurring at the microbe-mineral interface.

367 ANRV374-EA37-16 ARI 27 March 2009 15:20

geochemistry. Numerous recent reviews summarize our rapidly developing view of how microbial organisms transfer electrons to and from mineral surfaces, using Fe as the most extensively studied geomicrobiological system (see Kappler & Straub 2005, Weber et al. 2006, Lovley 2008). K-edge: the binding Synchrotron-based studies have also elegantly demonstrated how coupled biotic-abiotic chemical energy that photons transformations must be studied to understand the stability and transformations of Fe minerals must exceed to ionize (e.g., Hansel et al. 2003). Toexplore fundamentally related processes, we use the following sections electrons from the K of this review to integrate the recent ﬁndings from ﬁve other areas of geomicrobiological research (e.g., 1s) shell that have increasingly relied on synchrotron-based spectroscopic and microscopic tools to unravel microbial transformations of minerals and metals.

Sorption to Microbial Surfaces The concentration and speciation of metals, nutrients, and trace elements in rocks, soils, and aquatic environments can be dramatically influenced by sorption reactions at cell surfaces. Natural biofilms sequester several metallic and alkaline elements. For example, biofilms often contain K+, Ca2+,Cr3+,Mn2+,Fe3+,Ni2+,Cu2+,Zn2+, and Pb2+ at concentrations 2 to 4 orders of magnitude greater than those measured in the surrounding bulk fluid (Brown et al. 1994, Ferris et al. 2000). The affinity of bacterial surfaces for aqueous cations is due in part to the low isoelectric points of the surfaces, resulting in a negative surface charge over a wide pH range. However, the structure and charge density of bacterial cell walls can vary greatly (Beveridge & Murray 1980). There appears to be a continuum between metal ion sorption and precipitation reactions (Warren & Ferris 1998) in which bacteria or acidic mucopolysaccharides provide nucleation sites for the formation of metal oxides, carbonates, phosphates, silicates, sulfides, and organometallic complexes (Degens & Ittekkot 1982, Beveridge et al. 1983, Ferris et al. 1987, McLean & Beveridge 1990, Thompson & Ferris 1990, Schultze-Lam et al. 1992, Southam et al. 1995, Urrutia & Beveridge 1995, Fortin & Beveridge 1997). The mechanism of metal ion precipitation is assumed to follow multiple stages: Nucleation sites are first formed by the stoichiometric sorption of metal ions to functional groups on the cell walls, followed by the association of appropriate counter ions 2− 3− (e.g., CO3 ,PO4 ) and then additional metal ion deposition (Urrutia & Beveridge 1993). Considerable macroscopic and molecular-level research has been directed toward understanding the modes and mechanisms of metal ion sorption reactions on bacterial surfaces. Numerous recent studies (e.g., Fein et al. 1997, Yee & Fein 2003) have used surface complexation modeling to quantify the site densities, deprotonation constants, and metal-binding constants of the functional by University of Notre Dame on 08/23/14. For personal use only. groups present on bacterial surfaces. The carboxyl and phosphoryl sites are thought to dominate metal binding at low to neutral pH, and hydroxyl and amino groups have been invoked for metal complexation above pH 8. However, the surface complexation modeling can only tentatively infer

Annu. Rev. Earth Planet. Sci. 2009.37:367-391. Downloaded from www.annualreviews.org the chemical identity of sorption sites from their acidity constants, and the modeling is also only sensitive to the dominant sites involved in metal binding. X-ray absorption spectroscopy can be used to independently determine the modes of metal ion binding to specific functional groups on bacterial surfaces. Metal-bacteria studies have primarily 2+ 2+ 2+ 2+ been conducted for Zn ,Pb ,Cd , and UO2 . Todate, most spectroscopic studies have been in agreement with the surface complexation modeling (Kelly et al. 2002, Boyanov et al. 2003). Using Zn2+ as an example, metal K-edge EXAFS have identified the same dominant metal-binding sites on several different types of bacterial surfaces. Guine et al. (2006) found that Zn2+ binding at the surfaces of several bacterial species was dominated by phosphoester groups at all Zn2+ loadings. Similar data showing primary Zn association with phosphoryl groups were previously observed for Pseudomonas putida biofilms by Toneret al. (2005b) and for fungal cell walls by Sarret et al. (1998). Guine et al. conducted experiments across a broad range of Zn2+ loadings, and the resulting spectra

www.annualreviews.org • Synchrotron-Based Geomicrobiology 371 ANRV374-EA37-16 ARI 27 March 2009 15:20

(II) (III) (IV) Mn oxidase Mn(IV)Mn(M IV)

Mn(II)II) Nanoscale hexagonalhexagona Mn O2 2+ oxide sheet polymepolymers Mn (aq) SG-1 ssporepore

+ Ca2+ + Mn2+ μ + Mn2+ ed absorbance ( >500 M)

z μ SG-1 ( <500 M) Biooxide Hexagonal Normali

δ -MnO2 Feitknechtite β-MnOOH Hydrated hexagonal Ca-bearing C-disordered 6530 6550 6570 6590 birnessite orthogonal birnessites Energy (eV)

Figure 4 Mn K-edge XANES for Mn(II) reacted with spores of Bacillus sp. SG-1. Over time, Mn(II) is biologically oxidized to Mn(IV), forming nanoparticulate hexagonal Mn(IV)-oxide minerals. These biogenic oxides will then transform to a variety of Mn-oxide structures, such as Ca-bearing orthogonal birnessites, hydrated hexagonal birnessites, or feitkneichtite, depending on environmental conditions such as Ca2+ concentration or Mn2+concentration. (Conceptual ﬁgure prepared with John Bargar to synthesize ideas presented in Bargar et al. 2005 and Webb et al. 2005b.)

reduction of U(VI) to U(IV), the same process that may have originally formed many uraninite mineral deposits (e.g., Lovley et al. 1991). Synchrotron-based X-ray spectroscopy techniques have elucidated the mechanisms of adsorption of uranyl ions and other soluble uranium species to cell surfaces. Using EXAFS, several different studies have shown that at low pH uranium binds primarily to surface phosphoryl groups, but with increasing pH more complexes form with carboxyl groups (e.g., Kelly et al. by University of Notre Dame on 08/23/14. For personal use only. 2001, Panak et al. 2002, Nedelkova et al. 2007). The spectra also reveal the coordination of uranyl ions with the functional groups on the cell surfaces, indicating monodentate phosphate complexes and bidentate carboxyl bonds (Kelly et al. 2002, Kemner et al. 2005, Merroun et al. 2005).

Annu. Rev. Earth Planet. Sci. 2009.37:367-391. Downloaded from www.annualreviews.org Sorption of U(IV) species to biogenic metal-oxide minerals is also critically important. EXAFS spectroscopy has been used to determine the coordination chemistry of U complexes associated with biomineralized manganese (Webb et al. 2006) and iron oxides (e.g., O’Loughlin et al. 2003, Jeon et al. 2004, Sani et al. 2004, Kemner et al. 2005) and to determine how sorbed U complexes can inhibit microbial reduction of the associated mineral phases (Brooks et al. 2003, Kelly et al. 2007). It is feasible to promote uranium immobilization by precipitation of U(VI) by bacterial bioﬁlms (Beyenal et al. 2004) and microbial mats (Bender et al. 2000). Microbial metabolic processes can reduce uranium either by directly using U(VI) as an electron acceptor, or by releasing metabolic by-products that promote reduction. Fredrickson et al. (2000) demonstrated that dissimilatory metal-reducing bacteria such as Shewanella putrefasciens CN-32 can utilize both direct and indirect mechanisms of U reduction. More recently, Marshall et al. (2006) were able to show that the

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Fredrickson JK, Zachara JM, Kennedy DW, Dong HL, Onstott TC, et al. 1998. Biogenic iron mineralization accompanying the dissimilatory reduction of hydrous ferric oxide by a groundwater bacterium. Geochim. Cosmochim. Acta 62:3239–57 Fredrickson JK, Zachara JM, Kennedy DW, Duff MC, Gorby YA, et al. 2000. Reduction of U(VI) in goethite (alpha-FeOOH) suspensions by a dissimilatory metal-reducing bacterium. Geochim. Cosmochim. Acta 64:3085–98 Gilbert ES, Khlebnikov A, Meyer-Ilse W, Keasling JD. 1999. Use of soft X-ray microscopy for analysis of early-stage biofilm formation. Water Sci. Technol. 39:269–72 Guine V, Spadini L, Sarret G, Muris M, Delolme C, et al. 2006. Zinc sorption to three gram-negative bacteria: Combined titration, modeling, and EXAFS study. Environ. Sci. Technol. 40:1806–13 Hansel CM, Benner SG, Neiss J, Dohnalkova A, Kukkadapu RK, et al. 2003. Secondary mineralization pathways induced by dissimilatory iron reduction of ferrihydrite under advective flow. Geochim. Cosmochim. Acta 67:2977–92 Jacobsen C, Wirick S, Flynn G, Zimba C. 2000. Soft X-ray spectroscopy from image sequences with sub-100 nm spatial resolution. J. Microsc. 197:173–84 Jeon BH, Kelly SD, Kemner KM, Barnett MO, Burgos WD, et al. 2004. Microbial reduction of U(VI) at the solid-water interface. Environ. Sci. Technol. 38:5649–55 Kappler A, Straub KL. 2005. Geomicrobiological cycling of iron. In Molecular Geomicrobiology, ed. JF Banfield, J Cervini-Silva, KH Nealson, Rev. Mineral. Geochem. 59:85–108. Washington, DC: Mineral. Soc. Am. Kelly SD, Boyanov MI, Bunker BA, Fein JB, Fowle DA, et al. 2001. XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J. Synchrotron Radiat. 8:946–48 Kelly SD, Kemner KM, Brooks SC. 2007. X-ray absorption spectroscopy identifies calcium-uranyl-carbonate complexes at environmental concentrations. Geochim. Cosmochim. Acta 71:821–34 Kelly SD, Kemner KM, Carley J, Criddle C, Jardine PM, et al. 2008. Speciation of uranium in sediments before and after in situ biostimulation. Environ. Sci. Technol. 42:1558–64 Kelly SD, Kemner KM, Fein JB, Fowle DA, Boyanov MI, et al. 2002. X-ray absorption fine structure determination of pH-dependent U-bacterial cell wall interactions. Geochim. Cosmochim. Acta 66:3855–71 Kemner KM, Kelly SD, Lai B, Maser J, O’Loughlin EJ, et al. 2004. Elemental and redox analysis of single bacterial cells by X-ray microbeam analysis. Science 306:686–87 Kemner KM, O’Loughlin EJ, Kelly SD, Boyanov MI. 2005. Synchrotron X-ray investigations of mineral- microbe-metal interactions. Elements 1:217–21 Komlos JB, Mishra B, Lanzirotti A, Myneni SCB, Jaffe PR. 2008. Real-time speciation of uranium during active bioremediation and U(IV) reoxidation. J. Environ. Eng. 134:78–86 Koningsberger DC, Prins R. 1988. X-Ray Absorption: Principles, Applications, Techniques of EXAFS, SEXAFS, and XANES. New York: Wiley. 684 pp.

by University of Notre Dame on 08/23/14. For personal use only. Labrenz M, Druschel GK, Thomsen-Ebert T, Gilbert B, Welch SA, et al. 2000. Formation of sphalerite (ZnS) deposits in natural bioﬁlms of sulfate-reducing bacteria. Science 290:1744–47 Lawrence JR, Swerhone GDW, Leppard GG, Araki T, Zhang X, et al. 2003. Scanning transmission X-ray, laser scanning, and transmission electron microscopy mapping of the exopolymeric matrix of microbial

Annu. Rev. Earth Planet. Sci. 2009.37:367-391. Downloaded from www.annualreviews.org biofilms. Appl. Environ. Microbiol. 69:5543–54 Lee YJ, Prange A, Lichtenberg H, Rohde M, Dashti M, et al. 2007. In situ analysis of sulfur species in sulfur globules produced from thiosulfate by Thermoanaerobacter sulfurigignens and Thermoanaerobacterium thermosulfurigenes. J. Bacteriol. 189:7525–29 Lemelle L, Labrot P, Salome´ M, Simionovici A, Viso M, Westall F. 2008. In situ imaging of organic sulfur in 700–800 My-old Neoproterozoic microfossils using X-ray spectromicroscopy at the SK-edge. Org. Geochem. 39:188–202 Lemelle L, Salome´ M, Fialin M, Simionovici A, Gillet P. 2004. In situ identification and X-ray imaging of microorganisms distribution on the Tatahouine meteorite. Spectrochim. Acta B 59:1703–10 Lemelle L, Simionovici A, Susini J, Oger P, Chukalina M, et al. 2003. X-ray imaging techniques and exobiology. J. Phys IV 104:377–80 Liermann LJ, Barnes AS, Kalinowski BE, Zhou XY, Susan L. 2000. Microenvironments of pH in biofilms grown on dissolving silicate surfaces. Chem. Geol. 171:1–16

388 Templeton · Knowles Environ. Sci. Technol. 2009, 43, 7430–7436

and is a prerequisite for the effective management of aquatic Cadmium(II) Speciation in Complex and soil resources that are impacted by trace metals. Aquatic Systems: A Study with Spectroscopic studies have shown that metal adsorption onto iron oxides occurs through inner-sphere coordination Ferrihydrite, Bacteria, and an (3, 4) while ligands are generally considered to form a mix of both inner and outer-sphere complexes (5, 6). In ternary Organic Ligand metal-ligand-iron oxide systems, the ligand can either inhibit metal adsorption (by forming stable solution com- YANTAO SONG,†,⊥ plexes) or promote metal adsorption (by forming surface PETER J. SWEDLUND,*,‡ ternary complexes) depending on the conditions (7). Ternary NARESH SINGHAL,†,⊥ AND complexes involving an inner-sphere adsorbed metal with SIMON SWIFT§,⊥ the ligand bound to the metal by an inner- or outer-sphere 2+ Departments of Civil and Environmental Engineering, connection have been identified for Cd -phthalate- 2+- 2-- Chemistry, and Molecular Medicine and Pathology, University goethite and Pb SO4 goethite (8, 9). of Auckland, Private Bag 92019, Auckland, New Zealand Bacteria can present a significant reactive surface for metal ion adsorption due to their large surface area-to-volume ratio Received May 14, 2009. Revised manuscript received July and various surface functional groups. Spectroscopic and 27, 2009. Accepted July 29, 2009. modeling studies have shown that a discrete number of different functional groups are responsible for Cd2+ sorption onto many different bacterial strains (2, 10-12). In contrast, the interaction of ligands with bacterial surfaces is considered Understanding the chemical interactions that occur in to involve a hydrophobic attraction between a fully proto- complex natural systems is fundamental to their management. nated ligand and the cell membrane (13, 14). Metal adsorption In this work, the distribution of cadmium in the presence of onto bacteria-iron oxide mixtures depends on the relative proportion of bacteria in the composites but is typically less phthalic acid (H2Lp), ferrihydrite, and bacteria cells (Comamonas spp., heat killed) was measured and modeled for systems than expected from adding adsorption onto the pure phases because of surface sites being masked by the interaction with incrementally increasing complexity. In binary systems, between the bacteria and iron oxides (15-19). cadmium adsorption onto bacteria or ferrihydrite was accurately Surface complexation modeling has been widely and predicted using the nonelectrostatic four site model (NFSM) successfully used to describe sorption in increasingly complex and the diffuse layer model (DLM), respectively. Phthalic acid systems. In this work, the diffuse layer model (DLM) was 2+ (0.6 mM) enhanced Cd adsorption onto ferrihydrite (due used to describe sorption onto ferrihydrite, for reasons 2+ to surface ternary complex formation) but inhibited Cd adsorption previously described (7), while a nonelectrostatic four site onto bacteria to the same extent as predicted by Cd-phthalate model (NFSM) was used to describe sorption onto bacterial solution complex formation constants, implying no significant surfaces (12, 20, 21). The application of surface complexation surface ternary interaction occurred in this system. In modeling to describe metal speciation in systems containing Cd-ferrihydrite-bacteria systems, Cd2+ adsorption was up to both an iron oxide and bacteria is only in its infancy (22), 10% lower than that predicted by additive adsorption onto and the influence of an organic ligand on metal speciation the pure phases which suggests that an interaction between in such systems has not been previously reported. ferrihydrite and the bacteria is occupying or masking adsorption The objective of this study was to elucidate metal sites. By adding a generic reaction to the model for the speciation and distribution in a complex biogeochemical system by incrementally increasing the complexity from interaction between ferrihydrite and the bacteria, the adsorption simple binary systems up to quaternary systems containing 2+ - of Cd onto Comamonas spp. ferrihydrite was accurately a trace metal, organic ligand, iron oxide, and bacteria. 2+ predicted and Cd distribution and speciation in systems Comparing experimental and model results, as complexity containing ferrihydrite, Comamonas spp., and H2Lp could be is incrementally increased, provides qualitative and quan- predicted. titative insights into interactions between the system’s components that affect trace metal distribution. The study Introduction system used Cd2+, phthalic acid, ferrihydrite, and the Gram- Trace metal speciation in aquatic environments is inherently negative bacteria Comamonas spp. complex due to the large number of possible interactions with dissolved and particulate components. Trace metal Materials and Methods speciation is influenced by the formation of solution Materials. All solutions, except the bacterial growth broths, complexes with inorganic and organic ligands and by the were made from analytical grade reagents and 18 MΩ cm formation of complexes on the surfaces of mineral particles water that had been degassed with N2 to remove CO2. (such as iron oxides) and bacteria (1, 2). Understanding the Ferrihydrite (two-line) was synthesized from ferric nitrate interactions that occur in complex natural systems is and characterized as previously described (7). The bacterial important in predicting the fate and transport of trace metals strain Comamonas spp. was isolated from an aerobic activated sludge sample collected from the Rosedale waste- * Corresponding author tel: +64-9-923-4296; fax: +64-9-3737-422; water treatment plant in Auckland, New Zealand. It was e-mail: [email protected]. † identified using 16S rDNA sequencing as a member of Department of Civil and Environmental Engineering. Comamonas spp., with 99.8% identity to Comamonas strain ‡ Department of Chemistry. § Department of Molecular Medicine and Pathology. DJ12 (accession number AY600616.1). Comamonas spp. was ⊥ E-mail: [email protected] (Y.S.); [email protected] cultured in trypticase soy broth supplemented with 0.5% (N.S.); [email protected] (S.S.). (V/V) yeast extract.

7430 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 19, 2009 10.1021/es901434v CCC: $40.75  2009 American Chemical Society Published on Web 08/21/2009 A bacterial cell stock suspension was prepared by washing parameters (Table 1) optimized from the data. The proton cells five times in 0.01 M NaNO3 as described by Borrok et adsorption constants determined from this work are reason- al. (12). To avoid H2Lp metabolism, Comamonas spp. were ably close to previous reported values obtained using the killed at 70 °C for 15 min. Death was confirmed using the same model. The value of 3.16 for log KRHA from this work LIVE/DEAD BacLight Bacterial Viability Kit (Molecular is within the reported average values while log KRBH, log KRCH, Probes). The bacteria stock suspension in 0.01 M NaNO3 was and log KRDH are approximately half a log unit higher than prepared by vortex mixing for ∼1 min until a homogeneous the reported ranges. Bacterial surface site densities are suspension was observed. Unless stated, the bacterial cell typically between 2.0 × 10-5 and 10.0 × 10-5 mol g-1, concentration is expressed on a wet weight basis and the indicating the variety in functional group densities on various ratio of wet/dry weight was 5.4. Ferrihydrite-Comamonas bacterial surfaces. All Comamonas spp. site densities were spp. mixed suspensions were prepared by adding washed within the reported range of values ( one standard deviation. bacterial cell stock suspension to a ferrihydrite suspension The weighted average Comamonas spp. values from this work with thorough stirring. were used to model Cd2+ speciation in all systems with Adsorption Experiments. Acid-base titrations of bac- bacteria. terial cell suspensions used the method of Borrok et al. (12). Cadmium adsorption onto Comamonas spp. follows a Titrations were conducted under N2(g) using an autotitrator gradual sigmoid curve over 4 to 5 pH units (Figure 1b) shifting -1 with a drift value of 2 mV min . The electrode was calibrated to higher pH as the Cd(T)/Comamonas spp. ratio increases. by titrating the electrolyte. The pH was lowered to 2.5 prior This is typical of Cd2+ biosorption (2, 11, 20, 26) and contrasts to titrating the suspension with 0.114 M NaOH. To measure with Cd2+ adsorption onto ferrihydrite that increases from adsorption, a batch Comamonas spp. or ferrihydrite- zero to 100% over ≈2 pH units (1) which indicates a higher Comamonas spp. suspension was distributed to centrifuge degree of site heterogeneity at the bacteria surface. The NFSM 2+ 2+ tubes, H2Lp and/or Cd (as Cd(NO3)2 solution) was added, was used to model Cd adsorption data using the species + the pH was adjusted to yield a pH range, and samples were given in Table 1. The weighted average log KRACd of 3.11 is equilibrated on an end-over-end mixer for 72 h. This was in close agreement with the cited values, while the values for + + + sufficient for equilibration in all systems. The reported pH log KRBCd , log KRCCd , and log KRDCd are higher than the was then measured; samples were filtered with a 0.2 µm reported values. It is known that Cd2+ biosorption behavior + cellulose acetate membrane filter, and the dissolved Cd2 can vary between different bacterial species. For example, 2+ -1 and H2Lp were analyzed as previously described (7). The a comparison of 89 µMCd adsorption onto 10 g L Bacillus change in pH over the equilibration time was, in all cases, subtilis, Bacillus licheniformis (26), and Comamonas spp. at less than 1 pH unit. Identical experiments with no sorbents I ) 0.01 M showed that the bacteria exhibited affinity for + showed no Cd2 loss at the pH values of this work (pH < 8). Cd2+ in the following order Comamonas spp.> Bacillus subtilis Modeling Adsorption. The DLM was used to model >Bacillus licheniformis. In contrast, the parameters derived adsorption onto ferrihydrite using the parameters of Dzom- from this work for Comamonas spp. can accurately model - bak and Morel (1) including the surface area (600 m2 g 1) and Cd2+ adsorption onto Pseudomonas putida and Pseudomonas the site densities for high and low cation affinity sites, mendocina (20), which are Gram-negative like Comamonas - respectively, [≡FesOH] ) 0.005 mol (mol Fe) 1 and [≡FewOH] spp. It has been observed that there is a wide range of bacteria ) -1 ≡ x 0.2 mol (mol Fe) . Surface pKA’s of both site types ( Fe OH) that exhibit similar proton and metal adsorption behavior, were 7.29 and 8.93, and log K’s for the formation of the species and the metal adsorption onto these bacterial cells can be ≡ s + ≡ w + ≡ x 0 ≡ x - Fe OCd , Fe OCd , Fe HLp , and Fe Lp were 0.22, reasonably well modeled using only one set of averaged -2.90, 15.85, and 10.23, respectively, while the log K’s for the parameters for site densities and adsorption constants (2, 12). ≡ s formation of the ternary complexes Fe OHCdLp and However, this does not deny the existence of bacteria with ≡ w Fe OHCdLp were 9.86 and 7.56 (1, 7). The NFSM (12) was significantly different adsorption behavior (21). used to model sorption onto bacterial cells using the Most reported Cd2+ biosorption has used a limited range - - - - -1 components RA ,RB,RC, and RD to represent four of metal(T)/bacteria ratios from 5 to 27 µmol g (2, 10, 14, Comamonas spp. surface sites. All bacterial surface species 26-28). In our work, Cd2+ adsorption onto Comamonas spp. had activity coefficients set to 1. The point of zero charge was determined with Cd(T)/bacteria ratios ranging from 0.65 (PZC) of a bacterial surface is typically at pH < 2 where cell µ -1 + 2+ + to 32.4 mol g . Using the H and Cd adsorption constants damage occurs, so the total H concentration is not known derived from the experimental data in this work, cadmium and is an optimized parameter (23). Sorption constants were adsorption onto Comamonas spp. was reasonably well optimized in FITEQL as previously described (7). The initial reproduced over a wide range of Cd(T)/bacteria ratios and estimates for bacteria proton sorption parameters were from pH values (Figure 1b). While there have been some spec- Borrok et al. (12). The precise values did not influence the troscopic studies of Cd2+ sorption by bacteria, the distribution result, but log K’s needed to span a similar range of values of Cd2+ between the different model sites should be viewed for convergence. All modeling was performed at a fixed ionic as an average description of the surface rather than providing strength, and all equilibrium constants were reported at an specific mechanistic information. In addition, no attempt ionic strength of zero. Data sets and replicates were fitted was made to model bacteria sorption data with less site types separately, and weighted averages and 95% confidence because the objective of the study was to apply the existing intervals were calculated (1). All model curves presented generally accepted model to a complex system, rather than use weighted average values. Solution equilibrium constants to simplify the model. were from Gustaffson (24), and solution activity coefficients Experiments with Comamonas spp. showed that H2Lp were calculated with the Davies equation. adsorption was only observed at pH less than 3 with ≈7% H2Lp absorbed at pH 2. It has been proposed that organic Results and Discussion acid adsorption onto bacterial cells is determined by the Adsorption onto Comamonas spp. The titration data for balance between electrostatic and hydrophobic interactions. Comamonas spp. are given in Figure 1a with the modeled When an organic acid is fully protonated and, therefore, results calculated using previously reported parameters and uncharged (i.e., pH , pKA1), the hydrophobic interaction the values from this study. The parameters from the “universal between the organic acid and bacterial surface dominates model” (25) overestimated the titration data from this work the adsorption. With increasing pH, both the organic acid at pH < 7 and underestimated the data at pH > 7. The and the bacterial surface become more negatively charged, Comamonas spp. data were accurately modeled using the and absorption is observed to decrease due to the increasing

VOL. 43, NO. 19, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 7431 surface sites through the ferrihydrite-bacteria interaction which would diminish Cd2+ adsorption. On the other hand, ferrihydrite tends to increase Cd2+ sorption directly through Cd-ferrihydrite surface complexes. The former effect is more significant at lower pH but less significant at higher pH. The latter effect is more significant at higher pH. The ability of the model parameters to predict Cd2+ distribution in the quaternary system demonstrates that the model can provide insight into the processes that determine soluble and particulate metal species in complex biogeochemical environments. Iron oxides, bacteria, organic ligands, and metal ions are ubiquitous in natural environments and many engineered water treatment systems. Either bacteria or iron oxide can dominate metal adsorption (depending on the conditions), and the effect of organic ligands depends not only on the ligand but also on which phase dominates metal adsorption. The management of water treatment systems, especially those using iron oxides and bacteria- bearing activated sludge, can benefit from elucidating the interactions occurring between the components. Acknowledgments Y.S. acknowledges the New Zealand International Doctorial Research Scholarship and University of Auckland Interna- tional Doctoral Scholarship. Literature Cited (1) Dzombak, D. A.; Morel, F. M. M. Surface complexation modeling: hydrous ferric oxide; John Wiley & Sons: New York, 1990. (2) Yee, N.; Fein, J. Cd adsorption onto bacterial surfaces: A universal adsorption edge. Geochim. Cosmochim. Acta 2001, 65, 2037– 2042. FIGURE 4. (a) Experimental (symbols) and modeled (lines) (3) Peacock, C. L.; Sherman, D. M. Copper(II) sorption onto goethite, results for Cd2+ sorption in various systems. (b) Cd2+ speciation hematite and lepidocrocite: A surface complexation model based on ab initio molecular geometries and EXAFS spectroscopy. in quaternary system QS2. I ) 0.01 M NaNO3, ferrihydrite -1 Geochim. Cosmochim. Acta 2004, 68, 2623–2637. ) 0.1gL ,H2Lp(T) ) 0.6 mM, Cd(T) ) 1.78 µM. Bacteria wet -1 (4) Spadini, L.; Manceau, A.; Schindler, P. W.; Charlet, L. Structure and and dry weights are 0.275 and 0.05 g L , respectively. stability of Cd2+ surface complexes on ferric oxides: 1. Results from The significance of the bacterial surface absorbing EXAFS spectroscopy. J. Colloid Interface Sci. 1994, 168, 73–86. + (5) Peak, D.; Ford, R. G.; Sparks, D. L. An in situ ATR-FTIR Cd2 can be more clearly observed by decreasing the - investigation of sulfate bonding mechanisms on goethite. J. ferrihydrite Comamonas spp. (dry weight) ratio from 10 to Colloid Interface Sci. 1999, 218, 289–299. 2 (Figure 4). In this case, adsorption in the binary system (6) Catalano, J. G.; Park, C.; Fenter, P.; Zhang, Z. Simultaneous with Comamonas spp. was substantially greater than with inner- and outer-sphere arsenate adsorption on corundum and ferrihydrite at pH < 7. This indicates that 0.05 g L-1 (dry hematite. Geochim. Cosmochim. Acta 2008, 72, 1986–2004. weight) Comamonas spp. can accumulate more Cd2+ than (7) Song, Y.; Swedlund, P. J.; Singhal, N. Copper(II) and Cadmium(II) -1 < Sorption onto Ferrihydrite in the Presence of Phthalic Acid: 0.1gL ferrihydrite, particularly at pH 6.5. Adding H2Lp Some Properties of the Ternary Complex. Environ. Sci. Technol. 2+ inhibited Cd adsorption by Comamonas spp. but enhanced 2008, 42, 4008–4013. + + Cd2 adsorption onto ferrihydrite so that Cd2 adsorption in (8) Boily, J.-F.; Sjo¨berg, S.; Persson, P. Structures and stabilities of the presence of H2Lp was greater on ferrihydrite than Cd(II) and Cd(II)-phthalate complexes at the goethite-water Comamonas spp. at pH > 6. In all systems with ferrihydrite, interface. Geochim. Cosmochim. Acta 2005, 69, 3219–3235. the addition of 0.275 g L-1 Comamonas spp. significantly (9) Elzinga, E. J.; Peak, D.; Sparks, D. L. Spectroscopic studies of 2+ Pb(II)-sulfate interactions at the goethite-water interface. enhanced Cd sorption over the investigated pH range Geochim. Cosmochim. Acta 2001, 65, 2219–2230. (Figure 4a), indicating that the bacterial cells are important (10) Boyanov, M. I.; Kelly, S. D.; Kemner, K. M.; Bunker, B. A.; Fein, absorbent particularly at a lower pH range. The effect of J. B.; Fowle, D. A. Adsorption of cadmium to Bacillus subtilis adding H2Lp to the Comamonas spp.-ferrihydrite mixtures bacterial cell walls: a pH-dependent X-ray absorption fine was much lower with 0.275 g L-1 than 0.055 g L-1 Comamonas structure spectroscopy study. Geochim. Cosmochim. Acta 2003, 67, 3299–3311. sp. This is because the ferrihydrite surface, on which the (11) Kelly, S. D.; Boyanov, M. I.; Bunker, B. A.; Fein, J. B.; Fowle, ferrihydrite-Cd-Lp complexes form, is less important with - D. A.; Yeeb, N.; Kemnera, K. M. XAFS determination of the 0.275 g L 1 Comamonas sp. The net effect of these complex bacterial cell wall functional groups responsible for complex- interactions is that Cd2+ adsorption in the quaternary systems ation of Cd and U as a function of pH. J. Synchrotron Radiat. with 0.275 g L-1 Comamonas spp. is only slightly greater 2001, 8, 946–948. than that with 0.055 g L-1 Comamonas spp. This is in contrast (12) Borrok, D.; Fein, J. B.; Kulpa, C. F. Proton and Cd adsorption onto natural bacterial consortia: Testing universal adsorption to all systems that did not have both H2Lp and ferrihydrite behavior. Geochim. Cosmochim. Acta 2004, 68, 3231–3238. 2+ in which Cd adsorption was substantially greater with 0.275 (13) Daughney, C. J.; Fein, J. B. Sorption of 2,4,6-Trichlorophenol by - - gL 1 rather than with 0.055 g L 1 Comamonas sp. Bacillus subtilis. Environ. Sci. Technol. 1998, 32, 749–752. It is worth considering the effect of ferrihydrite on Cd2+ (14) Fein, J. B.; Delea, D. Experimental study of the effect of EDTA biosorption. Compared to Cd-bacteria binary systems, the on Cd adsorption by Bacillus subtilis: a test of the chemical presence of ferrihydrite decreased Cd2+ adsorption at pH equilibrium approach. Chem. Geol. 1999, 161, 375–383. ∼ ∼ (15) Ferris, F. G.; Hallberg, R. O.; Lyveń, B.; Pedersen, K. Retention less than 6 but increased adsorption at pH over 6. The of strontium, cesium, lead and uranium by bacterial iron oxides 2+ effects of ferrihydrite on Cd sorption were due to two from a subterranean environment. Appl. Geochem. 2000, 15, processes. On the one hand, ferrihydrite can mask bacterial 1035–1042.

VOL. 43, NO. 19, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 7435

Trace Metal Speciation in Complex Aquatic Environments

The Cu2+, Cd2+, Ferrihydrite, Phthalic Acid and Bacteria System

Yantao Song

A thesis submitted in fulfillment of the requirements for the degree of Doctor of Philosophy, The University of Auckland, 2009.

Chapter 1. – Introduction

Boyanov et al. 2003; Fowle and Fein 2000; Mullen et al. 1989). Since 1970s bacterial adsorption of metal ions has received increasing attention. The biosorption of metal ions onto bacterial cells mainly includes ionic interactions and complex formation between metal ions and organic acid functional groups at the cell surface (Boyanov et al. 2003; Gabr et al. 2008). Spectroscopic techniques such as IR and XAFS have demonstrated that there are numerous functional groups including carboxyl, phosphoryl, amino and phenolic groups (Boyanov et al. 2003; Kelly et al. 2001). For metal biosorption, numerous EXAFS spectroscopy studies have indicated that carboxyl and phosphoryl functional groups account for metal adsorption onto bacterial surface, no matter if the bacteria are Gram-positive (Boyanov et al. 2003; Kelly et al. 2001) or Gram-negative (Hennig et al. 2001; Toner et al. 2005). More recently EXAFS spectroscopy examined Cd2+ adsorption onto bacterial consortia containing both Gram- positive and Gram-negative bacteria growing from natural environment (river water) and engineered systems (manufacturing gas plant). It has been proposed that only a limited number of bacterial surface functional groups are responsible for Cd2+ biosorption despite of the complex mixture of the bacteria. Three types of bacterial surface functional groups (carboxyl, phosphoryl and sulfhydryl), or even only two types of bacterial surface functional groups (carboxyl and phosphoryl) were sufficient to model Cd2+ adsorption onto bacterial consortia from river water or a gas plant (Mishra et al. 2009).

Adsorption of a wide variety of metal ions onto both Gram-positive and Gram-negative bacteria has been qualitatively and quantitatively evaluated (Beveridge 1978; Beveridge and Koval 1981; Borrok and Fein 2005; Burnett et al. 2007; Daughney et al. 2000; Mullen et al. 1989). One important factor for metal adsorption onto bacterial surface is pH. The pH at which the total number of positive charges equals the number of negative charges is referred to as the point of zero charge (PZC). The PZC of bacterial cells was reported to be in the range of pH 1.75 ~ 4.15 (Harden and Harris 1953), and therefore bacterial cells were almost always negatively charged. In general metal adsorption onto bacterial cells increases with pH increasing from around 2 to 8 (Burnett et al. 2007; Ginn and Fein 2008; Johnson et al. 2007). However it has been observed that metal adsorption onto bacterial cells can also decrease as the pH increases above pH ≈ 8 (Burnett et al. 2007). Compared to metal adsorption onto ferrihydrite which increases from zero to 100 % over approximately 2 pH units (Dzombak and Morel 1990), metal sorption onto bacteria generally showed a more gradually sigmoid curve over about 3 ~ 5 pH units. This implies a higher degree of site heterogeneity at the surface of the bacteria than that on the ferrihydrite surface. The average pKA values for

Chapter 6. Cadmium and Phthalic Acid Sorption onto Comamonas spp.

6.3. Cadmium sorption onto Comamonas spp.

Cadmium adsorption onto Comamonas spp. in binary systems was determined over a pH -1 range from 2 to 8 for Cd(T)/Comamonas spp. ratios of 0.65, 1.08, 3.24 and 32.4 μmolg . The experimental data for Cd2+ sorption as a function of pH are shown in Figure 6.2. The pH dependence of Cd2+ biosorption has been observed for other bacteria (Borrok and Fein 2005; Burnett et al. 2006; Daughney and Fein 1998a; Kelly et al. 2001; Yee and Fein 2001). In general Cd2+ adsorption increases with an increase in pH up to 8, however Cd2+ adsorption sorption was found to decrease at pH > 8 (Burnett et al. 2006). This phenomenon was not observed for Comamonas spp. in this work because the pH investigated was less than 8.

Similar to Cd2+ adsorption onto ferrihydrite, Cd2+ sorption onto Comamonas spp. depends on pH and the Cd(T)/bacteria ratio. Cadmium sorption onto Comamonas spp. shows a more gradually sigmoid curve over about 5 pH units compared to Cd2+ adsorption onto ferrihydrite which increases from zero to 100 % over approximately 2 pH units (Figure 3.2). This implies a higher degree of site heterogeneity at the surface of the bacteria than that on the ferrihydrite surface. Bacteria’s having flatter titration curves than that of ferrihydrite (Figure 6.1) is evidence showing the presence of surface sites with a greater range of pKA values. The high degree of site heterogeneity on bacteria has been demonstrated by titration experiments and electrostatic force microscopy that showed pKA values for proton adsorption onto Shewanella putrefaciens ranged from 3.72 to 9.99 (Sokolov et al. 2001).

2+ The Cd adsorption edges in Figure 6.2 clearly shift to higher pH as the Cd(T)/Comamonas -1 2+ spp. ratio increases from 0.65 to 32.4 μmolg . The pH50 for Cd sorption onto Comamonas spp. occurs at pH ≈ 4.5, 4.8, 5.5 and 7.0, for the data with Cd(T)/Comamonas spp. ratio of 0.65, 1.08, 3.24, and 32.4 μmolg-1, respectively. A similar pH edge shift was observed for 2+ Cd adsorption onto ferrihydrite as a function of Cd(T)/Fe ratios. For example, about one pH -5 -3 -1 unit shift to the higher pH when the Cd(T)/Fe increased from 7.8×10 to 7.9×10 mol mol (Figure 3.2). The pH edges of Cd2+ adsorption onto Comamonas spp. suggested that for the same dry weight of Comamonas spp. and ferrihydrite, the bacteria absorb more Cd2+ at a 2+ given pH, i.e. Cd biosorption has a lower pH50 value. This conclusion is confirmed by -1 comparing 1.78 μM Cd(T) adsorption onto 0.1 gL ferrihydrite and Comamonas spp. (Figure 6.3). It shows that Comamonas spp. accumulates more Cd2+ than ferrihydrite over a pH range

References

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form similar to methylmercury acetate. These results are Using X-ray Microscopy and Hg L3 interpreted within the context of obtaining a “snapshot” of XANES To Study Hg Binding in the mercury methylation in progress.

Rhizosphere of Spartina Cordgrass Introduction

CYNTHIA PATTY,† BRANDY BARNETT,‡ Many challenges face the health of the San Francisco Bay BRIDGET MOONEY,‡ AMANDA KAHN,§ ecosystem today; among them are high levels of mercury | SILVIO LEVY,‡ YIJIN LIU, and invasion by exotic species. Mercury is a highly toxic heavy PIERO PIANETTA,† AND metal and neurotoxin that may bioaccumulate and bio- ,†,‡ JOY C. ANDREWS* magnify. In florae, mercury can greatly harm or even kill a Stanford Synchrotron Radiation Lightsource (SSRL), plant by interrupting photosynthetic processes (1). Wetlands 2575 Sand Hill Road, SLAC MS 69, Menlo Park, have commonly been regarded as sinks for mercury and other California 94025, Department of Chemistry and Biochemistry, pollutants. Recently, however, wetlands are emerging as California State University East Bay, 25800 Carlos Bee significant sources of mercury export into estuarine food Boulevard, Hayward, California 94542, Moss Landing Marine webs (2, 3). Laboratories, 8272 Moss Landing Road, Moss Landing, Little is known about the cycling of mercury and other California 95039, and Institute for High Energy Physics, Beijing, China heavy metals by plants or how changes in plant community structure might affect mercury bioavailability. Mercury enters the food chain when it is taken up by organisms at the base Received April 8, 2009. Revised manuscript received of the food web, including microbes and plants. Plants, and August 17, 2009. Accepted August 21, 2009. microbes in their rhizosphere, mobilize nutrients in the sediment environment, shifting chemical speciation of compounds (4, 5). Sulfate-reducing bacteria (SRB) (6, 7) and iron reducing bacteria (FeRB) (8, 9) have been shown to have San Francisco Bay has been contaminated historically by significant impact on Hg speciation, particularly methylation, mercury from mine tailings as well as contemporary industrial in the rhizosphere. sources. Native Spartina foliosa and non-native S. Plant communities inherently shape sediment composi- alterniflora-hybrid cordgrasses are dominant florae within tion and structure through their functions. Accordingly, the the SF Bay estuary environment. Understanding mercury uptake types of plants in a wetland and their abilities to take up and transformations in these plants will help to characterize metals affect mercury bioavailability (10). It follows then that the significance of their roles in mercury biogeochemical invasive plant species might alter the natural biogeochemical cycling in the estuarine environment. Methylated mercury can cycling of mercury in the Bay. Notably, through unsuccessful be biomagnified up the food web, resulting in levels in competition, native plants in the SF Bay are quickly being sport fish up to 1 million times greater than in surrounding replaced by exotic species. Spartina alterniflora-hybrids, waters and resulting in advisories to limit fish intake. including both S. alterniflora and the S. alterniflora x S. foliosa Understanding the uptake and methylation of mercury in the cross, are quickly out-competing and out-breeding the native S. foliosa, California cordgrass (11). S. alterniflora has shown plant rhizosphere can yield insight into ways to manage the ability to sequester metals in its above- and below-ground mercury contamination. The transmission X-ray microscope tissues to varying degrees (12), but the definitive relationship on beamline 6-2 at the Stanford Synchrotron Radiation between invasive S. alterniflora, or smooth cordgrass, and Lightsource (SSRL) was used to obtain absorption contrast changes in metal cycling is still unknown. images and 3D tomography of Spartina foliosa roots that were Despite the close relationship between S. alterniflora and exposed to 1 ppm Hg (as HgCl2) hydroponically for 1 week. S. foliosa, little research has been performed on the uptake, Absorption contrast images of micrometer-sized roots from S. storage, and translocation of heavy metals by S. foliosa or foliosa revealed dark particles, and dark channels within the hybrid. The purpose of this experiment is to compare the root, due to Hg absorption. 3D tomography showed that mercury uptake of native S. foliosa and non-native S. the particles are on the root surface, and slices from the alterniflora-hybrid and to study the speciation of Hg in the tomographic reconstruction revealed that the particles are rhizosphere. Understanding mercury uptake in these plants hollow, consistent with microorganisms with a thin layer of Hg and speciation in the chemically active rhizosphere will help to characterize their roles in mercury biogeochemical cycling on the surface. Hg L XANES of ground-up plant roots and 3 in the estuarine environment. Hg L micro-XANES from microprobe analysis of micrometer- 3 Mercury speciation can be determined using a variety of sized roots (60-120 µm in size) revealed three main types - techniques, including X-ray Absorption Spectroscopy (XAS). of speciation in both Spartina species: Hg S ligation in a form XAS has been used to determine mercury speciation in similar to Hg(II) cysteine, Hg-S bonding as in cinnabar minerals (13-15) and in plants (16, 17). Advantages of XAS and metacinnabar, and methylmercury-carboxyl bonding in a over the other techniques include the following: (1) it is element-specific, thus mercury can be analyzed despite other, more dominant species; (2) no chemical or physical alteration is required before sample analysis; and (3) sample quantities * Corresponding author phone: (650)926-4285; fax: (650)926-4100; can be small (less than 1 g). X-ray absorption near edge e-mail: [email protected]. structure (XANES) can be very useful to determine speciation † Stanford Synchrotron Radiation Lightsource. ‡ California State University East Bay. by comparison of spectra with appropriate model com- § Moss Landing Marine Laboratories. pounds. XANES has been used to determine the chemical | Institute for High Energy Physics. form of mercury in fish (18) and the binding of methylmercury

10.1021/es901076q CCC: $40.75  2009 American Chemical Society VOL. 43, NO. 19, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 7397 Published on Web 09/04/2009 showed signs of distress. While most inner leaves and/or stems were still green and vibrant, outer leaves were yellowing and withering, thus the treatment was ended after 1 week. Hg Speciation and Transformation by Microorganisms. The collection of mercury chemical species found within Spartina root hot spots may provide information about the chemical and biochemical transformations of Hg in the rhizosphere. Least-squares fitting of Hg L3 XANES of bulk roots of S. foliosa and S. alterniflora revealed that Hg speciation was similar in both cordgrasses. The micro- and bulk XANES fitting both indicated that most Hg chelation is in a form similar to Hg-S binding in Hg-cysteine and FIGURE 3. Microprobe map of Hg fluorescence on logarithmic dicysteine. This speciation is consistent with binding to thiol scale shows size and distribution of Hg within micro- groups of proteins, possibly on the cell walls of microbes (6) meter-sized S. foliosa roots (A). Micro-XANES points were or within root channels. Hg was clearly seen on the surface selected from the highest concentration areas (“hot spots”, of the microbes in the TXM images (Figure 1D,E), thus our indicated white). Fluorescent counts (range 0-603) in insets (B results are consistent with observations by Mason et al. (29) and C) were determined as a difference above the Hg edge that much of HgCl2 made available to microorganisms (12300 eV), minus below (12250 eV). (diatoms) ends up associated with the cell walls. There is some thought that binding to thiol groups may increase the determined to be the absorbing element in the dark spots availability of Hg for methylation (6). of the TXM images. In this study, all methylated mercury was bound to Spatial maps consistently showed the highest mercury carboxyl groups, possibly to organic acids. In studies of Cd (“hot spots”), or densely concentrated areas, in the tips of binding to Bacillus subtilis, Cd was bound to carboxyl groups the fine root structures (Figure 3B), but Hg was also found in the cell walls (30). However, Mason et al. (29) found that concentrated in channels along the center of the root (Figure in diatoms, most of the methylmercury they took up was 3A). Distribution and particle size of the Hg hot spots varied. found in the cytoplasm. The present work does not indicate For example, some mercury particles appeared in aggregates the exact methylmercury location. TXM images show that upto2or3µm in size, while others appeared smaller in size most of the Hg associated with microorganisms is exterior and more diffusely distributed. to the plant, within microbial cell walls. However, meth- Micro-XANES. Hg L3 micro-XANES spectra collected from the areas of highest Hg concentrations were fit to Hg model ylmercury could be in the cytoplasm of SRB or within plant compounds using the same least-squares method as for the roots. Because methylmercury is found in the same hot spots bulk samples. These areas represented more localized as cysteine-bound mercury (discussed below), the two species environments in the rhizosphere than the bulk root tissue are in close proximity. samples. Localized chemical species (Table 2) were propor- Hg as cinnabar, seen in some of the bulk samples, was tioned similarly to the bulk results, except for cinnabar which entirely absent from the micro-XANES. This inconsistency was absent. Hg-S binding as metacinnabar comprised may suggest that the cinnabar had different origins than the 0%-43.5% of the chemical species, Hg-S binding in a form species found in the micro-XANES. For example cinnabar similar to Hg(II)cysteine represented 20.5%-96.6%, and found in the bulk samples could have been due to residual Hg-O coordination (a sum of methyl mercury acetate and sediment from the sampling site, which was caught in root mercury diacetate) represented 0%-36.0%. tissue and ground with the sample. The bulk root samples contained thicker root portions as well as micrometer-sized Discussion roots, whereas TXM and microprobe (and micro-XANES) Hg Uptake by S. foliosa and S. alterniflora. After growing were performed only on micrometer-sized roots, where in the presence of mercury for 1 week, Spartina roots showed sediment was less likely to adhere. However, Hg as meta- significant accumulation of mercury, and S. foliosa roots took cinnabar was seen in clustered hot spots of microprobe Hg up significantly greater concentrations of mercury than S. maps via micro-XANES. Clusters of mercury nanoparticles alterniflora. Control Hg levels showed no significant differ- were found next to a metacinnabar crystal in TXM images ences between species (and therefore site). Accumulation (Figure 1A top and Figure 2). In the microprobe maps the Hg results suggest that S. foliosa is either more efficient at clusters appeared more diffuse, because the particles are removing mercury from its surroundings, or it has more less than the ∼0.5 µm resolution. established microbial populations in the rhizosphere. Whereas Overall, the proximity of various Hg species found in these our Hg shoot concentrations were low, longer-term studies roots further indicates a chemical and/or biochemical found that S. alterniflora transported more mercury to the relationship. It is noteworthy that three species of mercury, shoots (4, 10, 12). It is possible that translocation of Hg to namely Hg bound to cysteine, methymercury, and meta- shoots would have increased given a longer treatment time. cinnabar, were all found in the same regions in the micro- However, control and treated plants of both Spartina species XANES, indicating that they occur in close proximity. As

TABLE 2. Hg L3 Micro-XANES Least Squares Fitting Results of Micrometer-Sized S. foliosa Roots To Model Compounds (Percent)

mercury cysteine and sample cinnabar (HgS) meta-cinnabar (HgS) dicysteine methyl mercury acetate mercury diacetate S. foliosa root 1 region 1 a 18.6 ( 2.5 57.9 ( 3.9 9.6 ( 3.9 13.9 ( 1.6 S. foliosa root 1 region 2 s 29.1 ( 2.0 47.0 ( 2.9 7.9 ( 3.0 16.0 ( 1.2 S. foliosa root 2 region 1 ss96.6 ( 4.1 3.4 ( 4.1 s S. foliosa root 2 region 2 s 22.8 ( 2.6 64.5 ( 2.5 s 12.0 ( 1.5 S. foliosa root 3 region 1 s 43.5 ( 1.3 20.5 ( 2.0 36.0 ( 1.9 s a Less than 1% of total composition.

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7402 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 19, 2009 Chemistry Central Journal

Research article Open Access Quantum mechanical calculation of aqueuous uranium complexes: carbonate, phosphate, organic and biomolecular species James D Kubicki*1,2, Gary P Halada3, Prashant Jha3 and Brian L Phillips4

Address: 1Department of Geosciences, The Pennsylvania State University, University Park, PA 16802, USA, 2The Earth & Environmental Systems Institute, The Pennsylvania State University, University Park, PA 16802, USA, 3Department of Materials Science and Engineering, Stony Brook University, Stony brook, New York 11794-2275, USA and 4Dept. of Geological Sciences, Stony Brook University, Stony brook, New York 11794- 2275, USA Email: James D Kubicki* - [email protected]; Gary P Halada - [email protected]; Prashant Jha - [email protected]; Brian L Phillips - [email protected] * Corresponding author

Published: 18 August 2009 Received: 23 September 2008 Accepted: 18 August 2009 Chemistry Central Journal 2009, 3:10 doi:10.1186/1752-153X-3-10 This article is available from: http://journal.chemistrycentral.com/content/3/1/10 © 2009 Kubicki et al

Abstract Background: Quantum mechanical calculations were performed on a variety of uranium species representing U(VI), U(V), U(IV), U-carbonates, U-phosphates, U-oxalates, U-catecholates, U- phosphodiesters, U-phosphorylated N-acetyl-glucosamine (NAG), and U-2-Keto-3-doxyoctanoate (KDO) with explicit solvation by H2O molecules. These models represent major U species in natural waters and complexes on bacterial surfaces. The model results are compared to observed EXAFS, IR, Raman and NMR spectra. Results: Agreement between experiment and theory is acceptable in most cases, and the reasons for discrepancies are discussed. Calculated Gibbs free energies are used to constrain which configurations are most likely to be stable under circumneutral pH conditions. Reduction of U(VI) to U(IV) is examined for the U-carbonate and U-catechol complexes. Conclusion: Results on the potential energy differences between U(V)- and U(IV)-carbonate complexes suggest that the cause of slower disproportionation in this system is electrostatic 5- repulsion between UO2 [CO3]3 ions that must approach one another to form U(VI) and U(IV) rather than a change in thermodynamic stability. Calculations on U-catechol species are consistent 2+ with the observation that UO2 can oxidize catechol and form quinone-like species. In addition, outer-sphere complexation is predicted to be the most stable for U-catechol interactions based on calculated energies and comparison to 13C NMR spectra. Outer-sphere complexes (i.e., ion pairs bridged by water molecules) are predicted to be comparable in Gibbs free energy to inner-sphere complexes for a model carboxylic acid. Complexation of uranyl to phosphorus-containing groups in extracellular polymeric substances is predicted to favor phosphonate groups, such as that found in phosphorylated NAG, rather than phosphodiesters, such as those in nucleic acids.

Background capable of forming a wide variety of aqueous and surface The toxicity and radioactivity of U makes it a potentially complexes. Furthermore, redox reactions, mainly between hazardous element in the environment. In areas of high U U(VI) and U(IV), are common in subsurface environ- concentrations, understanding U chemistry is imperative ments (e.g., [1]). in order to predict its fate, transport, and risk. Uranium is

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tion of atoms in the model other than those involved in the monodentate uranyl complex, a uranyl-organophos- forming the second bond between the phosphoryl group phate complex (UO2-OrgHPO4; Fig. 10c) was investi- and the uranyl account for some of this lowering in gated. A stable monodentate configuration was obtained energy, it is clear that the bidentate configuration is lower when the phosphoryl group was protonated which in Gibbs free energy by much more that the expected resulted in a U---P distance of 3.8 Å (Table 4). The errors in the calculations. (Note: The energy minimiza- observed value is between the two calculated bidentate tions beginning with three different starting configura- and monodentate values, so either the monodentate com- tions resulted in final structures with energies within ± 25 plex calculation is overestimating the U---P distance by kJ/mol of the average. This order of magnitude is used as 0.2 Å or there exists a mixture of mono- and bidentate ura- an estimate of the uncertainty in the Gibbs free energies nyl-phosphoryl complexes on bacterial surfaces that result resulting from these energy minimization procedures.) in an average value of 3.6 Å. Furthermore, in case of the protonated phosphoryl group, the calculated U-O(P) Although the calculated energetics clearly favor the biden- bond was 2.46 Å, significantly longer than the 2.30 to tate configuration, the calculated U---P distance in these 2.37 Å [117] or 2.29 Å [116] observed. These discrepan- complexes is approximately 3.3 Å (Table 4) whereas the cies may be due to the role that multiple U-O-P linkages observed value is 3.64 to 3.68 Å [111,116]. This discrep- play in U binding to polyphosphate groups on bacterial ancy could be due to the fact that the Kelly et al. [47,111] surfaces [116,119]. These complexes were not modeled in experiments were conducted at a pH of 1.7 to 4.8, so the this study and should be included in future work. phosphoryl groups were probably protonated which pre- vented bidentate bonding. The model structures of UO2- UO2-GlcNPO4 OrgPO4 monodentate (Table 4) are also consistent with a Another possible explanation of the discrepancies in shorter U-Oeq bond of 2.30 to 2.37 Å measured by Koban interatomic distances between EXAFS and these calcula- et al. [117] using EXAFS in solutions of pH 3.5 to 5.5. tions may be that uranyl binds to a phosphoryl group Francis et al. [116] performed EXAFS on samples reacted unlike the phosphodiester model. To examine this possi- at pH 5 and obtained similar results as in Kelly et al. [111], bility, models of the glucosamine phosphoryl groups, but the expected pKa for phosphoryl groups on bacterial GlcNPO4 (Fig. 11) were also studied which represents a surfaces is approximately 7.2 [118], so experimental con- phosphonate compound. Phosphonates have been ditions may not have reached a state where the phospho- shown to complex U(VI) effectively [113,120]. The UO2- ryl was predominantly deprotonated. GlcNPO4 model calculations also predicted that the bidentate configuration had a lower Gibbs free energy In order to investigate the possibility that a protonated than the monodentate configuration but only by approx- phosphoryl (i.e., mimicking low pH conditions) favors imately -30 kJ/mol as compared with a difference of -520

Table 4: Model aqueous Gibbs free energies (without ZPE corrections) and interatomic distances calculated for uranium model complexes with biological ligands. U-X stands for the shortest U to P or C distance in the model.

Models Energy CN U=O U-O(P) U-O(H2)U-X

Expt [47] ----- 8 1.77 2.33 2.45 3.64

UO2-OrgPO4•27(H2O) bi -3601.0619 7 1.79 2.42* 2.57 3.18 UO2-OrgPO4•27(H2O)a mono, initial ≈-3600.4331* 7 1.79 2.30 2.53 3.40

UO2-OrgPO4a•27(H2O) bi, final -3601.0751 8 1.77 2.57 2.61 3.24

UO2-OrgPO4b•27(H2O) mono, initial ≈-3600.8669* 7 1.79 2.30 2.53 3.43

UO2-OrgPO4b•27(H2O) bi, final -3601.0662 7 1.77 2.47 2.66 3.27

UO2-OrgHPO4•27(H2O) mono -3601.5378 7 1.77 2.46 2.48 3.83

UO2-GlcNPO4•26(H2O) mono -3652.6059 7 1.81 2.23 2.51 3.61 UO2-GlcNPO4•26(H2O) bi** -3652.6170 7 1.80 2.43 2.52 3.18 Expt [47] ----- 8 1.77 2.33 2.45 2.89

UO2-KDO•26(H2O) mono -3029.6108 6 1.76 2.39 2.34 3.44 UO2-KDO•26(H2O) bi -3029.6144 7 1.75 2.42 2.51 2.89 UO2-KDO•26(H2O) OS -3029.6099 7 1.75 ----- 2.47 5.12

* – Gibbs free energies are estimated because no stable energy minimum was determined. + ** – One U-OH2 has deprotonated to form a U-OH and a H3O in the solvation sphere.

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Page 28 of 29 (page number not for citation purposes) American Mineralogist, Volume 94, pages 1377–1387, 2009

Optimizing experimental design, overcoming challenges, and gaining valuable information from the Sb K-edge XANES region

Sk y a E. Fa w c E t t ,1,* Ro b ER t a. Go R d o n ,2 a n d HE a t HER E. Ja m i ES o n 1

1Department of Geological Sciences and Geological Engineering, Queen’s University, Kingston, Ontario K7L 3N6, Canada 2Simon Fraser University and PNC-XOR, Advanced Photon Source, Argonne, Illinois 60439, U.S.A.

ab S t R a c t There are many challenges associated with collecting, processing, and interpreting high-energy XAS data. The most significant of these are broad spectra, minimal separation of edge positions, and high background owing to the Compton tail. Studies of the Sb system are a particular challenge owing to its complex bonding character and formation of mixed oxidation-state minerals. Furthermore, in environmental samples such as stream sediment containing mine waste, different Sb phases may coexist. Ways to overcome these challenges and achieve accurate and useful information are presented. Our investigations used Sb K-edge X-ray absorption near-edge spectroscopy (XANES) to elucidate Sb geochemical behavior. Several Sb mineral spectra are presented, including Sb sulfosalts, and contrasted based on the different hosting and coordination environments around the Sb atom in the crystal structure. These comparisons lead to the recognition of how the different hosting and coordination environments are manifested in the shape of the Sb mineral spectra. In fact from the shape of the spectra, the occupation of the Sb atom in a single or in multiple crystallographic sites, regardless of whether multiple phases are present in the sample, is discernible. Furthermore, we demonstrate that quantitative information can be derived from the XANES region using linear combination fitting of the derivative spectra, rather than the energy spectra. Particularly useful to the advancement of Sb research is the demonstration that a significant amount of information can be gained from the Sb K-edge XANES region. Keywords: Antimony, XANES, high-energy XAS, linear combination fitting, sulfosalt

in t R o d u c t i o n there have been great gains in the research on many elements, Synchrotron-based X-ray absorption near-edge structure most notably in complex natural materials. The oxidation state, (XANES) analysis provides a powerful and unique method local coordination environment, and relative proportions of As of characterizing the oxidation state, chemical form, and local phases have been studied in mine tailings (Foster et al. 1998), coordination environment of an element associated with various in pyrite and secondary weathering phases (Savage et al. 2000), minerals and phases. XANES is an especially valuable tool when in contaminated reservoir sediment (Kneebone et al. 2002), in the sample volume is small, the sample is fine grained, and the soil (Cances et al. 2008), and in processed ore (Walker et al. element is present in low concentrations or is sensitive to redox 2005). X-ray absorption spectroscopy (XAS) has been applied changes during sample preparation. The oxidation state of an to Pb in mine tailings (Morin et al. 1999), mine-impacted sedi- element can be determined by measuring shifts in the absorp- ment (O’Day et al. 1998), and soils (Morin et al. 1999). Further tion edge to higher energies as the valence charge increases examples of elements probed using XAS in complex matrices (owing to an increase in the effective nuclear charge). The local include: Se (Pickering et al. 1995); Zn (O’Day et al. 1998); Mn coordination environment of elements as sorbed phases, or in (Ressler et al. 2000; Manceau et al. 1992); Fe (Patterson et al. minerals, liquids, and biological samples can be determined 1997; Savage et al. 2000; Manceau et al. 1992; Waychunas 1987); by comparing spectral features of model compounds with Ti (Waychunas 1987); Cr (Patterson et al. 1997); and Ni and Co well-defined structures, to the spectra of unknown compounds. (Manceau et al. 1992). Furthermore, using principal component analysis (PCA) and However, synchrotron research into elements requiring high- linear combination fitting (LCF), quantitative information can energy experimental configurations (those with higher atomic be obtained from XANES spectra (Ressler et al. 2000; Pickering numbers) has been limited due to the broad nature of high-energy et al. 1995; Foster et al. 1998). With advances in the applica- adsorption spectra. Spectral broadening occurs at higher energies tion of synchrotron science to mineralogy and sorbed phases, due to a short core-hole lifetime effect (D’Acapito et al. 2002), which results in poor resolution and reduced separation of the * Present address: Lorax Environmental, Vancouver, British edge positions of elements in different mineral phases. Never- Columbia V6J 3H9, Canada. E-mail: [email protected] theless, many researchers have obtained valuable results from

0003-004X/09/0010–1377$05.00/DOI: 10.2138/am.2009.3112 1377 1378 FAWCETT ET AL.: GAINING VALUABLE INFORMATION FROM THE Sb K-EDGE XANES REGION elements with higher K-edges including: Au at 80 725 eV and Yb noise ratio, and can be a valuable tool if the data are analyzed at 61 332 eV (D’Acapito et al. 2002); La at 38 925 eV (Kubozono carefully. This study focuses on extracting useful information et al. 2001; Yamamoto et al. 2001); Ce at 40 445 eV (Martin et from the XANES region, including the ability to identify those al. 2003; Quartieri et al. 2002); Nd at 43 569 eV (Quartieri et minerals in which Sb atoms occupy multiple crystallographic al. 2002); Te at 31 814 eV (Hayakawa et al. 2001); and Eu at sites within the crystal structure. Also novel in this study is the 48 519 eV and W at 69 525 eV (Borowski et al. 1999). In fact, attainment of quantitative information from unknown Sb spectra there are many problems, including investigations into complex using linear combination fitting analyses, and a detailed account matrices and environmental samples, which can be resolved only of how to reduce ambiguities in fitting analyses. Sb-sulfosalts and at the K-edges. For this reason the field of higher-energy XAS is stibnite mineral spectra are examined to expand on the spectral developing rapidly, most notably at third generation synchrotron database established by Scheinost et al. (2006). Ways in which sources (Quartieri et al. 2002). Higher K-edge elements have to improve spectral resolution, to quantify the limitations of Sb even been analyzed in more complex natural samples including XANES analyses in complex matrices, and to illustrate ways Cd (26 711 eV) complexation in bacteria (Kelly et al. 2001), and in which the experimenter can take advantage of high-energy Sn (29 200 eV) in sediment and antifouling paint (Sakakibara experimental designs, are presented. et al. 2004). Antimony is an environmentally significant, high K-edge ele- ba c k GR o u n d ment, found in anthropogenically impacted sediment (Brannon and Patrick 1985; Chen et al. 2003; Craw et al. 2004; Fillela et al. Advantages and challenges of high-energy XANES 2002), soil (Mitsunobu et al. 2006; Scheinost et al. 2006; Takaoka experiments et al. 2005; Lintschinger et al. 1998), mine tailings (Wilson et XAFS analyses on Sb are typically analyzed over the K-edge al. 2004a), surface water (Fillela et al. 2002a; Craw et al. 2004; (30 491 eV) or the L-edges (4132–4698 eV). The greatest advan- Wilson et al. 2004b), groundwater (Niedzielski and Siepak 2005), tage to working at higher energies is minimal (often non-existent) plants (He 2007; Murciego 2007), and biota (Koch et al. 2000). overlaps with other elements, or other edges of the same element. Antimony is a recognized carcinogen and is emerging as a pol- For example, in the study of the Sb-Te system, Hayakawa et al. lutant of priority interest (Council of the European Union 1998; (2001) found the Sb and Te L-edges to be too close to measure, U.S. Environmental Protection Agency 1979). Yet, broad gaps in but obtained successful analysis at their K-edges. At the Sb the knowledge of its geochemical behavior persist (Filella et al. K-edge, the next nearest element edge positions are Sn and Te, 2002a; Shotyk et al. 2005). Understanding the bonding character both separated from Sb by more than 1290 eV. Antimony L- of Sb is becoming increasingly critical. Extensive synchrotron- edges have numerous potential overlaps. Within 100 eV of the based investigations into the mobility and reactivity of Sb in Sb L-edges are the U M-edge, the Te and Sn L-edges, and most the environment could greatly improve our understanding of its notably, the intense Ca K-edge, an element almost ubiquitous in geochemical behavior. Synchrotron-based investigations into Sb most environmental samples. Distinguishing the Sb L-emission have been conducted on materials with simpler matrices includ- line in the presence of these elements is possible if a detector of ing catalysts (Millet et al. 2003), alloys (Hayakawa et al. 2001; sufficient resolution, such as a wavelength dispersive detector, Tani et al. 2001), nanocrystals (Rockenberger et al. 2000), and can be obtained. Only a few beamlines have the combination of hydrothermal solutions (Sherman et al. 2000). energy range and energy resolution to be able to study the Sb More recently, Sb was studied in soils. Antimony was found L-edges in the presence of the above-mentioned elements. Also, to persist exclusively in its oxidized form (Sb5+) in mine-impacted 30 keV X-rays are useful in studies relating to Sb in heteroge- soils (Mitsunobu et al. 2006; Takaoka et al. 2005) and in the neous systems because they probe deeper into the samples than Sb5+ and Sb0 oxidation states in shooting range soil (Scheinost do the 4 keV X-rays. et al. 2006). These researchers found, at most, two Sb oxidation- Conversely, there are many challenges associated with states in soil. However, the bonding character of Sb can be very high-energy XAS collection. The greatest challenge associated complex (Makovicky 1989; Takeuchi and Sadanaga 1969) and with Sb K-edge analysis is contending with poor edge-position other materials may contain Sb in compounds more difficult separation resulting from spectral broadening, the nature of the to recognize. In fact, the Sb mineral system includes multiple bonding environment, and the types of Sb minerals (i.e., mixed mixed-oxidation-state minerals and minerals in which Sb atoms oxidation state minerals whose edge position lies between the occupy multiple crystallographic sites (Amador et al. 1988; Sb3+-O and Sb5+-O edge positions). Bayliss and Nowacki 1972; Buerger and Hahn 1954; Kyono et al. 2002; Rouse et al. 1998; Thornton et al. 1977; Zubkova et Antimony mineralogy al. 2000). The accurate identification and quantification of Sb in In this study, XANES spectra were collected from Sb oxides, anthropogenically impacted environmental samples is especially Sb-sulfosalts, and the primary Sb-sulfide (stibnite). There are 3+ challenging as coexisting phases and complex hosting of Sb has four main oxides in the Sb-O system: Sb2 O3 (including senar- 5+ 3+ 5+ been observed (i.e., Fawcett 2009). Most of the XAS work on Sb montite and valentinite), Sb2 O5, Sb Sb O4 (alpha and beta to date uses the edge position to identify the oxidation state and form), and Sb6O13 (not discussed) (Earnshaw and Greenwood the extended X-ray absorption fine structure (EXAFS) region 1997). They are found most often as industrial products (i.e., to identify the phase (Scheinost et al. 2006; Mitsunobu et al. flame retardants, catalysts, by-products of high-temperature 2006; Takaoka et al. 2005). However, XANES collection is much processing), but can also be the product of the weathering of faster than EXAFS, has a higher signal, and a better signal-to- stibnite. Antimony atoms in Sb2O3 (senarmontite) are bonded in 1386 FAWCETT ET AL.: GAINING VALUABLE INFORMATION FROM THE Sb K-EDGE XANES REGION

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Modelling Bioremediation of Uranium Contaminated Aquifers

by Ben E. G. Rotter

A thesis submitted to the College of Science and Engineering in conformity with the requirements for the degree of Doctor of Philosophy

The University of Edinburgh Edinburgh, United Kingdom 2007

microbial residence. These observations may be valuable in the context of U remediation schemes, since DMRB tend to form nano-sized particles of uraninite (Suzuki et al.,

2002) which may undergo oxidative dissolution (Kelly et al., 2001). The presence of such uraninite particles in less accessible porous media regions of low flow velocity may prevent reoxidation or transportation of the mineral. In such cases, the results presented here would suggest that systems with bioactive immobile regions may hold advantages over those without. Whilst U(VI) immobilisation efficiency in systems with bioactivity in both regions is the same irrespective of the porosity ratio, the distribution of uraninite in the media may nevertheless vary depending on this porosity ratio. Furthermore, systems with bioactivity in both regions tend to exhibit uraninite precipitation further upgradient compared to those with bioactivity in the mobile region only or immobile region only.

Simulations revealed that spatially varying mineral presence can exert a significant impact of U(VI) bioreduction efficiency. Efficiency was found to be dependent on the combination of spatial mineral distribution and spatial microbial distribution. The results highlight the potential significance of spatially varying mineralogical and biological conditions within heterogeneous porous media and suggest that future research into the sensitivities of systems to mineralogy may prove to bring valuable insight to remediation efficiency.

Multi-regional models that assume microbial residence in both “mobile” and

“immobile” conceptual flow regions may significantly overestimate microbial efficiency

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heavy metal contamination of our environment. Over the Stability of Lead(II) Complexes of past two decades, there have been significant advances made Alginate Oligomers in efforts aimed at quantifying the nature of the physicochemical interactions occurring between heavy metals and biointerfaces (1). In particular, it is to a large degree the nature †,4 ,§ THOMAS A. DAVIS, of the biological chemistry, that is, the natural biopolymer ,‡ JOSE P. PINHEIRO,* composition and macromolecular conformation of the HANS GRASDALEN,§ OLAV SMIDSRØD,§ AND † constituents in the cell envelope that ultimately dictates, (i) HERMAN P. VAN LEEUWEN the relative strength and distribution of trace metal binding Laboratory of Physical Chemistry and Colloid Science, at the outer surface of a microorganism and, (ii) the gel Wageningen University, P.O. Box 8038, 6700 EK Wageningen, structure of that envelope which is, at least partially, related The Netherlands, Centro de Biomedicina Molecular e Estrutural, to the identity and variety of the particular metals bound Departamento de Química, Bioquímica e Farmácia, Fac. de therein. These two parameters are ultimately responsible Ciências e Tecnologia, Universidade do Algarve, Campus de for, or related to, a wide array of features such as metal Gambelas, 8005-139 Faro, Portugal, and Department of accumulation in the cell envelope (e.g. refs 2 and 3), Biotechnology, Norwegian University of Science and Technology, contaminant trace metal fate in groundwater systems (e.g. N-7491, Sem Sælands vei 6/8, Trondheim, Norway refs 4 and 5,), metal selectivity in biosorption (e.g. refs 6 and 7) or the binding of metals to plant root system cell walls Received September 26, 2007. Revised manuscript received (8, 9) to name only a few. November 30, 2007. Accepted December 12, 2007. The current work seeks to combine extraction, purification, and characterization techniques for uronic acid oligomers that are the building blocks for certain cell wall The current work reports on the Pb(II) complexes formed polysaccharides found in both bacterial and algal cell walls with oligomeric uronic acids (carboxylated saccharide residues) (6), with the electrochemical determination of their Pb(II) binding constants. Lead was chosen as the target metal ion, found polymerized in the cell walls and envelopes of algae not only for its contaminant history, but also because of its and bacteria alike. The application of partial acid hydrolysis, size- strong binding to carboxylates. Uronic acid oligomers were 1 exclusion chromatography (SEC), H NMR, and scanned chosen as the model cell wall polysaccharide material chiefly deposition stripping chronopotentiometry (SSCP) has permitted because of the dearth of literature available for this class of the determination of stability constants for Pb(II) with both naturally occurring charged polymer. Specifically, alginic acid mannuronic (M) and guluronic (G) acid oligomers ranging from is the common name given to a family of linear polysac- the dimer to the pentamer. The determined logarithm of the charides containing 1,4-linked -D-mannuronic (M) and R-L- stability constants range between 4.11 ( 0.05 and 5.00 ( 0.04 guluronic (G) acid residues arranged in a nonregular, mol-1 ·dm3 for the eight oligomers studied (pH 6; I ) 0.1 blockwise order along the chain (10). The residues typically mol ·dm-3). Additional experiments under the same experimental occur as (-M-)n, (-G-)n, and (-MG-)n sequences or blocks (Supporting Information Figure S1A). M- and G- block conditions employing galacturonic and glucuronic acid ( ( sequences differ significantly, and the proportions in which oligomers yielded slightly lower values (2.19 0.10 to 4.02 they are present in the alginate determine the physical -1 3 0.07 mol ·dm ) that were expected based on their structure, properties and reactivity of the polysaccharide (11). The whereby the monomers which were not included in the carboxylic acid dissociation constants of mannuronic and alginate oligomer series (unavailable by SEC), yielded the guluronic acids have been determined as pKa ) 3.38 and pKa lowest stability constants. This work demonstrates the applicability ) 3.65, respectively, with similar pKa values for the polymers of the SSCP technique for the determination of stability (12). Polymannuronic acid is a flat ribbon-like chain; its constants for metal–ligand complexes in which the ligands molecular repeat is 1.035 nm, and it contains two diequa- f display relatively low molecular mass. Previous studies on heavy torially (1e 4e) linked -D-mannuronic acid residues in metal interaction with the matrix polysaccharide alginate the chair conformation (13). In contrast, polyguluronic acid contains two diaxially (1a f 4a) linked R-L-guluronic acid have largely been restricted to the whole polymer that forms residues in the chair form which produces a rod-like polymer a gel upon binding to network bridging ions such as calcium. The with a molecular repeat of 0.87 nm (14). This key difference results will be discussed in this context with the emphasis in molecular conformation between the two homopolymeric being placed on the relevance of these findings to processes blocks is believed to be chiefly responsible for the variable occurring at the biointerface and results from the relevant affinity of alginates for heavy metals. Generally speaking, literature. alginates are quite abundant in nature since they occur as a structural component in marine brown algae (Phaeo- Introduction phyceae), comprising up to 40% of the dry matter, and as capsular polysaccharides in soil bacteria (6). The study of the physicochemistry of heavy metal binding Much of the early work on the interactions between taking place at biointerfaces represents an important part of aqueous cations and cell walls has been qualitative, deter- quantifying, modeling and assessing the risks associated with mining only the relative binding strength for different metals (e.g. refs 15–17). A more quantitative approach (18) deter- * Corresponding author e-mail: [email protected]. mined binding constants for interactions of metals with the † Wageningen University. 4 entire cell wall, albeit without differentiation between the Present address: Department of Chemistry, Université de various surface sites involved in the metal binding. Significant Montréal, C.P. 6128, succursale Centre-ville, Montreal, (QC) H3C- 3J7; [email protected]. progress was made by the work of Xue et al. (1988) (19) and ‡ Universidade do Algarve. Plette et al. (1995) (20) who constrained the values of § Norwegian University of Science and Technology. deprotonation constants for particular surface sites. In

10.1021/es702350w CCC: $40.75  2008 American Chemical Society VOL. 42, NO. 5, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 1673 Published on Web 01/31/2008 particular, Xue et al.’s (1988) work was significant because Materials and Methods they attempted to quantify the absolute stabilities of Materials and Reagents. All solutions were prepared in surface-metal complexes by employing thermodynamic ultrapure water from a Barnstead EasyPure UV system equilibrium constants for specific binding sites. The use of (resistivity >18 MΩ · cm). Pb(NO3)2 was prepared from acid–base titrations for the determination of deprotonation dilution of a certified standard (0.100 M Metrohm), NaNO3 constants for specific functional groups (4, 5, 21) allowed was prepared from solid NaNO3 (Merck, p.a.). Stock solutions more predictive models of metal binding to be developed as of MES (2-(N-morpholino) ethanesulfonic acid) buffers were a function of both pH and ionic strength. An essential feature prepared from the solids (Fluka, Microselect, >99.5%). of modeling metal binding lies in identifying the nature, Solutions prepared from HNO3 (Merck, suprapur) and NaOH quantity and conformation of specific binding sites present (0.1 M standard, Merck) were used to adjust the pH. for a given cell envelope (i.e., carboxylate, phosphate, etc.). D-Glucuronic acid Na salt was from Sigma; D-(+)-Galactu- Thus, detailed studies of the biochemical constitution of the ronic acid was from Sigma. 3-(trimethylsilyl)-propionic- outer surfaces of both microorganisms and algae alike (e.g. 2,2,3,3,-d4 acid Na salt (TSP) was from Aldrich (St. Louis MO). 2 2 refs 7, 22–26) are essentially linked to quantifying and NH4Ac was from SDS (Peypin, France). H2O and NaO H predicting metal binding behavior. were from Chiron (Trondheim, Norway); HCl was from Merck An ideal scenario for the quantification of thermodynamic (Darmstadt, Germany). The samples for electrochemical metal binding constants is the extraction, isolation, and analysis were removed from deep-freeze and placed several purification of whole biopolymers or oligomers of high purity days into a desiccator prior to weighing. The dissolved samples were prepared at a concentration ranging from 1 to which can then be used in controlled metal-binding experi- 4 mol · dm-3, in order to make the final required concentration ments. However, this is often difficult to achieve because of for the metal binding experiments. the low yields of highly purified material combined with the Mannuronan and Highly Enriched Guluronate Prepa- time intensive protocols necessary to obtain sufficient ration. The starting material was high molar mass mannu- quantities. On the analytical side, sensitive, electrochemical ronan that was isolated from a fermentation broth of a stripping techniques have been widely employed to study mannuronanC-5epimerase(AlgG)negativestrainofPseudomo- metal complexation with natural complexants (27, 28). nas fluorescens. The purification of this material and its However, their utility has been hampered by secondary effects deacetylation were carried out as previously described (32). such as adsorption of organic matter at the electrode surface. The absence of guluronate signals in the 1H NMR spectrum Recently, the attractive features of depletive stripping chro- (molar fraction FG < 0.001) confirmed that homopolymeric nopotentiometry, especially when operated at scanned mannuronan was formed (33). Alginate rich in G-blocks were deposition potential (SSCP) mode, have been demonstrated isolated from Laminaria hyperborea stipes and used as the (29). Depletive SSCP has a detection limit comparable to starting material. Successive fractional precipitation in that of pulsed stripping voltammetric modes, i.e. in the solutions of CaCl2 enriched the precipitated material in subnanomolar range, while being largely insensitive to guluronate according to previously described methodologies adsorption interferences (30, 31). In this work we test the (11, 34). This yielded a polydisperse preparation with a FG 1 application of SSCP on small quantities of purified oligomeric ) 0.94 with the remainder ) FM as determined by H NMR uronic acids isolated from either the stipes of the brown (35). algae Laminaria hyperborea or the mannuronan C-5 epi- Partial Acid Hydrolysis (for 1H NMR Analysis). Samples merase (AlgG) negative strain of Pseudomonas fluorescens. of either mannuronan or guluronate enriched alginate Isolation, purification, and characterization by 1H NMR of starting material were subjected to prehydrolysis (36). Further the uronic acid residues results in a small yet sufficient details on the procedure can be found on Page S5 of the quantity of highly purified and defined starting material Supporting Information. required for the Pb-binding experiments and, henceforth, Purified Oligomer Preparation. Oligomannuronic and determination of the stability constants, K. An additional oligoguluronic acid hydrolysis mixtures (80–100 mg) were objective was to determine the suitability of the SSCP filtered (0.45 µm, Millipore) and then chromatographed on technique for the determination of stability constants for two columns of preparative grade Superdex 30 (2.6 cm × 95 metal-complexes of low molar mass, yet biologically relevant cm, serially connected). The flow rate employed was 0.8 mL/ to the field of metal sequestration by cell walls or biofilms min in a supporting buffer of 0.1 M NH4Ac (pH 4.5) at room resulting from aggregation of cells. temperature. The relative concentrations of oligomers eluted This work focuses on the oligomeric equivalents of both from the column were determined with an online refractive mannuronic and guluronic acid ranging from the dimer to index (RI) detector (Shimazdu RID-6A). Many successive the pentamer (Supporting Information Figure S1B), since column runs of both oligomannuronic and oligoguluronic their binding behavior and physicochemical interactions with samples were eluted from the column, whereby fractions of 4 mL were collected and pooled on the basis of peaks metals has received much less attention than binding by identified from the RI profile. In order to ensure high purity, their larger polysaccharide equivalents and corresponding for each of the individual peaks, the lower third of both sides gels. These oligomers represent not only the fundamental of the sample peak were discarded. The samples were stored units of the alginates themselves, but they also occur as repeat at 4 °C during pooling and subsequently passed through at structures in the B-band O-side chains of lipopolysaccharides least five cycles of freeze-drying to remove all NH Ac and of Pseudomonas aeruginosa (PAO1), previously studied for 4 provide the solid purified oligomeric material. The final this strain’s metal binding capacity (25). freeze-dried samples were permanently stored at -70 °C. The results of the metal binding studies should be also 1H NMR. The freeze-dried oligomers (5 mg) were dissolved compared with experiments performed on commercially in deuterium oxide (600 µL) and the pH adjusted with NaO2H available uronic acids, namely galacturonic and glucuronic to 6.8; 1% TSP in deuterium oxide (5 µL) was also added as acid oligomers. The potential use of SSCP for the charac- an internal reference standard for some samples. 1-D 1H terization of metal binding constants with other highly NMR spectra were obtained on a Bruker DPX 400 spec- purified cell wall material is promising, and the results of trometer at 90 °C, using a 30° pulse angle, a spectral width this study are useful in determining the applicability for such of 4789 Hz, a 32 K data block size and either 128 or 256 scans. future studies of trace metal interactions with the biochemical Both the identity and purity of the oligomeric samples were constituents present in a wide array of biointerfaces. confirmed by comparing the ratio of integrated H-1 signals

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VOL. 42, NO. 5, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 1679 Appl Microbiol Biotechnol (2008) 79:293–299 DOI 10.1007/s00253-008-1415-4

ENVIRONMENTAL BIOTECHNOLOGY

Investigating the interaction mechanism between zinc and Saccharomyces cerevisiae using combined SEM-EDX and XAFS

Chen Can & Wang Jianlong

Received: 6 January 2008 /Revised: 12 February 2008 /Accepted: 12 February 2008 /Published online: 15 April 2008 # Springer-Verlag 2008

Abstract The interaction mechanism between zinc and the efficient remediation strategies and biotechnology treatment intact yeast cells of Saccharomyces cerevisiae was investigated facilities (Boyanov et al. 2003). Biosorption, using biomate- by using the scanning electron microscopy with energy- rials such as bacteria, fungi, yeast and algae, is regarded as a dispersive X-ray analysis, as well as X-ray absorption fine cost-effective biotechnology for the treatment of high-volume structure spectroscopy (XAFS). Displacement of H+,K+,Mg2 and low-concentration complex wastewaters containing +,andNa+ during zinc uptake confirmed the existence of both heavy metal(s). It has been reported that those biomass covalent interactions and ionic interactions between Zn2+ and materials can efficiently adsorb a variety of aqueous metal the microbe. Ion exchange mechanism played a role in zinc cations (Wang and Chen 2006). The macroscale research on uptake. The local environment of Zn accumulated in the intact metal ion sorption performance by the biomass was yeast cells was determined by XAFS, which suggests that the performed extensively. However, substantially less is known nearest neighboring atom of the bound zinc ion on the about the mechanisms of biosorption (Boyanov et al. 2003). biomass is oxygen atom. The adsorbed zinc ion on the intact Saccharomyces cerevisiae is ideal for mechanism research, cells of S. cerevisiae is a tetrahedron structure, with the Zn–O as well as for application in waste water treatment bond length of 1.97 Å, and the coordination number is only (Schneegurt et al. 2001; Naeem et al. 2006; Wang and Chen 3.2 of Zn–O structure in the first shell. 2006). Metallic zinc is one of the most important metals. Recovering or lowering the levels of zinc ion in the Keywords Zn2+ . Saccharomyces cerevisiae . Interaction . contaminated (waste)water is of environmental and agricul- XAFS . Mechanism tural importance (Taniguchi et al. 2000; Norton et al. 2004). Some methods such as Fourier transform infrared spectroscopy (FTIR), scanning electron microscopy Introduction (SEM), acid/base titrations, as well as surface complexation modeling method, were used to explore biosorption Heavy metal pollution is a serious environmental problem mechanisms (Fein et al. 1997; Wang and Chen 2006). We (Wang and Chen 2006). The released metals contaminate the have indicated that carboxyl played an important role in atmosphere, soils, and aquatic ecosystems. Microorganisms zinc uptake by FTIR. However, direct evidence of zinc in soils, sediments, and natural waters or in manmade speciation on the zinc loaded biomass was absent. Recently, treatment facilities influence the transport and fate of trace synchrotron-based X-ray absorption fine structure spectros- and contaminant metals. Exploring the metal–microbe copy (XAFS) provides a means for studying element interaction is a prerequisite for designing effective and complexation in environmental samples varying from mineralogical to biological in nature. XAFS, which consists : of two different complimentary techniques, the X-ray C. Can W. Jianlong (*) absorption near edge spectroscopy (XANES) and extended Laboratory of Environmental Technology, INET, X-ray absorption fine structure (EXAFS), is a technique Tsinghua University, Beijing 100084, People’s Republic of China increasingly used to study biological systems (Fomina et al. e-mail: [email protected] 2007), which can offer direct information of metal 298 Appl Microbiol Biotechnol (2008) 79:293–299

4 Fit Although Sarret et al. (1998) indicated that zinc phosphate Zn-cell sample dehydrate could offer good fit for zinc adsorption on 2 Penicillium chrysogenum cell walls, more reported results

3 suggest zinc uptake by cells or cell wall involving in 0 multifunctional groups such as phosphoester, carboxylate, x(k) k hydroxyl, and sulfhydryl (Webb et al. 2001; Guine et al. -2 2006). The disorder effect on atomic neighbors in the intact S. cerevisiae was in our study, which was observed to have -4 a coordination number of 3.2 for oxygen, as Sarret et al. 24681012 (1998) reported that coordination numbers for oxygen k /(0.1 nm)-1 range only between 3.2 and 2.3 for the adsorbed Zn(II) on Fig. 6 k3-Weighted back-transformed Zn–O peak (first shell) from the the fungal P. chrysogenum cell wall. They suggested that FT spectra of Zn(II) adsorbed onto the intact cells showing the fit of this kind of loss of atomic neighbors has no structural the theoretical curve to the experimental curve (solid line) reality, maybe because of the multiplicity of binding sites with slightly different chemical composition, interatomic Here, H+ displacement implied that Zn2+ exhibited a certain distances, and geometry. degree of covalent binding with the biomass (Brady and Some studies reported that Zn coordination to first shell Tobin 1955). Usually, the release of K+,Mg2+,Na+, and O usually falls into two categories, tetrahedral and Ca2+ from biosorbents in binding heavy metal ions was octahedral. The Zn–O bond length can be used as an regarded as an indicator of the mechanism of ion exchange indicator for octahedral and tetrahedral coordination (Sarret for heavy metal binding (Reddad et al. 2002). The result et al. 1998; Pan et al. 2004; Toner et al. 2006). The average here indicated that ion exchange played a role in zinc Zn–O distance in octahedral coordination is typically uptake. Ion exchange of K+,Mg2+,Na+,orCa2+ with Zn2+ 2.10 Å (range of 2.00–2.20 Å), whereas a value of 1.95 Å during sorption indicated a certain degree of ionic binding (range of 1.90–2.00 Å) is commonly reported in tetrahedral interaction between Zn and the biomass. coordination. Zinc EXAFS spectra collected from samples In our view, the reason of deformation of the intact yeast containing tetrahedrally coordinated Zn exhibit four O cells before biosorption was due to the cell living state. atoms in the first shell of the nearest neighboring atoms When live cells entered into bad-conditioned environment and an interatomic distance of approximately 1.97 Å (Sarret (in our study, the intact cells were cleaned and centrifuged et al. 1998; Bochatay and Persson 2000; Pan et al. 2004). and stored in deionized water without any buffer and The Zn–O bond length in solid zinc oxide ZnO(s) was nutrients) from cultural solution, they would change longer 1.96 Å. Toner et al. (2006) proposed that Zn–O interatomic from oviform to sausage form (Zheng et al. 1992). SEM-EDX distance of 2.07±0.02 Å is a weighted average distance clearly evidenced that Zn sorption, directly by the surface of between tetrahedral and octahedral Zn-sorbed species in the cells as well as precipitation containing zinc ions with biofilm-embedded Mn oxides. Kelly et al. (2001) suggest higher content than that adsorbed by the surface directly, is that the zinc center is 4 coordinate with mixed N/O ligation one of the biosorption mechanisms. It was reported that at an average distance of approximately 1.98 Å for the cells Aspergillus niger was grown in an industrial gold mining exposed to 250 and 500 mmol L−1 Zn2+ in plant Datura solution containing cyano-metal complexes in which gold, innoxia cells. Schneegurt et al. (2001) found that the silver, copper, iron, and zinc were accumulated, and the primary interaction between Zn ions and the dried yeast major mechanism of metal uptake involved metal and plant mixture from the production of ethanol from corn precipitation on the cell surface (Gomes et al. 1999). was via an oxygen atom. The fitting results of the data Precipitation of Cu, Pb, and Cd on the surface of the waste indicated an average Zn–O coordination number of 5±1, an S. cerevisiae was also observed by Marques et al. (2000). average Zn–O radial distance of 2.10±0.015 Å, and a During Zn uptake, not only inorganic matter such K and change in the XAFS Debye–Waller factor (relative to the Mg released but also organics such as protein probably zinc oxide standard) of −0.003±0.002 Å2. At the cell containing N element are also released from the biomass. surface for all diatom species studied, marine (Skeletonema Soares et al. (2002) used S. cerevisiae NCYC 1190 cells costatum) and freshwater (Achnanthidium minutissimum, that accumulated copper and found a leakage of UV- Navicula minima, and Melosira varians) diatoms, Zn was absorbing compounds and inorganic phosphate, which tetrahedrally coordinated with oxygen at similar to 2.00± could bind appreciable quantity of Cu. In our study, N 0.02 Å and monodentately bonded to one or two released was obvious but not for P reported by Soares et al. carboxylate groups (Pokrovsky et al. 2005). (2002). Whether the released organics in our study could In our study, the average Zn–O bond length of the three bind Zn was not clear and needs to be further studied. samples was 1.97 Å, indicating that the absorbed zinc ion Appl Microbiol Biotechnol (2008) 79:293–299 299 on the intact yeast cell of S. cerevisiae was the main Kelly RA, Andrews JC, DeWitt JG (2002) An X-ray absorption tetrahedron structure, and the nearest neighboring atom of spectroscopic investigation of the nature of the zinc complex – accumulated in Datura innoxia plant tissue culture. Microchem J the bound metal ion was an oxygen atom with the Zn O 71(2–3):231–245 bond length of about 1.97 Å, and coordination number was Marques P, Rosa MF, Pinheiro HM (2000) pH effects on the removal 2+ 2+ 2+ only 3.2. The results accorded with the XANES analysis of Cu ,Cd and Pb from aqueous solution by waste brewery – mentioned above. biomass. Bioprocess Engineering 23(2):135 141 Naeem A, Woertz JR, Fein JB (2006) Experimental measurement of proton, Cd, Pb, Sr, and Zn adsorption onto the fungal species Acknowledgments The work was financially supported by the Saccharomyces cerevisiae. Environ Sci Technol 40(18):5724– National Natural Science Foundation of China (Grant No. 5729 50278045) and the Basic Research Fund of Tsinghua University Norton L, Baskaran K, McKenzie T (2004) Biosorption of zinc from (Grant No. JC2002054). We thank Mr. Xie Yaning and Dr. Du aqueous solutions using biosolids. Adv Environ Res 8(3–4):629– Yonghua (Beijing Synchrotron Radiation Facility, Institute of High 635 Energy Physics, Beijing, China) for their help with X-ray absorption Pan G, Qin YW, Li XL et al (2004) EXAFS studies on adsorption– spectroscopy. desorption reversibility at manganese oxides–water interfaces I. Irreversible adsorption of zinc onto manganite (gamma-MnOOH). J Colloid Interface Sci 271(1):28–34 References Parsons JG, Aldrich MV, Gardea-Torresdey JL (2002) Environmental and biological application of extended X-ray absorption fine structure (EXAFS) and X-ray absorption near edge structure Bochatay L, Persson P (2000) Metal ion coordination at the water– (XANES) spectroscopies. Applied Spectroscopy Reviews 37 manganite (gamma-MnOOH) interface II. An EXAFS study of (2):187 zinc(II). J Colloid Interface Sci 229(2):593–599 Pokrovsky OS, Pokrovski GS, Gelabert A et al (2005) Speciation of Boyanov MI, Kelly SD, Kemner KM et al (2003) Adsorption of Zn associated with diatoms using X-ray absorption spectroscopy. cadmium to Bacillus subtilis bacterial cell walls: a pH-dependent Environ Sci Technol 39(12):4490–4498 X-ray absorption fine structure spectroscopy study. Geochim Reddad Z, Gerente C, Andres Y et al (2002) Adsorption of several Cosmochim Acta 67(18):3299–3311 metal ions onto a low-cost biosorbent: kinetic and equilibrium Brady JM, Tobin JM (1955) Binding of hard and soft metal ions to studies. Environ Sci Technol 36(9):2067–2073 Rhizopus arrhizus biomass. Enzyme Microb Technol 17(9):791– Sarret G, Manceau A, Spadini L et al (1998) Structural determination 796 of Zn and Pb binding sites in Penicillium chrysogenum cell walls Fein JB, Daughney CJ, Yee N et al (1997) A chemical equilibrium by EXAFS spectroscopy. Environ Sci Technol 32(11):1648–1655 model for metal adsorption onto bacterial surfaces. Geochim Schneegurt MA, Jain JC, Menicucci JA et al (2001) Biomass Cosmochim Acta 61(16):3319–3328 byproducts for the remediation of wastewaters contaminated Fomina M, Charnock JM, Hillier S et al (2006) Zinc phosphate with toxic metals. Environ Sci Technol 35(18):3786–3791 transformations by the Paxillus involutus/pine ectomycorrhizal Soares EV, Duarte A, Boaventura RA et al (2002) Viability and association. Microb Ecol 52(2):322–333 release of complexing compounds during accumulation of heavy Fomina M, Charnock J, Bowen AD et al (2007) X-ray absorption metals by a brewer’s yeast. Appl Microbiol Biotechnol 58(6): spectroscopy (XAS) of toxic metal mineral transformations by 836–841 fungi. Environ Microbiol 9(2):308–321 Taniguchi J, Hemmi H, Tanahashi K et al (2000) Zinc biosorption by a Gomes NCM, Figueira MM, Camargos ERS et al (1999) Cyano-metal zinc-resistant bacterium, Brevibacterium sp strain HZM-1. Appl complexes uptake by Aspergillus niger. Biotechnol Lett 21 Microbiol Biotechnol 54(4):581–588 (6):487–490 Toner B, Manceau A, Webb SM et al (2006) Zinc sorption to biogenic Guine V, Spadini L, Sarret G et al (2006) Zinc sorption to three gram- hexagonal-birnessite particles within a hydrated bacterial biofilm. negative bacteria: Combined titration, modeling, and EXAFS Geochim Cosmochim Acta 70(1):27–43 study. Environ Sci Technol 40(6):1806–1813 Tsezos M, Remoudaki E, Angelatou V (1995) A systematic study on Han RP, Zhang JH, Shi J et al (2004) Contribution of carboxylate equilibrium and kinetics of biosorptive accumulation—the case groups to lead biosorption by yeast. Journal of ZhengZhou of Ag and Ni. Int Biodeterior Biodegrad 35(1–3):129–153 University (Natural Science ed.) 36(1):86–91 Wang JL, Chen C (2006) Biosorption of heavy metal by Saccharo- Kazy SK, Das SK, Sar P (2006) Lanthanum biosorption by a Pseudomonas myces cerevisiae: a review. Biotechnol Adv 24(5):427–451 sp.: equilibrium studies and chemical characterization. J Ind Wang QW, Liu WH (1994) X ray absorption fine structure and its Microbiol Biotech 33(9):773–783 application. Science Press, Beijing Kelly SD, Boyanov MI, Bunker BA et al (2001) XAFS determination Webb SM, Gaillard JF, Jackson BE et al (2001) An EXAFS study of zinc of the bacterial cell wall functional groups responsible for coordination in microbial cells. J Synchrotron Radiat 8:943–945 complexation of Cd and U as a function of pH. J Synchrotron Zheng SL, Hu BL, Sheng ZD et al (1992) Basis of microbiolosy. Radiat 8:946–948 Chemical Industry, Beijing, China View Article Online / Journal Homepage / Table of Contents for this issue

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Interaction of uranium(VI) with lipopolysaccharide†

Astrid Barkleit,* Henry Moll and Gert Bernhard

Received 11th October 2007, Accepted 14th March 2008 First published as an Advance Article on the web 18th April 2008 DOI: 10.1039/b715669c

Bacteria have a great influence on the migration behaviour of heavy metals in the environment. Lipopolysaccharides form the main part of the outer membrane of Gram-negative bacteria. We 2+ investigated the interaction of the uranyl cation (UO2 ) with lipopolysaccharide (LPS) from Pseudomonas aeruginosa by using potentiometric titration and time-resolved laser-induced fluorescence spectroscopy (TRLFS) over a wide pH and concentration range. Generally, LPS consists of a high density of different functionalities for metal binding such as carboxyl, phosphoryl, amino and hydroxyl groups. The dissociation constants and corresponding site densities of these functional groups were determined using potentiometric titration. The combination of both methods, potentiometry and TRLFS, show that at an excess of LPS uranyl phosphoryl coordination dominates, whereas at a slight deficit on LPS compared to uranyl, carboxyl groups also become important for uranyl coordination. The stability constants of one uranyl carboxyl complex and three different uranyl phosphoryl complexes and the luminescence properties of the phosphoryl complexes are reported.

2+ Introduction In this paper, we present the results of uranium(VI)(UO2 ) complexation studies with LPS from P. aeruginosa S10. The Microorganisms are very important for the bioremediation of complexation was studied on the one hand by potentiometric the environment because they are able to adsorb radionuclides titration with a slight deficit to nearly equal metal and ligand and other heavy metals. They significantly influence mobilization concentrations and on the other hand by time-resolved laser- 1 and immobilization of metal ions in soils. Some bacteria, like induced fluorescence spectroscopy (TRLFS) with a very low the Gram-negative Pseudomonas aeruginosa, are to be found uranyl concentration and an excess of LPS, which simulates 2 under harsh environmental conditions like uranium waste piles environmental conditions. 3 or sewage from uranium mill tailings. To predict the radionuclide TRLFS is a very convenient tool for detecting and char- transport in the environment and to improve technical bioremedi- acterizing uranyl speciation, even at very low concentrations, ation strategies, understanding of the binding mechanisms on the because of its high sensitivity. The method detects changes in molecular level is essential. the chemical environment of the fluorescent actinide through For Gram-negative bacteria, the main binding sites for metal shifts of fluorescence spectra and variations of the lifetime of Published on 18 April 2008. Downloaded by University of Notre Dame 23/08/2014 20:09:50. ions are carboxyl, phosphoryl, hydroxyl and amino groups, located the excited state.10,11 The obtained spectroscopic data (spectra, 1 in the lipopolysaccharide (LPS) unit in the outer membrane. peak maxima, intensities and lifetimes) from model complexes Various investigations with bacteria show that phosphoryl groups give valuable information identifying the speciation of actinides in 4,5 seem to be quite effective for uranium binding. Some studies bio-systems.10,12–14 assume that carboxyl groups are also involved in uranium binding.5 There are few studies examining systematically the complex behaviour of bacterial cell-wall components that provide Experimental the potential binding sites for the uranyl ion.6 Some investigations dealing with metal binding on LPS are published,7 and they all Potentiometric titrations suppose phosphoryl groups as the main binding partner. However, The potentiometric titration experiments were carried out in a the exact mechanisms of the interactions between cell surfaces and glove box under inert gas atmosphere (nitrogen) at 22 ± 1 ◦C. In metals remain unclear and, to our knowledge, investigations of each case, the starting volume was 50 mL. For each ligand titration, uranium complexes with LPS have not been done so far. 10 mg LPS (lipopolysaccharide from Pseudomonas aeruginosa Fig. 1 demonstrates the structure of LPS from P. aeruginosa S10, prepared by trichloroacetic acid extraction, purchased from (based on8,9). The whole LPS molecule can be divided into three Sigma) was dissolved with carbonate-free deionized water. The main parts: the lipid A, the core region and the repeating unit, ionic strength was kept constant at 0.1 M by adding NaClO ·H O the O-antigen. The core region mainly contains a high density of 4 2 (Merck, p.A.). 3 samples were acidified with HClO (carbonate- phosphoryl and carboxyl groups. 4 free) to obtain a starting pH of about 4 and were titrated with 1 mM NaOH (carbonate-free, Merck, Titrisol), and 3 samples Institute of Radiochemistry, Forschungszentrum Dresden-Rossendorf e.V., were alkalized with NaOH to obtain a starting pH of about P.O box 510119, D-01314, Dresden, Germany. E-mail: [email protected]; Fax: +49 (0)351 260 3553; Tel: +49 (0)351 260 2148 10andweretitratedwith1mMHClO4 (carbonate-free; the † Electronic supplementary information (ESI) available: Summary of all exact molarity was determined with 0.01 M NaOH, Merck, TRLFS measurement series (Table S1). See DOI: 10.1039/b715669c Titrisol). For complexation titration (5 measurements), in each

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uranyl ion. The stabilization of the LPS assembly in the outer Fein, D. A. Fowle, N. Yee and K. M. Kemner, J. Synchrotron Radiat., membrane through divalent metal ions like Ca or Mg, both intra- 2001, 8, 946–948. 42,43 5 M. L. Merroun, J. Raff, A. Rossberg, C. Hennig, T. Reich and S. and intermolecular, is established experimentally as well as Selenska-Pobell, Appl. Environ. Microbiol., 2005, 71, 5532–5543; S. D. by computer simulation.3,42,44 For example, it was found that on Kelly, K. M. Kemner, J. B. Fein, D. A. Fowle, M. I. Boyanov, B. A. average a Ca 2+ ion interacts beneath carboxyl and hydroxyl with Bunker and N. Yee, Geochim. Cosmochim. Acta, 2002, 66, 3855–3871. three phosphates.44 These findings also support the assumption 6 J. D. Van Horn and H. Huang, Coord. Chem. Rev., 2006, 765– 775. of the coordination of uranyl with more than one phosphoryl 7 F. G. Ferris and T. J. Beveridge, Can. J. Microbiol., 1986, 32, 52–55; group. B. Klapcinska, Can. J. Microbiol., 1994, 40, 686–690; S. Langley and T. J. Beveridge, Appl. Environ. Microbiol., 1999, 65, 489–498; E. T. Oh, H. S. Yun, T. R. Heo, S. C. Koh, K. H. Oh and J. S. So, J. Microbiol. Conclusion Biotechnol., 2002, 12, 296–300; S. M. Strain, S. W. Fesik and I. M. Armitage, J. Biol. Chem., 1983, 258, 3466–3477. The combination of potentiometric titration and time-resolved 8 Y. A. Knirel, CRC Crit. Rev. Microbiol., 1990, 17, 273–304. laser-induced fluorescence spectroscopy for the study of the 9 Y.A. Knirel, O. V. Bystrova, N. A. Kocharova, U. Zahringer and G. B. Pier, J. Endotoxin Res., 2006, 12, 324–336. complex behaviour between the uranyl ion and LPS provides a 10 G. Geipel, Coord. Chem. Rev., 2006, 250, 844–854. broad overview over a wide concentration and pH range. With 11 C. Moulin, Radiochim. Acta, 2003, 91, 651–657; C. Moulin, I. potentiometry a nearly equimolar ratio of uranyl and functional Laszak, V. Moulin and C. Tondre, Appl. Spectrosc., 1998, 52, 528– 535. groups of LPS was studied, whereas with TRLFS the behaviour 12 A. Gunther,¨ G. Bernhard, G. Geipel, T. Reich, A. Rossberg and H. at more environmentally relevant conditions, relatively low uranyl Nitsche, Radiochim. Acta, 2003, 91, 319–328. concentration and with an excess of LPS was examined. As a result, 13 A. Gunther,¨ G. Geipel and G. Bernhard, Radiochim. Acta, 2006, 845– we observed a coordination of the uranyl ion with both carboxyl 851. 14 A. Koban and G. Bernhard, Polyhedron, 2004, 23, 1793–1797; A. Koban and phosphoryl groups. While at conditions with a slight deficit and G. Bernhard, J. Inorg. Biochem., 2007, 101, 750–757. on phosphoryl groups carboxyl groups also play an important role 15 G. Geipel, A. Brachmann, V. Brendler, G. Bernhard and H. Nitsche, for uranyl coordination, the coordination via phosphoryl groups Radiochim. Acta, 1996, 75, 199–204. 16 B. F. Turner and J. B. Fein, Comput. Geosci., 2006, 32, 1344–1356. dominates with ligand excess. These findings support the rare 17 B. F. Turner, ProtoFit Version 2.0, a program for determining surface studies that suggest an involvement of carboxyl groups in uranyl speciation constants from titration data, University of Notre Dame, coordination,5 while the general assumption is that in microbial Notre Dame, USA, Department of Civil Engineering and Geological systems phosphoryl groups are the main binding partner for the Sciences, 2005. 4 18 J. S. Cox, D. S. Smith, L. A. Warren and F. G. Ferris, Environ. Sci. uranyl ion. Fig. 4 depicts for demonstration the speciation of the Technol., 1999, 33, 4514–4521. observed complex species at both edge conditions: equimolar and 19 N. Yee and J. B. Fein, Geochim. Cosmochim. Acta, 2001, 65, 2037– excess of ligand. 2042. Anyway, the binding of uranyl to LPS seems to be very strong, 20 A. C. Texier, Y.Andres, M. Illemassene and P. Le Cloirec, Environ. Sci. Technol., 2000, 34, 610–615. and a potential use for bioremediation technologies should not be 21 R. E. Martinez, D. S. Smith, E. Kulczycki and F. G. Ferris, J. Colloid disregarded. Interface Sci., 2002, 253, 130–139. 22 D. M. Borrok and J. B. Fein, J. Colloid Interface Sci., 2005, 286, 110– 126. Acknowledgements 23 P. Gans, A. Sabatini and A. Vacca, Talanta, 1996, 43, 1739–1753.

Published on 18 April 2008. Downloaded by University of Notre Dame 23/08/2014 20:09:50. 24 D. C. Harris, Quantitative Chemical Analysis,W.H.Freeman,and This work was funded by the BMWi under contract number Company, New York, 3rd edn, 1991. 25 R. Guillaumont, T. Fanghanel,¨ J. Fuger, I. Grenthe, V. Neck, D. A. 02E9985. Furthermore we thank J. Schott for carrying out the Palmer and M. H. Rand, Update on the Chemical Thermodynamics of potentiometric titration experiments and U. Schaefer for ICP-MS Uranium, Neptunium, Plutonium, Americium and Technetium, Elsevier, measurements. Amsterdam, 2003. 26 S. Sachs, V. Brendler and G. Geipel, Radiochim. Acta, 2007, 95, 103– 110. References 27 A. Koban, G. Geipel, A. Rossberg and G. Bernhard, Radiochim. Acta, 2004, 92, 903–908. 1 T. J. Beveridge and R. J. Doyle, Metal ion and bacteria, John Wiley & 28 C. Jacopin, M. Sawicki, G. Plancque, D. Doizi, F.Taran, E. Ansoborlo, Sons, Inc., New York, N.Y., 1989. B. Amekraz and C. Moulin, Inorg. Chem., 2003, 42, 5015–5022. 2 P. Panak, S. Selenska-Pobell, S. Kutschke, G. Geipel, G. Bernhard and 29 H. Moll, G. Geipel, T.Reich, G. Bernhard, T.Fanghanel¨ and I. Grenthe, H. Nitsche, Radiochim. Acta, 1999, 84, 183–190. Radiochim. Acta, 2003, 91, 11–20; A. Gunther,¨ G. Geipel and G. 3 M. R. Shroll and T. P. Straatsma, Biopolymers, 2002, 65, 395–407. Bernhard, Polyhedron, 2007, 26, 59–65. 4 P. Panak, J. Raff, S. Selenska-Pobell, G. Geipel, G. Bernhard and H. 30 G. Bernhard, G. Geipel, T. Reich, V. Brendler, S. Amayri and H. Nitsche, Radiochim. Acta, 2000, 88, 71–76; P. J. Panak, R. Knopp, Nitsche, Radiochim. Acta, 2001, 89, 511–518. C. H. Booth and H. Nitsche, Radiochim. Acta, 2002, 90, 779–783; 31 S. Scapolan, E. Ansoborlo, C. Moulin and C. Madic, J. Radioanal. M. Merroun, C. Hennig, A. Rossberg, G. Geipel, T. Reich and S. Nucl. Chem., 1997, 226, 145–148. Selenska-Pobell, Biochem. Soc. Trans., 2002, 30, 669–672; M. Merroun, 32 A. Brachmann, G. Geipel, G. Bernhard and H. Nitsche, Radiochim. C. Hennig, A. Rossberg, T. Reich and S. Selenska-Pobell, Radiochim. Acta, 2002, 90, 147–153. Acta, 2003, 91, 583–591; M. Merroun, M. Nedelkova, A. Rossberg, C. 33 I. Bonhoure, S. Meca, V. Marti, J. De Pablo and J. L. Cortina, Hennig and S. Selenska-Pobell, Radiochim. Acta, 2006, 94, 723–729; Radiochim. Acta, 2007, 95, 165–172. M. L. Merroun, G. Geipel, R. Nicolai, K. H. Heise and S. Selenska- 34 S. Scapolan, E. Ansoborolo, C. Moulin and C. Madic, J. Alloys Compd., Pobell, BioMetals, 2003, 16, 331–339; M. Nedelkova, M. L. Merroun, 1998, 271, 106–111. A. Rossberg, C. Hennig and S. Selenska-Pobell, FEMS Microbiol. 35 G. Geipel, G. Bernhard, M. Rutsch, V. Brendler and H. Nitsche, Ecol., 2007, 59, 694–705; R. Knopp, P. J. Panak, L. A. Wray, N. S. Radiochim. Acta, 2000, 88, 757–762. Renninger, J. D. Keasling and H. Nitsche, Chem.–Eur. J., 2003, 9, 2812– 36 R. A. Binstead, A. D. Zuberbuhler¨ and B. Jung, SPECFIT Global 2818; C. Hennig, S. Selenska-Pobell, W.Matz and P.Panak, Radiochim. Analysis System, Version 3.0.37, Spectrum Software Associates, Marl- Acta, 2001, 89, 625–631; S. D. Kelly, M. I. Boyanov, B. A. Bunker, J. B. borough, USA, 2005.

This journal is © The Royal Society of Chemistry 2008 Dalton Trans., 2008, 2879–2886 | 2885 Chemical Geology 244 (2007) 493–506 www.elsevier.com/locate/chemgeo

Divalent metal adsorption by the thermophile Anoxybacillus flavithermus in single and multi-metal systems ⁎ Peta-Gaye G. Burnett a, Kim Handley b, Derek Peak a, , Christopher J. Daughney c

a Department of Soil Science, University of Saskatchewan, 51 Campus Drive, Saskatoon, SK Canada S7N 5A8 b Department of Earth Sciences, University of Manchester, Manchester M13 9PL, United Kingdom c Institute of Geological and Nuclear Sciences, 1 Fairway Drive, Avalon, Lower Hutt, New Zealand Received 3 July 2007; accepted 6 July 2007

Editor: D. Rickard

Abstract

Recent studies have applied surface complexation models (SCMs) to describe metal adsorption by mesophilic bacteria. However, only one SCM has been developed for metal biosorption by a thermophile and the study was limited to adsorption of Cd. In this study, we quantify adsorption of a variety of metals onto the thermophile Anoxybacillus flavithermus with surface complexation modeling. We conduct both single and multi-metal sorption studies using Cd, Cu, Mn, Ni, Pb, and/or Zn, in order to compare the relative affinities of the metals for the cell surface and investigate if these individual affinities change in the presence of other metals. We also use linear free energy relationships (LFERs) to compare the stability constants for the metal-bacteria surface complexes with a variety of aqueous metal–ligand stability constants, and we compare the metal binding capacity of Cd, Cu, Mn, Pb, Ni, and Zn of A. flavithermus to that of the commonly studied mesophile Bacillus subtilis. We find that the metals exhibit different preferences for the bacterial surface in the order Mn≈NibZnbCdbPb≈Cu. Metal binding is described best by formation of a M–carboxyl complex and either a MOH–carboxyl, M–phosphoryl, or MOH– phosphoryl complex (where M is the metal cation). Stability constants determined from the single metal and multi-metal systems are comparable. It is therefore possible to obtain reasonable stability constants in multi-metal systems. This has the potential to greatly simplify the acquisition of metal-bacteria thermodynamic data. However, stability constants obtained from the single metal systems remain more accurate because the experimental conditions are more ideal. Under our experimental conditions, competition effects are observed for the low affinity metals in the presence of higher affinity metals. The LFER plot using the formation of M– carboxyl and MOH–carboxyl with M–acetate complexes yields linear correlation coefficients (r) of 0.82 and 0.73, respectively. We also observe that A. flavithermus generally adsorbs less metal than B. subtilis at similar metal-to-biomass concentration ratio over the pH range studied. © 2007 Elsevier B.V. All rights reserved.

Keywords: Metal adsorption; Surface complexation modeling; Bacteria; Thermophile; Linear free energy relationship

1. Introduction

Microbial surfaces have several types of organic ⁎ Corresponding author. Fax: +1 306 966 6881. functional groups (carboxylic, phosphoryl, hydroxyl, E-mail address: [email protected] (D. Peak). amino, etc.) that can react with dissolved metals.

0009-2541/$ - see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.chemgeo.2007.07.006 P.-G.G. Burnett et al. / Chemical Geology 244 (2007) 493–506 497 be significant even for bacteria cultured in the same Burnett et al. (2006b) for the adsorption of Cd onto manner. Hence additionally, V(Y)s for SCMs that A. flavithermus and by Fowle and Fein (1999) for the attempt to simultaneously fit data from many cultures adsorption of Cu and Pb onto B. subtilis and are often higher. Similarly, these values also depend on B. licheniformis. (Note that Fowle and Fein (1999) the pH range and biomass concentration that the data included a third reaction in their SCM). Most other span. For single metal adsorption systems, SCM values investigations of metal–bacteria adsorption have pro- for stability constants were determined for several posed SCMs involving metal adsorption to carboxyl and independent batch adsorption sets (independent cul- to phosphoryl sites within the cell wall (Fein et al., 1997; tures), and the results were used to calculate weighted Daughney and Fein, 1998; Daughney et al., 1998; Sarret 2σ uncertainties. For multi-metal adsorption systems, et al., 1998; Fowle and Fein, 2000; Fowle et al., 2000; errors were generated by consideration of data sets Daughney et al., 2001; Fein et al., 2001; Kelly et al., having different biomass concentrations rather than 2001, 2002; Ngwenya et al., 2003; Borrok et al., 2004a). from data pertaining to independent cultures. Our best-fitting SCM assumes that metal adsorption involves only carboxyl sites within the cell wall, but it 3. Results and discussion allows for adsorption of the first metal hydrolysis product. However, we acknowledge that the two-site 3.1. Adsorption of metals in single metal systems two-reaction SCMs, that is, the formation of 1) M– carboxyl and M–phosphoryl or 2) M–carboxyl and The single metal adsorption experiments indicate that MOH–phosphoryl fit the data almost as well. Addition- all five metals (Cu, Mn, Ni, Pb, and Zn) bind to A. ally, we acknowledge that it is possible that only some of flavithermus (Fig. 1). As thoroughly documented in the the metals may preferentially bind in M–phosphoryl literature for metal adsorption onto mesophiles, these configurations whereas others are present as a hydroly- metals show similar relationships between pH, biomass sis product. This can be observed for Ni, where concentration, and percent metal adsorbed, that is, complexation with phosphoryl groups produces a increasing metal adsorption with increasing pH and slightly better fit. The model we have chosen performs bacteria-to-metal concentration ratio. Comparable ad- better when all metals are considered together with the sorption behaviour has been reported for the sorption of minimum amount of complexity. It is possible that it will other divalent metals onto different microbial surfaces need to be modified to include additional complexes if (Fein et al., 1997; Daughney and Fein, 1998; Daughney spectroscopic studies determine that phosphoryl groups et al., 1998; Fowle and Fein, 1999; Daughney et al., are participating with some metals. 2001; Yee and Fein, 2003; Borrok et al., 2004a,b; Yee To determine if the reactions proposed in our SCM et al., 2004; Borrok and Fein, 2005). For a given are occurring, spectroscopic examination of the biomass concentration and pH, the quantity of metal adsorbed metal is required. Burnett et al. (2006b) adsorbed (per gram bacteria) follows the order: examined Cd adsorption onto A. flavithermus and were Mn≈ NibZnbCdbPb≈ Cu. Note that the relative able to verify, via X-ray Absorption Spectroscopy, affinity of Cd is based on results by Burnett et al. formation of a Cd-carboxyl surface complex. However, (2006b). A similar affinity series has been reported for their data did not corroborate a second reaction metal adsorption onto B. subtilis and B. licheniformis mechanism. Spectroscopic evidence may be able to (Daughney and Fein, 1998; Daughney et al., 1998) and confirm if the reaction mechanisms are the same for all Padina sp. (Sheng et al., 2004). six metals, as is implied by the batch macroscopic For most of the metals investigated here, the best- adsorption data. If the same reaction mechanisms are fitting SCM included the formation of M–carboxyl required to account for metal adsorption as physio- (RCOOM+) and MOH–carboxyl (RCOOMOH0) com- chemical parameters are varied, regardless of the plexes (Eqs. (3) and (4) below): identity of the six metals investigated here, then the SCM may be applicable to a range of aqueous divalent − R–COO þ M2þ⇔R–COOMþ ð3Þ metal species and geochemical conditions. Bivariate scattergrams were used to compare the − 2þ 0 þ R–COO þ M þ H2O⇔R–COOMOH þ H ð4Þ predictions from the overall best-fitting SCMs (Eqs. (3) and (4)) to the experimental data for each metal (Fig. 2) In all cases with the exception of Mn, a two-reaction and principal axes analyses were conducted using SCM provided the best fits to the experimental data formulations described in Sokal and Rohlf (1995) (Table 2). These reactions have been proposed by (Table 3). Graphically, the SCMs are able to predict 506 P.-G.G. Burnett et al. / Chemical Geology 244 (2007) 493–506

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Zinc sorption to biogenic hexagonal-birnessite particles within a hydrated bacterial bioﬁlm

Brandy Toner a,*, Alain Manceau b, Samuel M. Webb c, Garrison Sposito a a Department of Environmental Science, Policy and Management, Division of Ecosystem Sciences, University of California, Berkeley, CA 94720-3114, USA b Environmental Geochemistry Group, L.G.I.T.-Maison des Geosciences, Universite J. Fourier, B.P. 53, 38041 Grenoble Cedex 9, France c Stanford Synchrotron Radiation Laboratory, Bldg. 137, MS 69, 2575 Sand Hill Rd., Menlo Park, CA 94025, USA

Received 14 February 2005; accepted in revised form 8 August 2005

Abstract

Biofilm-embedded Mn oxides exert important controls on trace metal cycling in aquatic and soil environments. The speciation and mobility of Zn in particular has been linked to Mn oxides found in streams, wetlands, soils, and aquifers. We investigated the mechanisms of Zn sorption to a biogenic Mn oxide within a biofilm produced by model soil and freshwater MnII-oxidizing bacteria Pseudo- monas putida. The biogenic Mn oxide is a c-disordered birnessite with hexagonal layer symmetry. Zinc adsorption isotherm and Zn and Mn K-edge extended X-ray absorption fine structure (EXAFS) spectroscopy experiments were conducted at pH 6.9 to characterize Zn sorption to this biogenic Mn oxide, and to determine whether the bioorganic components of the biofilm affect metal sorption properties. The EXAFS data were analyzed by spectral fitting, principal component analysis, and linear least-squares fitting with reference spectra. Zinc speciation was found to change as Zn loading to the biosorbent [bacterial cells, extracellular polymeric substances (EPS), and biogenic Mn oxide] increased. At low Zn loading (0.13 ± 0.04 mol Zn kg1 biosorbent), Zn was sorbed to crystallographically well-defined sites on the biogenic oxide layers in tetrahedral coordination to structural O atoms. The fit to the EXAFS spectrum was consistent with Zn sorption above and below the MnIV vacancy sites of the oxide layers. As Zn loading increased to 0.72 ± 0.04 mol Zn kg1 biosorbent, Zn was also detected in octahedral coordination to these sites. Overall, our results indicate that the biofilm did not intervene in Zn sorption by the Mn-oxide because sorption to the organic material was observed only after all Mn vacancy sites were capped by Zn. The organic functional groups present in the biofilm contributed significantly to Zn removal from solution when Zn concentrations exceeded the sorption capacity of the biooxide. At the highest Zn loading studied, 1.50 ± 0.36 mol Zn kg1 biosorbent, the proportion of total Zn sorption attributed to bioorganic material was 38 mol%. The maximum Zn loading to the biogenic oxide that we observed was 4.1 mol Zn kg1 biogenic Mn oxide, corresponding to 0.37 ± 0.02 mol Zn mol1 Mn. This loading is in excellent agreement with previous estimates of the content of cation vacancies in the biogenic oxide. The results of this study improve our knowledge of Zn speciation in natural systems and are consistent with those of Zn speciation in mineral soil fractions and ferromanganese nodules where the Mn oxides present are possibly biogenic. 2005 Elsevier Inc. All rights reserved.

1. Introduction phores), (3) alteration of chemical microenvironment (e.g., O2 and pH), (4) redox reactions with metal electron donors Microorganisms in soils, sediments, and natural waters and acceptors, and (5) mineral formation and dissolution influence the transport and fate of trace and contaminant (Manceau and Combes, 1988; Bargar et al., 2000; Kraemer metals through: (1) incorporation into biomass during et al., 2002; Boyanov et al., 2003; Nelson and Lion, 2003; growth, (2) metal complexation by cell/spore surfaces, Wang et al., 2003; Kemner et al., 2004). The precipitation extracellular polymers, and diffusible exudates (e.g., sidero- and subsequent surface reactivity of biogenic minerals is an expanding field of study, as these minerals are often * Corresponding author. small reactive particles that are spatially distributed as E-mail address: [email protected] (B. Toner). coatings at interfaces where biofilms form (Templeton

The solid material of the Zn-sorbed samples was collect- 3–12 A˚ 1. The real and imaginary parts of the Zn ed by centrifugation. The sample pastes were stored frozen FT[v(k)k3] were fit with phase and amplitude functions cal- until the time of analysis. The effect of sample freezing on culated with FEFF6 (Rehr et al., 1992) from the crystal Zn speciation was not examined. The sample material was structure of chalcophanite (Post and Appleman, 1988). packed into a PCTFE sample holder with Lexan windows The quality of each fit was evaluated quantitatively with 2 and the sample holder was sealed with Kapton tape. The the reduced chi square ðvm Þ and R-factor parameters (Ra- Lexan windows were used to prevent the Kapton tape vel, 2000). All fitted parameters [interatomic distance, R ˚ adhesive from reacting with MnOx species. The samples (A), coordination number, N, and mean-square displace- were kept moist and stored on ice until the EXAFS spectra ment of bond length, r2 (A˚ 2)] are presented with the esti- were collected. mated errors from IFEFFIT calculations (Newville, Zinc acetate, Zn phytate, Zn citrate, and Zn sorbed to 2001). The error estimates reported by the IFEFFIT fitting dMnO2 (Zn dMnO2) reference materials were prepared. routine are the diagonal elements of the covariance multi- 2 1=2 The organic reference materials were chosen to reflect the plied by ðvm Þ . The atom pairs used to model the Zn EX- anticipated functional groups presented to solution by AFS spectra were linked such that a single value for the Gram-negative bacterial outer membranes (Fein et al., change in threshold energy, DE0 (Teo and Joy, 1981), 2001; Kelly et al., 2001) and EPS, as well as those used was obtained for each fit. In fits to both reference and 2 for similar studies (Sarret et al., 1998). Zinc dMnO2 was experimental spectra the amplitude reduction factor ðS0Þ prepared as a reference material for Zn sorbed to a speci- was set to 0.86, as determined by fitting of the EXAFS men MnIV oxide that has structural characteristics similar spectrum of chalcophanite. to the biogenic Mn oxide produced by P. putida (Villalobos The Zn EXAFS data set was subjected to principal com- et al., 2003). A summary of the conditions used to produce ponent analysis (PCA) to aid in determining the number the reference materials is presented in Table 1. Unlike Zn and identity of the Zn species present in the Zn sorbed sam- acetate and Zn citrate, the stability constants for Zn phy- ples. The reader is referred to Manceau et al. (2002b) for a tate were not readily available. For Zn K-edge EXAFS detailed treatment of the use of PCA in EXAFS spectro- spectra collection, Zn acetate, Zn phytate, and Zn citrate scopic analysis. The number of principal components rep- complexes were added to a PCTFE sample holder with resented by the data set was determined using the Lexan windows as solutions, and the sample holder was minimum of the indicator function (Malinowski, 1977). sealed with Kapton tape. The dMnO2 particles were col- Target transformation analyses were conducted to evaluate lected by vacuum filtration, and the resulting wet paste the suitability of the Zn reference spectra for inclusion in was packed into a PCTFE sample holder with Lexan win- the spectral data set using the SPOIL parameter (Malinow- dows and sealed with Kapton tape. The Zn EXAFS spec- ski, 1978). Values for SPOIL of less than 1.5 are considered trum of the Zn–Mn oxide chalcophanite was used as a excellent, 1.5–3 good, 3–4.5 fair, and 4.5–6 poor (Malinow- reference in this study; the oxide has been characterized ski, 1978). The PCA was performed with unsmoothed, elsewhere (Manceau et al., 2002a). unfiltered k3-weighted experimental spectra. The Zn and Mn EXAFS experimental and reference Using the component spectra identified by PCA, linear spectra were processed and analyzed using SixPack soft- least-squares fitting of k-weighted experimental EXAFS ware (Webb, 2005) and a home-made software for data spectra [v(k)k3] with the reference spectra was conducted. normalization and Fourier transformation (Villalobos In the fits, the energy was not allowed to float and the com- et al., in press). The v(k) spectra were k-weighted and Fou- ponent sum was not forced to equal 1.0. The numerical re- rier transformed (FT) without smoothing in the k range of sults of these linear fits are interpreted as the fractional contribution of each reference spectrum to the experimental spectrum, and therefore, are taken to indicate the pro- Table 1 portion of each Zn species in the sample. In general, the Summary of Zn EXAFS reference material preparation precision of linear fitting is approximately ±10% (Manceau Zn acetate Zn phytate Zn citrate Zn dMnO2 et al., 2000b; Isaure et al., 2002). However, it can be as low Ligand (M) 2.0 Solubility 2.0 0.09 g L1b as approximately ±5% when (1) the spectrum from a spe- a limit cies of interest has high amplitude or a distinctive feature, Zn (M) 0.5 <0.01 0.5 7.05 · 105 pH 6 4 5 7 (2) the dataset is of high quality and well constrained c Complex1 10% Zn[ace] — 12% ZnH2[cit] — (ODay et al., 2004; Sarret et al., 2004; Bargar et al., Complex2 90% Zn[ace]2 — 12% Zn[cit]2 — 2005). The goodness of the linear fits was assessed by the Complex3 — 76% Zn[cit] — 2 minimization of the vm parameter (Newville, 2001). Note. The dMnO2 was provided by Dr. Villalobos (National University of Mexico), the synthesis recipe is published in Villalobos et al. (2003). 2.3. 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SIMULATIONS AND SURFACE COMPLEXATION MODELS

A Dissertation

Submitted to the Graduate School

of the University of Notre Dame

in Partial Fulfillment of the Requirements

for the Degree of

Doctor of Philosophy

Kelly J. Johnson

______Jeremy B. Fein, Director

Graduate Program in Civil Engineering and Geological Sciences

Notre Dame, Indiana

July 2006 includes three related studies that apply modeling techniques to describe the mechanisms and quantify the adsorption of heavy metals to bacterial cell walls.

In Chapter 2 we apply molecular simulations techniques to model heavy metal adsorption to the bacterial cell wall. In recent years, the adsorption of heavy metal cations onto bacterial surfaces has been extensively studied using both laboratory and field techniques. Metal adsorption has generally been modeled as a bulk partitioning process, ignoring the specific site interactions and only determining the quantity of metal adsorbed. Bulk adsorption measurements, involving both protons and aqueous metal cations, conducted as a function of pH and/or metal-bacteria concentration ratio, can be used to indirectly constrain important adsorption reactions and to determine the equilibrium constants for those reactions (e.g., Fein et al., 1997; Cox et al., 1999;

Martinez et al., 2001; Ngwenya et al., 2003; Borrok et al., 2004). Additional direct constraints on the metal-bacterial cell wall binding environment have been offered by X- ray absorption fine-structure spectroscopy (XAFS) investigations (e.g., Sarret et al.,

1998; Kelly et al., 2001; Boyanov et al. 2003a; Templeton et al., 2003; Francis et al.

2004). X-ray absorption spectroscopy can provide excellent constraints on the first and second nearest neighbors to a metal of interest on the cell wall. However, this approach only yields an averaged view of the binding environment consisting of numerous ligands and binding orientations. Molecular simulation methods have the potential to be a complementary third approach for studying metal-bacteria adsorption reactions, providing a more detailed and atomistic understanding of how metal cations interact with specific functional group types within the bacterial cell wall. Chapter 2 describes a study in which we applied energy minimizations and molecular dynamics based simulations to

3 for those reactions (e.g., Fein et al., 1997; Cox et al., 1999; Martinez et al., 2001;

Ngwenya et al., 2003). More direct constraints on the metal-bacterial cell wall binding environment have been offered by X-ray absorption fine-structure spectroscopy (XAFS) investigations (e.g., Sarret et al., 1998; Kelly et al., 2001; Boyanov et al. 2003a;

Templeton et al., 2003; Francis et al. 2004). X-ray absorption spectroscopy can provide excellent constraints on the first and second nearest neighbors to a metal of interest on the cell wall. However, this approach only yields an averaged view of what may be a complex binding environment consisting of numerous ligands and binding orientations.

Molecular simulation methods have the potential to be a complementary third approach for studying metal-bacteria adsorption reactions, providing a more detailed and atomistic understanding of how metal cations interact with specific functional group types within the bacterial cell wall.

Molecular simulations have previously been applied to model components of the bacteria cell. However, these studies focused on the lipid bilayer of the cell wall, not the metal-binding macromolecules within the cell wall (Bandyopadhyay et al., 2001; Shelley et al., 2001a,b). For example, Shelley et al. (2001b) simulated the self-assembly of the lipid bilayer starting with a random configuration, providing insight into two different phospholipid phases. The lipid bilayer research utilized coarse-grained simulation models to demonstrate the properties of the layer, grouping individual atoms with similar functionality into one entity. This scaling process enables modeling of relatively large systems and has reasonably low computational cost. However, because of the simplifications, the approach describes only bulk processes as opposed to atomic level interactions.

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108 Geochimica et Cosmochimica Acta 70 (2006) 5253–5269 www.elsevier.com/locate/gca

Cd adsorption onto Anoxybacillus ﬂavithermus: Surface complexation modeling and spectroscopic investigations

Peta-Gaye G. Burnett a,*, Christopher J. Daughney b, Derek Peak a

a Department of Soil Science, University of Saskatchewan, 51 Campus Drive, Saskatoon, Sask., Canada S7N 5A8 b Institute of Geological and Nuclear Sciences, P.O. Box 30368, Lower Hutt, New Zealand

Received 1 February 2006; accepted in revised form 3 August 2006

Abstract

Several recent studies have applied surface complexation theory to model metal adsorption behaviour onto mesophilic bacteria. How- ever, no investigations have used this approach to characterise metal adsorption by thermophilic bacteria. In this study, we perform batch adsorption experiments to quantify cadmium adsorption onto the thermophile Anoxybacillus flavithermus. Surface complexation models (incorporating the Donnan electrostatic model) are developed to determine stability constants corresponding to specific adsorption reactions. Adsorption reactions and stoichiometries are constrained using spectroscopic techniques (XANES, EXAFS, and ATR– FTIR). The results indicate that the Cd adsorption behaviour of A. flavithermus is similar to that of other mesophilic bacteria. At high bacteria-to-Cd ratios, Cd adsorption occurs by formation of a 1:1 complex with deprotonated cell wall carboxyl functional groups. At lower bacteria-to-Cd ratios, a second adsorption mechanism occurs at pH > 7, which may correspond to the formation of a Cd–phosphoryl, CdOH–carboxyl, or CdOH–phosphoryl surface complex. X-ray absorption spectroscopic investigations confirm the formation of the 1:1 Cd–carboxyl surface complex, but due to the bacteria-to-Cd ratio used in these experiments, other complexation mechanism(s) could not be unequivocally resolved by the spectroscopic data. 2006 Elsevier Inc. All rights reserved.

1. Introduction face of interest (e.g., carboxylic, phosphoryl, hydroxyl, amino, etc.) (Beveridge et al., 1982; Fein et al., 1997; The adsorption of metals by bacteria can influence the Daughney et al., 1998; Fowle et al., 2000; Fein et al., speciation and mobility of metals in geologic and aquatic 2001; Yee and Fein, 2001). Most SCMs constructed for environments. Previous studies of metal–bacteria adsorp- metal biosorption are primarily based on macroscopic tion have utilised surface complexation theory to quantify data, and only a limited number of investigations have used the amount of metal adsorbed (e.g., Fein et al., 1997; spectroscopic techniques to identify the functional groups Daughney and Fein, 1998; Daughney et al., 1998; Fowle on microbial surfaces (Sarret et al., 1998; Kelly et al., and Fein, 1999, 2000; Fein et al., 2001; Daughney et al., 2001, 2002; Boyanov et al., 2003; Benning et al., 2004; 2001; Yee and Fein, 2001; Yee et al., 2004a,b). Surface Sheng et al., 2004; Wei et al., 2004; Yee et al., 2004a,b). complexation theory is a specific type of chemical equilib- Obtaining direct evidence of the reaction mechanisms in- rium approach that correlates experimental adsorption volved in metal biosorption and the chemical composition data with electrical double layer (EDL) theory and bonding of the metal–bacterial complexes formed is vital for valida- mechanisms to describe adsorption. Surface complexation tion of SCMs. models (SCMs) explicitly describe the chemical reactions Despite the importance of metal–microbe interactions, that occur between the solute and specific sites on the sur- no extensive metal adsorption investigations have been conducted on thermophilic bacteria. Thermophiles are * Corresponding author. Fax: +1 30 69 666 881. extremophilic microorganisms that have adapted to grow E-mail address: [email protected] (P.-G. G. Burnett). at elevated temperatures ranging from approximately 45

(pH 5.2), and 1:2 and 1:50 in the cadmium acetate solutions standard deviations of fits constructed for each scan (stan- (pH 7.3 and 5.7, respectively). All solution samples were dards) and for the averages of fewer scans (biomass sam- transferred into liquid cells. ples). These standard deviations provide a measure of The bacterial pellets from a set of eighteen independent- precision not accuracy. ly grown 400-mL cultures were combined to make three samples (pHs 3.9, 5.8, and 7.0), while a different set of eigh- 2.6. Fourier transform infrared spectroscopy analyses teen cultures were combined to make three other samples (pHs 5.2, 7.6, and 8.5). Each sample (400 mL) was exposed 2.6.1. Standards and bacterial samples 1 2 3 to 5 mg L Cd and the pH adjusted to the desired value Solutions (50 mM) of H2PO4 , HPO4 , and PO4 using 1 M NaOH. After agitation for 2 h, samples were were prepared by dissolving the appropriate sodium phos- centrifuged (6079g, 15 min) and the supernatant decanted. phate salt in 18.2 MX deionized H2O (Barnstead Nano- The bacterial residues were filtered (0.45 lm) during centri- Pure, USA). Solutions of H3PO4 (50 mM) with and fugation (3522g, 30 min) two more times to ensure maxi- without 1 M Cd(NO3)2 were also prepared and both sam- mum removal of entrained solution. The final ples acidified with 0.75 mL concentrated HNO3. Multiple concentration of dissolved Cd was determined by FAAS 50 mM lecithin:Cd samples were prepared with molar (Varian SpectrAA 220, Victoria, Australia) as described ratios 1:0 and 1:2. A range of pH values (2.6–10.5) were previously. Bacterial pellets were loaded directly into a slot- obtained by addition of 1 M HNO3 or 1 M NaOH. Note ted sample holder, sealed with kapton tape, and cooled to that lecithin (R1R2HPO4) is a mixture of phospholipids 77 K using liquid N2. This experimental method is analo- and that R1 and R2 refer to (CH2)2N(CH3)3 and gous to that employed by Kelly et al. (2001), Kelly et al. CH2CHR3CH2R3, respectively, where the latter bears (2002), and Boyanov et al. (2003) for EXAFS on microbial hydrocarbon chains (R3) of varying lengths. We estimated systems. Note that the Cd loading was too low to obtain a a molecular weight of 750 U assuming the main R3-group reasonable signal from the pH 3.9 sample and so it will not is linoleic acid. be presented in this paper. Additionally, the pH 7.6 sample For bacterial sorption samples, the pellets from six inde- was excluded because of poor data quality (unknown pendently grown 400-mL cultures were combined to pro- cause). From this point forward, the pH 5.2 and 5.8 sam- vide sufficient biomass for the experiments and to ples will be referred to as the pH 5 and 6 samples, eliminate interculture variability. After the final rinse step, respectively. washed bacteria were resuspended in a known volume of electrolyte and then divided among two reaction vessels 2.5.2. Data collection and analysis of 400 mL each. One of the samples was exposed to XANES and EXAFS spectra were collected at the Ad- 0.5 mM Cd and both samples adjusted to the desired pH. vanced Photon Source at Argonne National Laboratory Samples were centrifuged (6079g, 15 min) after agitation with the electron storage ring operating at 4.5 GeV and for 2 h. The supernatant was decanted and analysed for in top up mode (100 ± 0.5 mA) for all experiments. Mea- dissolved Cd content via FAAS as described previously. surements at the Cd K-edge (26.711 keV) were made on the PNC-CAT 20-BM beamline using a Si(111) reflection 2.6.2. Data collection and analysis of the double crystal monochromator and with no focusing FTIR experiments were performed on a Bruker Equinox optics. Standards were analysed at room temperature in 55 Infrared Spectrometer (Ettlingen, Germany) equipped transmission mode. Energy calibration for solution stan- with a purge gas generator and a mercury cadmium tellu- dards was carried out by measuring the absorption of ride (MCT) detector. Liquid samples (1 mL) and wet bac- Cd3(PO4)2 simultaneously with that of the sample. Biomass terial pastes were loaded directly into a multiple bounce samples were placed in a cryostat and analysed at 77 K horizontal attenuated total reflectance (ATR) cell (Pike using a multi-element detector cooled with liquid N2. Ener- Technologies, USA) equipped with a 45 degree germanium gy was calibrated using a reference Ag foil for solid stan- (Ge) ATR crystal. Spectra were collected from 4000 to dards and biomass samples. In all cases, multiple scans 700 cm1 using 4 cm1 resolution and 1024 co-added scans were collected to improve signal to noise ratio. Data anal- per sample. The ATR crystal was thoroughly cleaned with ysis was carried out using WINXAS 3.1. Prior to process- distilled H2O and 95% ethanol between samples. The sam- ing the data, multiple scans were averaged. Calibration and ple chamber was continuously purged using H2O and CO2- background subtraction using standard procedures were free air to improve data quality. All data were interpreted always executed. Radial structure functions (RSFs) were using Opus 4.2. Background subtraction, baseline correc- generated by Fourier Transformation of the normalised, tion, and normalisation were performed on all spectra. weighted k3v(k) data over an approximate k range of For the baseline correction, a linear baseline was subtract- 2–9.5 A˚ 1. The EXAFS oscillations were quantitatively ed over the range 4000–700 cm1. Normalisation was analysed by curve fitting the data in R-space (out to 4.5 performed on baseline-corrected spectra over the range and 3.5 A˚ for standards and biomass samples, respectively) 1250–850 cm1 and 1800–700 cm1 for standards and bac- to theoretical scattering paths using FEFF 7.0. Uncertain- terial samples, respectively. No further smoothing or ties (2r) in each parameter were estimated by taking the manipulation was performed. Cd adsorption by Anoxybacillus flavithermus 5259

R–COO þ Cd2þ () R–COOCdþ ð6Þ 95% confidence limits, the regression line is not significantly different from the y = x line. Hence, based on principal R–COO þ CdOHþ () R–COOCdOH0: ð7Þ axes analysis, the SCM is suitably accurate in view of our In no instance was a three-site or three-reaction SCM re- experimental uncertainties. Moreover, the SCM is able to quired to satisfactorily account for the observed adsorption match all of the experimental data to within approximately (results not tabulated). This SCM yielded stability con- ±10%, covering a wide range of conditions from pH 3.6 to 1 stants of 2.38 ± 0.62 and 5.36 ± 0.50 for the Cd–carboxyl 9.0 and biomass concentration from 0.07 to 1.09 dry g L . and CdOH–carboxyl complexes, respectively. However, To examine whether the bacterial cells were affected by we stress that the other two-reaction SCMs fit most of Cd, we measured DOC concentrations in filtrates after the individual trials almost as well, and as such, identifica- the bacterial suspensions were equilibrated for 2 h with tion of the second adsorption reaction mechanism Cd at various pHs (low, mid, and high) and ionic strengths (formation of a Cd–phosphoryl, CdOH–carboxyl, or (0.001, 0.01, and 0.1 M). Based on DOC analyses conduct- CdOH–phosphoryl complex) is not possible based on the ed on bacterial suspensions equilibrated at different pHs batch adsorption data alone. Hence, spectroscopic mea- and ionic strengths during 2 h titrations (Burnett et al., 1 surements are ultimately required to confirm the SCM 2006), the addition of 5 mg L Cd had no significant effect reactions we have proposed (see below). on DOC concentrations after 2 h of equilibration, regard- We graphically and statistically compared the SCM fit less of pH or ionic strength (p > 0.05). Confocal laser scan- (above one-site two reaction SCM with overall variance ning microscopy examination indicates that most of the of 10.2) to all of our experimental data using a bivariate bacterial cells remain intact after equilibration with Cd. scattergram (Fig. 2). Principal axes analyses were conduct- We assume that intact, non-viable cells will interact with ed using formulations described in Sokal and Rohlf (1995). ions identically to cells that are intact and viable, but met- The unbiased values for the slope (first eigenvalue of the abolically inactive. Thus, we conclude that the presence of variance–covariance matrix) and intercept of the principal Cd does not affect the integrity of the cells of A. flavither- axis of the relationship between the experimental data mus under the conditions of our experiments. and the SCM predictions are 1.03 (0.99,1.08) and 2.36 (4.23,0.56), respectively. (Values in the parentheses cor- 3.2. XAS analyses respond to the 95% confidence intervals. The uncertainties in the slope and intercept were estimated using eigenvalues A small number of spectroscopic studies have been con- of the principal and secondary axes.) The degree of disso- ducted to identify the functional groups on biological sur- ciation between the observed and predicted values was faces and those that are responsible for metal binding. Wei evaluated using the product-moment correlation coefficient et al. (2004) found using ATR–FTIR that carboxylic, phos- (r) which was estimated at 0.96. The slope of this regression phate, and carbohydrate groups of both Gram-positive and line is not significantly different from 1 and also overlaps Gram-negative bacteria were sensitive to changes in solu- (95% confidence interval) with its ideal value. Although tion pH. Yee et al. (2004a,b) observed using synchrotron the ideal value for the intercept (zero) lies outside the radiation FTIR that the cell surface functional groups on the cyanobacteria Calothrix are a combination of proteins, lipids, and carbohydrates. These authors also noted an increase in peak intensity at 1400 cm1 over the pH range 100 3.2–6.5 during sequential titration of the cyanobacteria V(Y) = 10.2 and concluded that these vibrational changes corresponded 80 to the formation of deprotonated carboxyl sites. According to Kelly et al. (2001), uranyl binds to phosphoryl groups, whereas Cd binds to carboxyl functional groups on the sur- 60 face of Bacillus subtilis at low pH. Uranyl has also been reported to interact with phosphoryl and carboxyl func- 40 tional groups via inner-sphere complexation on the surface n = 141 of B. subtilis (Kelly et al., 2002). Sarret et al. (1998) found 7 independent cultures

20 via EXAFS that Zn binds predominantly to phosphoryl % Cd Adsorbed, Predicted Adsorbed, Cd % ca. 5 mg/L Cd pH 3.6- 9.0 (95%) and minor carboxyl groups (5%) on Penicillium 0.07 - 1.09 g/L chrysogenum, but that Pb had a reversed affinity for the 0 0 20 40 60 80 100 same surface. A study by Toner et al. (2005) also attributed % Cd Adsorbed, Observed Zn sorption by a bacterial biofilm to Zn–phosphoryl complexes, with a smaller contribution from carboxyl-type Fig. 2. Observed versus predicted extent of metal sorption, with predic- complexes. Independent investigations using X-ray Photo- tions based on adsorption reactions given in Eqs. (6) and (7). V(Y) is the variance between the SCM and the experimental data. Solid line is the electron Spectroscopy and FTIR revealed the chelating linear regression line based on principal axes analysis and the dashed lines character of the ion coordination to carboxyl groups and are the 95% confidence limits. confirmed that carboxyl, ether, hydroxyl, and amino 5268 P.-G. G. Burnett et al. 70 (2006) 5253–5269

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John Wiley and Sons, Inc., New York, pp. 1–16. 6 Mapping Biopolymer Distributions in Microbial Communities

John R. Lawrence, Adam P. Hitchcock, Gary G. Leppard, and Thomas R. Neu

CONTENTS

6.1 Introduction ...... 122 6.2 Methodology ...... 123 6.2.1 Handling Flocs for Microscopic Examination ...... 123 6.2.2 Epiﬂuorescence Microscopy...... 124 6.2.3 CLSM and 2P-LSM ...... 124 6.2.3.1 CLSM Limitations...... 125 6.2.3.2 2P-LSM Limitations...... 125 6.2.4 Synchrotron Radiation (Soft x-ray Imaging)...... 126 6.2.4.1 STXM Limitations...... 127 6.3 Targets and Probes ...... 127 6.3.1 Polysaccharides ...... 127 6.3.1.1 General Probes...... 127 6.3.1.2 Lectins...... 128 6.3.1.3 Antibodies...... 131 6.3.2 Proteins–Lipids ...... 131 6.3.3 Nucleic Acids...... 131 6.3.4 Charge/Hydrophobicity ...... 132 6.3.5 Permeability ...... 133 6.4 Examination of EPS Bound and Associated Molecules ...... 133 6.5 Digital Image Analyses...... 134 6.5.1 Quantitative In Situ Lectin Analyses ...... 135 6.6 Deconvolution...... 136 6.7 3D rendering ...... 136 6.8 Conclusions ...... 137 Acknowledgments ...... 137 References ...... 137

“L1615_C006” — 2004/9/9 — 16:56 — page 121 — #1 Mapping Biopolymer Distributions in Microbial Communities 127 x-ray analyses also have potential for application to bioﬁlm–ﬂoc materials having been used for bacterial cell–metal interaction studies.33

6.2.4.1 STXM Limitations Limitations to STXM include: suitability of the model compounds relative to bioﬁlm/ﬂoc material, data acquisition without undue radiation damage, requirement for very thin samples (<200 nm equivalent thickness of dry organic components, less than 5 micron of water when wet), use of fragile silicon nitride windows, sample preparation, that is, encapsulation in a wet cell, and absorption saturation distortion of analysis in thick regions of a specimen.

6.3 TARGETS AND PROBES The in situ analyses of hydrated biofilms may be carried out using a variety of probes targeted generally at polysaccharides, proteins, lipids, or nucleic acids. In addition, other probes such as dextrans, ficols, and polystyrene beads may be used to assess general properties such as charge, hydrophobicity, permeability, or the determination of diffusion coefficients. Probes are most frequently conjugated to fluors although colloidal reflective conjugates (gold, silver) may be used.27 Recently, quantum dots (QDs) have shown great promise as multiwavelength fluorescent labels. Colloidal QDs are semiconductor nanocrystals whose photoluminescence emission wavelength is proportional to the size of the crystal. Kloepfer et al.34 reported that cell surface molecules, such as glycoproteins, made excellent targets for QDs conjugated to wheat germ agglutinin. This new class of fluorescent labels may open opportunities for in situ detection of matrix chemistry. As indicated above, the option exists for probe independent examination of major biopolymers and other constituents in hydrated biofilm and floc material providing a basis for detailed examination of these structures and ground truthing of the fluorescent and reflection based probe dependent approaches.

6.3.1 POLYSACCHARIDES 6.3.1.1 General Probes A range of stains with specificity for beta-d-glucan polysaccharides are used as general stains, these include calcofluor white and congo red. Ruthenium red has also been used as a light microscopy stain for detection of EPS. Probes for glycoaminoglycan such as Alcian blue may also be used as a general stain for “polysaccharides.” Wetzel et al.35 demonstrated its use for determination of total EPS in microbial biofilms, in this case it was used indirectly and not for microscopy. Due to the complexity of the EPS the likelihood of finding a true total polysaccharide probe appears to be very limited.

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33. Kelly, S.D., Boyanov, M.I., Bunker, B.A., Fein, J.B., Fowle, D.A., Yee, N., and Kemner, K.M. XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J Synchrotron Radiat 8: 946–948, 2001. 34. Kloepfer, J. A., Mielke, R. E., Wong, M. S., Nealson, K. H., Stucky, G., and Nadeau, J. L. Quantum dots as strain- and metabolism-specific microbiological labels. Appl Environ Microbiol 69: 4205–4213, 2003. 35. Wetzel. R.G., Ward, A.K., and Stock, M. Effect of natural dissolved organic matter on mucilaginous matrices of biofilm communities. Arch Hydrobiol 139: 289–299, 1997. 36. Sharon, N. and Lis, H. A century of lectin research (1888–1988). Trends Biochem Sci 12: 488–491, 1987. 37. Sharon, N. and Lis, H. Lectins as cell recognition molecules. Science 246, 227–234, 1989. 38. Elgavish, S. and Shaanan, B. Lectin-carbohydrate interactions: Different folds, common recognition principles. Trends Biochem Sci 22: 462–467, 1997. 39. Bog-Hansen, T.C. Lectins — Biology, Biochemistry, Clinical Biochemistry. Volume 1, 2, and 3. de Gruyter, Berlin 1981. 40. Doyle, R.J. and Slifkin, M. Lectin-Microorganism Interactions. Marcel Dekker, New York, 1994. 41. Weis, W.I. and Drickamer, K. Structural basis of lectin-carbohydrate recognition. Annu Revi Biochem 65: 441–473, 1996. 42. Brooks, S.A., Leathem, A.J.C., and Schuhmacher, U. Lectin Histochemistry. Bios Scientific Publishers, Oxford, U.K., 1997. 43. Amann, R. Who is out there? Microbial aspects of biodiversity. Syst Appl Microbiol 23:1–8, 2000. 44. Staudt, C., Horn, H., Hempel, D.C., and Neu, T.R. Screening of lectins for stain- ing lectin-specific glycoconjugates in the EPS of biofilms. In P. Lens, A.P. Moran, T. Mahony, P. Stoodley, and V. O’Flaherty Biofilms in Medicine, Industry and Environmental Biotechnology. pp 308–327, Edited by IWA Publishing, UK, 2003. 45. Jones, A.H., Lee, C.-C., Moncla, B.J., Robinovitch, M.R., and Birdsell, D.C. Surface localization of sialic acid on Actinomyces viscosus. J Gen Microbiol 132, 3381–3391, 1986. 46. Morioka, H., Tachibana, M., and Suganuma, A. Ultrastructural localization of carbohydrates on thin sections of Staphylococcus aureus with silver methenamine and wheat germ agglutinin-gold complex. J Bacteriol 169: 1358–1362, 1987. 47. Merker, R.I. and Smit, J. Characterization of the adhesive holdfast of marine and freshwater Caulobacters. Appl Environ Microbiol 54: 2078–2085, 1988. 48. Hood, M.A. and Schmidt, J.M. The examination of Seliberia stellata exopolymers using lectin assays. Microb Ecol 31: 281–290, 1996. 49. Michael, T. and Smith, C.M. Lectins probe molecular films in biofouling: Charac- terization of early films on non-living and living surfaces. Mar Ecol Prog Ser 119: 229–236, 1995. 50. Liss, S.N., Droppo, I.G., Flannigan, D.T., and Leppard, G.G. Floc architecture in wastewater and natural riverine systems. Environ Sci Technol 30: 680–686, 1996. 51. Mohamed, M.N., Lawrence, J.R., and Robarts, R.D. Phosphorus limitation of heterotrophic biofilms from the Fraser River, British Columbia, and the effect of pulp mill effluent. Microb Ecol 36: 121–130, 1998. 52. Cummings, R.D. Use of lectins in analysis of glycoconjugates. Meth Enzymol 230: 66–86, 1994.

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Gordon E. Brown, Jr.1,2*, Jeffrey G. Catalano2, Alexis S. Templeton2,3, Thomas P. Trainor2,4, Francois Farges2,5, Benjamin C. Bostick2,6, Tom Kendelewicz2, Colin S. Doyle2, Alfred M. Spormann7, Kristin Revill2, Guillaume Morin8, Farid Juillot8 and Georges Calas8

1Stanford Synchrotron Radiation Laboratory, SLAC, MS 99, 2575 Sand Hill Road, Menlo Park, CA 94025, USA 2Dept. of Geological & Environmental Sciences, Stanford University, Stanford, CA 94305-2115, USA 3Marine Biology Research Division, Scripps Institution of Oceanography, University of California, San Diego, La Jolla, CA 92093, USA 4Consortium for Advanced Radiation Sources, University of Chicago, 5640 South Ellis Avenue, Chicago, IL 60637, USA 5 Laboratoire des géomatériaux, Université de Marne-la-Vallée, FRE CNRS 2455 “G2I”, 77454 Marne la Vallée cedex 2, France 6Dept. of Earth Sciences, Dartmouth College, 6105 Fairchild Hall, Hanover, NH 03755, USA 7Dept. of Civil and Environmental Engineering, Stanford University, Stanford, CA 94305, USA 8Laboratoire de Minéralogie-Cristallographie, Universités de Paris 6 et 7, IPGP and UMR CNRS 7590, 2 place Jussieu, 75252 Paris cedex 05, France

Received June 26, 2003; accepted November 4, 2003 pacs number: 61.10.Ht

Abstract (STXM), transmission X-ray microscopy (TXM)) to complex During the past decade, applications of synchrotron radiation (SR) methods environmental samples has led to the development of a new to environmental science have grown significantly, resulting in a new field referred to as molecular environmental science (MES) interdisciplinary field known as molecular environmental science, which focuses [5–7]. MES is concerned with determining the speciation on the molecular-level speciation of environmental pollutants and the fundamental of environmental pollutants and with understanding the chemical and biological processes determining their behavior in complex chemical and biological processes that affect their speciation, systems. Here we review some recent applications of XAFS spectroscopy and related SR methods, including micro-XANES and X-ray Standing Wave transformation, sequestration, transport, toxicity, and potential (XSW) spectroscopy, micro-XRD, X-ray fluorescence micro-tomography, and bioavailability. The term speciation, as used here, refers to (1) photoemission spectroscopy to heavy metals and actinides in model systems the identity of the element (or elements), (2) its (their) oxidation and more complex environmental samples. Examples include XAFS studies of state, (3) its (their) physical state (i.e., phase association; presence heavy metals and actinides at environmental interfaces, combined XAFS and in a liquid, gas, or solid phase (amorphous or crystalline), XSW studies of heavy metal interactions with microbial biofilm-coated metal oxide surfaces, and combined XANES and photoemission studies of microbial colloidal particle, animal or plant cell, or biofilm; presence as interactions with metal sulfide surfaces. a surface coating or thin film on a solid, as a sorption complex (monomeric or polymeric) on a solid, colloidal particle, or an organic substance; etc.), (4) its stoichiometry, and (5) its detailed 1. Introduction molecular structure. Knowledge of the molecular-level speciation Since X-ray absorption fine structure (XAFS) spectroscopy of pollutants is essential for predicting their environmental was introduced over 30 years ago as an element-specific properties, i.e., their stability, mobility, toxicity, and potential local structural probe [1, 2], it has become the technique of bioavailability to humans and other organisms [5, 8], as well as choice for determining the local coordination environment of for designing effective remediation strategies (e.g., [9, 10]). specific elements in solids (both amorphous and crystalline) Because of the complexity of environmental samples, which and liquids at low concentrations, where X-ray and neutron often contain dissolved and solid phases, including amorphous diffraction methods are relatively insensitive. Because of materials (natural organic matter), as well as biological this attribute and the ability to examine samples under components (plant matter, fungi, microorganisms), it is essential in situ conditions with minimal preparation, XAFS spectroscopy to conduct SR-based studies of simplified model systems in has become the most important method for determining the parallel with studies of natural samples in order to gain a speciation of environmental pollutants such as zinc, lead, and more fundamental understanding of the chemical and biological arsenic, which are often present in complex, multi-phase, processes that control the behavior and toxicity of environmental natural samples at concentrations ranging from greater than a pollutants. This need has led to numerous studies of adsorption weight percent to less than a ppm in a variety of chemical processes at environmental interfaces, including interfaces forms, including those in which such elements are sorbed at between minerals and microbial biofilm coatings, as well as / / solid/aqueous solution interfaces [3, 4]. Applications of XAFS plant-root water and microbe water interfaces [4]. To date, and micro-XAFS spectroscopy and other SR methods (micro- XAFS studies of environmental samples have focused mainly synchrotron X-ray fluorescence (-SXRF), X-ray standing on inorganic pollutants, particularly heavy metals such as Cr, wave (XSW) spectroscopy, photoemission spectroscopy (PES), Ni, Zn, Cd, Hg, and Pb, metalloids such as As and Se, and micro-X-ray diffraction (-XRD), X-ray fluorescence and actinides such as U, Pu, and Np. Such elements have micro-tomography, scanning transmission X-ray microscopy relatively high K- or L-edge energies, thus can be studied at low concentrations (currently ≤50ppm with insertion device beam lines on 3rd-generation SR sources) even in strongly absorbing *e-mail: [email protected] matrices, assuming that spectral interferences are minimal

Physica Scripta T115 C Physica Scripta 2005 82 Gordon E. Brown, Jr. et al. the increasing number of coordinatively unsaturated surface Ti of northern Idaho, U.S.A. Mn and Zn occur as isolated nodules of sites with decreasing nanoparticle size. In a similar EXAFS study mixed-metal carbonate (rhodochrosite/hydrozincite) on the root of nanoparticulate -Fe2O3, surface Fe sites were found to be surface. In most cases, the molecular-scale mechanisms of metal undercoordinated relative to Fe in the bulk structure and are transformations or hyperaccumulation by bacteria or plants are restructured to octahedral sites when the nanoparticles are reacted unknown. with enediol ligands [41]. Nanoparticles play an important role in the transport of heavy metal and organic contaminants in the environment. For example, 3. XAFS Studies of Metal Ion Adsorption at Solid/Aqueous Kersting et al. [42] concluded that colloid-facilitated transport of Solution Interfaces Pu is responsible for its subsurface movement from the Nevada Most chemical reactions involving environmental materials take Test Site in the United States to a site 1.3km south over a relatively place at interfaces, most importantly solid/aqueous solution short time period (≈40 years), based on the unique 240Pu/239Pu interfaces [6, 25]. Although not inherently surface (or interface) isotope ratios of the samples. EXAFS spectroscopy and -SXRF sensitive, when the element of interest occurs dominantly at a are now shedding light on how Pu and other actinide ions sorb on surface (or interface), XAFS spectroscopy is often capable of colloidal mineral particles and undergo redox transformations in providing unique information about the structure and composition some cases [43, 44]. of an adsorbate, as well as its mode of attachment to a solid surface (i.e., inner-sphere vs. outer-sphere, mononuclear vs. multinuclear, 2.5. Inorganic and organic ligands and coatings can have a and (if inner-sphere) monodentate vs. bidentate or tridentate). major impact on the sorption behavior of metal ions Because of this capability, XAFS studies have now been carried There are a number of common natural inorganic ligands (e.g., out on over 200 metal ion/mineral sorption systems (see [4] for a 2− 3− 2− − SO4 ,PO4 ,CO3 ,Cl ) and organic ligands (e.g., oxalate, recent review). Here we review the results of two such studies, one citrate, humic materials) that can enhance or inhibit the sorption on Zn(II) sorption on alumina [67] and one on Au(III) sorption of metal ions from aqueous solutions onto mineral surfaces. For on Fe(III)- and Al-oxyhydroxides [68, 69]. example, sulfate enhances the sorption of Pb(II) on goethite The products of sorption reactions of Zn(II) with high surfaces over the pH range 5 to 7 due to the reduction of surface area alumina (Linde-A) were characterized by XAFS the positive surface charge of goethite at these pH values spectroscopy as a function of Zn sorption density ( = 0.2 to 2− / 2 and the formation of a stable ternary Pb(II)/SO4 /goethite 3.3 mole m ) at pH values of 7.0 to 8.2 by Trainor et al. surface complex [45]. In contrast, chloride ligands inhibit [67] (see Fig. 1). Over equilibration times of 15–111hrs and the sorption of Hg(II) on goethite at similar pH values at low Zn sorption densities ( = 0.2 to 1.1mole/m2), we because of the formation of a stable Hg(II)-chloride aqueous found that Zn(II) forms dominantly inner-sphere bidentate surface complex [46]. Organic ligands can have similar effects. For complexes with AlO6 polyhedra. At higher sorption densities example, citrate [47] and malonic acid [48] can enhance the sorption of U(VI) and Pb(II), respectively, onto goethite through the formation of ternary surface complexes, as revealed by EXAFS and FTIR spectroscopy. Combined in situ XAFS and ATR-FTIR spectroscopy studies are beginning to reveal some of the molecular-level details of ternary surface complexes involving metal ions, organic (or inorganic) molecules, and mineral surfaces (e.g., [49–51]).

2.6. Biota can have a major effect on the speciation and sequestration of contaminant and pollutant species Many microorganisms and plants can sequester toxic chemical species and transform them into less toxic forms. EXAFS spectroscopy studies of microorganism-metal interactions are beginning to reveal the types of cell-wall functional groups to which metal ions bind [52–61]. For example, in an EXAFS study of the sorption products of Cd(II) and U(VI) on Bacillus subtilis, Kelly et al. [58] found that U(VI) binds to phosphoryl groups and that Cd(II) binds to carboxyl groups at low pH. The transformations and sequestration of metal ions in hyperaccumulating plants has also been studied in several systems using XAFS and -XAFS spectroscopy (e.g., [62–66]) and X-ray fluorescence -tomography [66]. For example, several groups have found that aqueous Cr(VI) is accumulated in and rapidly reduced to Cr(III) in aquatic ferns [62, 63]. In one of the first X-ray fluorescence -tomography studies (coupled with Fig. 1. Zn K-edge EXAFS spectra (k3 weighted) (A) and corresponding Fourier XAFS), Hansel et al. [66] showed that Pb and As are associated transforms (B) of Zn sorbed onto high surface area alumina at the following surface coverages ( = molesm−2): 0.2 (1a), 0.3 (1b), 0.8 (1c), 1.1 (1d), 1.7 (1e), 1.6 with iron plaque (in the form of ferrihydrite, goethite, and siderite) (1f), 1.5 (1g), and 3.3 (1h), compared with Zn-hydrotalcite (Zn-HTLC) and Zn2+ 2+ on plant roots of Phalaris arundinacea, a common aquatic plant in a 10mM Zn(NO3)2 aqueous solution (Zn (aq)) (after [67]). The structure and from a mine waste impacted wetland in the Coeur d’Alene Basin composition of Zn/Al-hydrotalcite (Zn/Al-HTLC) are shown.

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Table I. EXAFS fit results for Borehole 102 sample 61A compared 29. Brown, G. E., Jr. and Parks, G. A., Internat. Geol. Rev. 43, 963 (2001). with fit results for boltwoodite [(K,Na)2(UO2)2(SiO3OH)2 30. Fenter, P. A., Rev. Mineral. Geochem. 49, 149 (2002). 31. Fenter, P. A. et al., J. Colloid Interface Sci. 225, 154 (2000). ·3H2O)] and uranophane [Ca(UO2)2(SiO3OH)2·5H2O], two 32. Trainor, T. P. et al., Langmuir 18, 5782 (2002). members of the uranophane group (from [88]). 33. Banfield, J. F. and Zhang, H., Rev. Mineral. Geochem. 44, 1 (2001). 34. Gleiter, H., Adv. Materials (Weinheim, Germany) 4, 474 (1992). = Sample U Oax U-Oeq1 U-Oeq2 U-Si U-U 35. Siegel, R. W., Scientific American 275, 74 (1996). 36. McHale, J. M. et al., Science 277, 788 (1997). 61A 2–1.81Å 2.9–2.29Å 1.8–2.46Å 0.9–3.17Å 1.8–3.94Å 37. Chen, L. X. et al., J. Phys. Chem. B 101, 10688 (1997). Boltwoodite 2–1.81Å 3–2.30Å 2–2.45Å 1–3.15Å 2–3.95Å 38. Chen, L. X. et al., Nucl. Instrum. Meth. B 133, 8 (1997). Uranophane 2–1.81Å 3–2.28Å 2–2.46Å 1–3.16Å 2–3.93Å 39. Chen, L. X. et al., J. Synchrotron. Rad. 6, 445 (1999). 40. Rajh, T. et al., J. Phys. Chem. B 103, 3515 (1999). 41. Chen, L. X. et al., J. Phys. Chem. B 106, 8539 (2002). cells, microbial biofilms, natural organic matter, and plants. 42. Kersting, A. B. et al., Nature 397, 56 (1999). Micro-XAFS methods are contributing unique insights about 43. Kaplan, D. I. et al., Environ. Sci. Technol. 28, 1186 (1994). 44. Allen, P. G. et al., Abstracts of Papers, 222nd American Chem. Soc. Nat. the association of heavy metal and metalloid pollutants with Mtg., Chicago, IL, GEOC-042 (2001). environmental materials. With the increasing availability of high 45. Ostergren, J. D. et al., J. Colloid Interface Sci. 225, 483 (2000). brightness synchrotron X-ray sources, it is likely that information 46. Kim, C. S., Rytuba, J. J. and Brown, G. E., Jr., J. Colloid Interface Sci. (2003, on molecular-level speciation of environmental pollutants will be in press). increasingly utilized in making decisions about environmental 47. Redden, G. D., Bargar, J. R. and Bencheikh-Latmani, R., J. Colloid Interface Sci. 244, 211 (2001). remediation. 48. Lenhart, J. J., Bargar, J. R. and Davis, J. A., J. Colloid Interface Sci. 234, 448 (2001). 49. Bargar, J. R., Persson, P. and Brown, G. E., Jr., Geochim. Cosmochim. Acta Acknowledgments 63, 2957 (1999). 50. Bargar, J. R., Reitmeyer, R. and Davis, J. A., Environ. Sci. Technol. 33, 2481 We thank the U.S. National Science Foundation (CRAEMS Grant CHE-0089215, (1999). GEB), the U.S. Department of Energy (DOE-BES Grant DE-FG03-93ER14347- 51. Fitts, J. P. et al., J. Colloid Interface Sci. 220, 133(1999). A008, GEB), and the French CNRS (GC, GM, FJ, FF) for financial support of the 52. Templeton, A. S. et al., J. Synchrotron Rad. 6, 642 (1999). research reported here. We also thank the staffs of SSRL, APS, LURE, and ESRF, 53. Templeton, A. S. et al., Proc. Nat. Acad. Sci. U.S.A. 98, 11897 (2001). where the data used in our studies were collected. SSRL is supported by the U.S. 54. Hennig, C. et al., Radiochim. Acta 89, 625 (2001). Department of Energy (Office of Basic Energy Science) and the U.S. National 55. Kelly, S. D. et al., J. Synchrotron Rad. 8, 946 (2001). Institutes of Health. 56. Webb, S. M. et al., J. Synchrotron Rad. 8, 943 (2001). 57. Panak, P. J. et al., Radiochim. Acta 90, 315 (2002). 58. Kelly, S. D. et al., Geochim. Cosmochim. Acta 66, 3855 (2002). 59. Templeton, A. S. et al., Environ. Sci. Technol. 37, 300 (2003). References 60. Templeton, A. S., Spormann, A. M. and Brown, G. E., Jr., Environ. Sci. 1. Lytle, F. W., Adv. X-ray Anal. 9, 398 (1966). Technol. 37, 2166 (2003). 2. Sayers, D. E., Stern, E. A. and Lytle, F. W., Phys. Rev. Lett. 27, 1204 (1971). 61. Templeton, A. S. et al., Geochim. Cosmochim. Acta 67 (2003, in press). 3. Brown, G. E., Jr., Parks, G. A. and Chisholm-Brause, C. 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accumulate U phases that may be more representative of Uranium Complexes Formed at those that exist in U-contaminated aquifers than those Hematite Surfaces Colonized by which accumulate in closed experimental systems. These phases should be considered in models attempting to Sulfate-Reducing Bacteria predict U transport through subsurface environments.

ANDREW L. NEAL,†,‡,§ | JAMES E. AMONETTE, Introduction BRENT M. PEYTON,⊥ AND GILL G. GEESEY*,†,‡ There is now compelling evidence that microorganisms play Department of Microbiology, Montana State University, a significant role in the cycling of U in anaerobic environ- 3+ Bozeman, Montana 59717-3520, Center for Biofilm ments. Almost coincident with the identification of the Fe - 2- Engineering, Montana State University, and SO4 -reducing zones as sites for U(VI) reduction in Bozeman, Montana 59717-3980, Environmental Molecular sediments (1, 2), Fe(III)-reducing bacteria of the genera Sciences Laboratory, Pacific Northwest National Laboratory, Geobacter and Shewanella and SRB of the genus Desulfovibrio Richland, Washington 99352, and Center for Multiphase (inter alia) were identified as being responsible for biogenic Environmental Research, Washington State University, U(VI) f U(IV) reduction (3, 4). The reduction process by Pullman, Washington 99164 both groups of organisms is enzymatic (3, 5), with cytochromes, principally of the c type, implicated (6, 7). The affinity of U(VI) for microbial cell wall components (neutral -POH and deprotonated -COOH groups), and the Modeling uranium (U) transport in subsurface environments resulting increase in electron density U(VI) imparts to the requires a thorough knowledge of mechanisms likely to cell walls, has been exploited for many years to facilitate restrict its mobility, such as surface complexation, visualization of microbial cells by electron microscopy (8, 9). precipitation, and colloid formation. In closed systems, sulfate- Our knowledge of specific interactions between U(VI) and reducing bacteria (SRB) such as Desulfovibrio spp. bacterial cells comes a great deal from the pioneering work demonstrably affect U immobilization by enzymatic reduction of Lovley and colleagues, who identified uraninite [UO2(s)] 3+ of U(VI) species (primarily the uranyl ion, UO 2+, and its formation resulting from cytochrome activity of both Fe 2 and SO 2- reducers (3, 5, 10-12). Continued research on complexes) to U(IV). However, our understanding of such 4 U(VI)-bacterial interaction stems from its importance in interactions under chronic U(VI) exposure in dynamic defining the mobility and bioavailability of U in groundwater. systems is limited. As a first step to understanding such Geochemical models predicting transport of the various interactions, we performed bioreactor experiments under contaminants through the vadose and saturated zones are continuous flow to study the effect of a biofilm of the sulfate- based on the principle of mass conservation in the aqueous reducing bacterium Desulfovibrio desulfuricans attached phase along a flow path (13). One of the greatest limitations to specular hematite (R-Fe2O3) surfaces on surface-associated of the mass conservation approach is an inability to predict U(VI) complexation, transformation, and mobility. Employing the fate and transport of contaminants arising from the use real-time microscopic observation and X-ray photoelectron of linear, reversible sorption as the only mechanism covering spectroscopy (XPS), we show that the characteristics of the the interaction between aqueous and solid phases during transport. The use of empirical sorption coefficients in the U(VI) complex(es) formed at the hematite surface are evaluation of contaminant fate and transport overlooks other influenced by the composition of the bulk aqueous phase geochemical processes such as mineral precipitation, forma- flowing across the surface and by the presence of surface- tion of colloids, and oxidation/reduction reactions by associated SRB. The XPS data further suggest higher levels subsurface microorganisms. The potentially complex nature of U associated with hematite surfaces colonized by of the insoluble precipitates arising from various bacterially SRB than with bacteria-free surfaces. Microscopic driven redox reactions in groundwater results in serious observations indicate that at least a portion of the U(VI) limitations of available geochemical models to accurately that accumulates in the presence of the SRB is exterior to describe the behavior of U. To confound the problem, our the cells, possibly associated with the extracellular knowledge of the products of bacterial uranyl reduction/ biofilm matrix. The U4f core-region spectrum and U5f2 precipitation to date results from the study of closed systems, 7/2 that is to say, batch cultures with a single addition of U(VI). valence-band spectrum provide preliminary evidence that It is therefore imperative that we better understand the likely the SRB-colonized hematite surface accumulates both fate of U(VI) when bacteria are present under conditions of U(VI) and U(IV) phases, whereas only the U(VI) phase(s) chronic U exposure in flowing systems. accumulates on uncolonized hematite surfaces. The results As a first step to understanding U-bacterial dynamics in suggest that mineral surfaces exposed to a continuously flowing systems, we studied the effects of SRB biofilms replenished supply of U(VI)-containing aqueous phase will associated with the iron(III) oxide hematite (R-Fe2O3)onthe removal of U(VI) from solution under hydrodynamically * Corresponding author phone: (406) 994-3820; fax: (406) 994- relevant conditions. Hematite was chosen as a suitable 4926; e-mail: [email protected]. substratum as it is a component of many soils and sediments † Department of Microbiology, Montana State University. (14) and in its specular habit forms flat macroscopic faces ‡ Center for Biofilm Engineering, Montana State University. well suited to microscopic, spectroscopic, and flow studies. § Present address: Advanced Analytical Center for Environmental Moreover, we have previously described iron sulfide forma- Studies, Savannah River Ecology Laboratory, Drawer E, Aiken, SC 29802. tion by SRB under similar conditions (15). The goal of this | Pacific Northwest National Laboratory. work is to document the changes in U complexes formed at ⊥ Washington State University. a hematite surface as a result of surface colonization by a

10.1021/es030648m CCC: $27.50  2004 American Chemical Society VOL. 38, NO. 11, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 3019 Published on Web 04/17/2004 FIGURE 4. Low-resolution XP spectra of (top) the bacteria-free hematite surface following 30 days of exposure to flowing, sterile, UMA-containing, anaerobic culture medium and (bottom) the hematite surface following 30 days of exposure to cells of G20gfpmut-2 in the presence of flowing, UMA-containing, anaerobic culture medium.

C-O/C-N at 286.3 eV, and CdO/OsCsO at 288.2 eV (22, 48-50) (Figure 5a). The C1s region of the XP spectrum of the hematite surface colonized by cells of G20gfpmut-2 in the presence of anaerobic culture medium containing UMA shows evidence of these same features at approximately the same binding energies, although not as well resolved as in the absence of UMA (Figure 5b). The feature centered between binding energies of 288.9 and 289.5 eV, where contributions from carboxyl groups (OsCdO) are predicted (22), contributes less than 5% of the C1s intensity in either the presence or absence of uranyl. Although there appears to be a 0.7 eV shift to higher energy for the carboxyl peak in FIGURE 5. C1s core region of the XP spectrum of the hematite the presence of U (Figure 5a,b), this shift falls within the surface following (a) 30 days of colonization by cells of G20gfpmut-2 range of uncertainty of the peak-fitting routine. That the C1s in the presence of flowing, UMA-free anaerobic culture medium region was sensitive to the bacteria on the hematite surface and (b) 30 days of colonization by cells of G20gfpmut-2 in the presence is demonstrated by the increase in intensity of the C-O/ of flowing, UMA-containing, anaerobic culture medium and (c) C1s C-N peak relative to the C-C/C-H and CdO/OsCsO peaks core region of the XP spectrum of the sterile hematite surface in spectra of colonized surfaces compared to the relative following 30 days of exposure to flowing, sterile, UMA-free anaerobic intensities of these peaks in spectra of hematite surfaces culture medium. Component bands were fit to the C1s region exposed to sterile culture medium (Figure 5c). Thus, it is following baseline subtraction. The data (b), fitted model (s), and unlikely that the increased U complexation at the hematite component curves (---) are shown. Data presented across the top surface colonized by G20gfpmut-2 involved carboxyl groups are the deviations from the curve fitting. of biofilm components. Similarly, the lack of carbonate and bicarbonate contributions (289.6-291.5 eV) (51, 52)tothe abundance of U(VI) associated with G20gfpmut-2-colonized C1s spectrum of hematite surfaces colonized by bacteria hematite surfaces. Kelly et al. (54) found that phosphoryl indicates that inner-sphere adsorption of U(VI) to carbonate groups, rather than carboxyl groups, were responsible for at the hematite surface is not responsible for the increased uranyl binding to cell walls of Bacillus subtilis. Although U accumulation in the presence of bacteria (31, 53). further work is required to resolve the types of U(VI) Uranyl coordination by phosphoryl groups in the cell wall complexes formed at hematite surfaces colonized by cells of of G20gfpmut-2 may have contributed to the increased G20gfpmut-2, the results presented here suggest that the

3024 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 38, NO. 11, 2004 (46) Sutherland, I. W. In Surface carbohydrates of the procaryotic (54) Kelly, S. D.; Boyanov, M. I.; Bunker, B. A.; Fein, J. B.; Fowle, D. cell; Sutherland, I., Ed.; Academic Press: New York, 1977; pp A.; Yee, N.; Kemner, K. M. J. Synchrotron Radiat. 2001, 8, 946- 27-96. 948. (47) Allen, G. C.; Shuh, D. K.; Bucher, J. J. E., N. M.; Reich, T.; Denecke, (55) Suzuki, Y.; Kelly, S. D.; Kemner, K. M.; Banfield, J. F. Nature M. A.; Nitsche, H. Inorg. Chem. 1996, 35, 784-787. 2002, 419, 134. (48) Rouxhet, P. G.; Mozes, N.; Dengis, P. B.; Dufrene, Y. F.; Gerin, (56) Finch, R.; Murakami, T. In Uranium: mineralogy, geochemistry P. A.; Genet, M. J. Colloids Surf., B 1994, 2, 347-369. and the environment Burns, P. C., Finch, R., Eds.; Mineralogical Society of America: Washington, DC, 1999; Vol. 38, pp 91- (49) Beech, I. B.; Gubner, R.; Zinkevich, V.; Hanjagsit, L.; Avci, R. - 179. Biofouling 2000, 16,93 104. (57) Locock, A. J.; Burns, P. C. Am. Mineral. 2003, 88, 240-244. (50) Laszlo, K.; Josepovits, K.; Tombacz, E. Anal. Sci. 2001, 17 (suppl), (58) Pietzsch, K. B.; Hard, C.; Babel, W. J. Basic Microbiol. 1999, 39, i1741-i1744. 365-372. (51) Stipp, S. L.; Hochella, M. F. J. Geochim. Cosmochim. Acta 1991, - 55, 1723 1736. Received for review October 2, 2003. Revised manuscript - (52) Heuer, J. K.; Stubbins, J. F. Corrosion 1999, 41, 1231 1243. received March 5, 2004. Accepted March 15, 2004. (53) Barger, J. R.; Reitmeyer, R.; Davis, J. A. Environ. Sci. Technol. 1999, 33, 2481-2484. ES030648M

VOL. 38, NO. 11, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 3027 APPLIED AND ENVIRONMENTAL MICROBIOLOGY, Feb. 2004, p. 771–780 Vol. 70, No. 2 0099-2240/04/$08.00ϩ0 DOI: 10.1128/AEM.70.2.771–780.2004 Copyright © 2004, American Society for Microbiology. All Rights Reserved.

Determination of Cu Environments in the Cyanobacterium Anabaena ﬂos-aquae by X-Ray Absorption Spectroscopy X. C. Kretschmer,1* G. Meitzner,2 J. L. Gardea-Torresdey,1 and R. Webb1 Department of Environmental Science and Engineering1 and Materials Research and Technology Institute,2 The University of Texas at El Paso, El Paso, Texas 79968

Received 18 July 2003/Accepted 4 November 2003 Downloaded from Whole cells and peptidoglycan isolated from cell walls of the cyanobacterium Anabaena ﬂos-aquae were lyophilized and used at pH 2 and pH 5 in Cu(II) binding studies. X-ray absorption spectra measured at the Cu K-edge were used to determine the oxidation states and chemical environments of Cu species in the whole-cell and peptidoglycan samples. In the whole-cell samples, most of the Cu retained at both pH values was coordinated by phosphate ligands. The whole-cell fractions contained signiﬁcant concentrations of Cu(I) as well as Cu(II). An X-ray absorption near-edge spectrum analysis suggested that Cu(I) was coordinated by amine and thiol ligands. An analysis of the peptidoglycan fractions found that more Cu was adsorbed by the peptidoglycan

fraction prepared at pH 5, due to increased chelation by amine and carboxyl ligands. The peptidoglycan http://aem.asm.org/ fractions, also referred to as the cell wall fractions, contained little or no Cu(I). The Cu loading level was 30 times higher in the cell wall sample prepared at pH 5 than in the sample prepared at pH 2. Amine and bidentate carboxyl ligands had similar relative levels of importance in cell wall peptidoglycan samples prepared at both pH values, but phosphate coordination was insigniﬁcant.

Heavy metals damage the environment since they are non- Much research has been accomplished recently in the study of biodegradable and have toxic effects on plants, animals, and cyanobacterial protein interactions with metals, metal availabil- humans. Precipitation, coagulation, membrane-based pro- ity in cyanobacteria, and Cu homeostasis in other microorgan- on August 23, 2014 by UNIV OF NOTRE DAME cesses, and ion exchange are some of the current approaches to isms (12, 62, 69). We are concerned here with metals adsorbed the remediation of contaminated sites, but these methods are by lyophilized cells and not essential metals that are present, expensive and lose effectiveness as metal concentrations fall for example, as metalloenzyme cofactors in live cells (15, 17). (54). The use of microbes for the bioremediation of trace metals Several groups have used XAS to investigate the chemistry has shown potential for enabling a site to meet regulatory spec- of accumulated metals in biological samples. Previous XAS ifications. Experiments have demonstrated the ability of bacteria studies of Cu in plants have examined Solanum elaeagnifolium to immobilize and concentrate metal ions (6, 14, 28, 37, 46). (silverleaf nightshade) (68), hops (systematic name not given) Mechanisms by which cells accumulate, sequester, and de- (49), and Larrea tridentata (creosote bush) (22, 51). These toxify metals have been investigated. One indirect approach reports identified only the immediate oxygen, nitrogen, or sul- has been the determination of uptake and retention as a func- fur neighboring atoms and inferred hydroxyl, amine, carboxyl, tion of pH or competition with some other cation (14, 16, 20, or thiol ligands. Conclusions were corroborated by investiga- 35). Another has been to chemically modify plant or cell mat- tions of the effect of esterifying prospective ligands upon Cu ter (for example, by esterification) to render specific classes of binding. Decreased Cu binding was interpreted as evidence prospective ligands unavailable and to evaluate the impact on that a carboxyl was the principal ligand. However, it is not clear metal binding capacity (7, 21). In a study of binding sites in cell that the esterification chemistry was specific for carboxyl walls of gram-positive bacteria, teichoic acid, presenting phos- groups. An extended X-ray absorption fine structure (EXAFS) phate sites, was extracted with aqueous alkali in order to quan- study of Zn coordination in anaerobic bacteria concluded that tify its contribution to metal binding (7). XAS is capable of discriminating thiol, O/N, and phosphate The chemical studies described are informative but involve ligands (72), but did not quantify the respective fractions co- many assumptions due to the complexity of biological systems. ordinating Zn in the samples. A study by Sarret et al. (60) of Also, because of that complexity, few analytical probes can Zn(II) and Pb(II) binding by cell walls from the fungus Peni- illuminate the chemistry of metals associated with cells. X-ray cillium chrysogenum combined titration with the metal cations absorption spectroscopy (XAS) is element specific, returning of binding sites and EXAFS of adsorbed cations. Carboxyl-Pb information for a single atomic number in an arbitrarily com- complexes were formed at low Pb concentrations, followed by 3Ϫ plicated mixture (52). XAS is nondestructive, and no sample (PO4)n-Pb complexes. In contrast, Zn saturated PO4 sites reduction or digestion is required which would alter the chem- first. The study developed a detailed model for metal binding istry of the element of interest. XAS can be used to determine sites. Nevertheless, most XAS studies of cell-bound Cu and the oxidation state or states of the element as well as the phys- other metals have presented a simplified picture, approximat- ical structure of sites surrounding it to a radius of about 5 A˚ . ing the metal environment by the presumably most abundant ligand and assuming the extended structure of this ligand. In * Corresponding author. Mailing address: Department of Biological this report, we describe the use of a more elaborate model to Sciences, The University of Texas at El Paso, El Paso, TX 79968. Phone: interpret the EXAFS of Cu bound by purified peptidoglycan of (915) 747-6876. Fax: (915) 747-5808. E-mail: [email protected]. cell walls and to differentiate and quantify amine and carboxyl

771 VOL. 70, 2004 Cu ENVIRONMENTS IN ANABAENA FLOS-AQUAE 779

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Effects of cultivation conditions on acid–base titration properties of Shewanella putrefaciens

Johnson R. Haas*

Department of Geosciences/Environmental Studies Program, Western Michigan University, Kalamazoo, MI 49008, USA

Received 19 June 2003; accepted 19 April 2004 Available online

Abstract

Acid–base titrations were conducted on pure strain laboratory cultures of the Gram-negative, facultatively aerobic bacterium Shewanella putrefaciens, cultivated under a range of conditions. Bacteria used in acid–base titrations were cultivated for periods of 24 to 100 h (exponential growth phase to late stationary phase), in varying media compositions (rich to minimal), and under aerobic and anaerobic conditions. In addition, two mutant strains of S. putrefaciens were tested, which are respiration- deficient with respect to U(VI) and Fe(III). Titration data were used to constrain the abundances and proton-exchange properties of bacterial surface functional groups using a discrete multisite surface complexation model. Estimated properties for titratable functional groups at the bacteria–water interface were compared on the basis of cultivation conditions. Results demonstrate that the acid–base behavior of the S. putrefaciens surface is largely insensitive to variations in culture age and media composition, yielding values for a wide range of aerobic cultivation conditions that are not significantly different on a statistical basis. Conversely, results for anaerobic growth, in which Fe(III) was supplied as sole terminal electron acceptor, show significantly lower bacterial site densities and differing functional-group acid dissociation constants. It is recommended that different profiles for aerobic and anaerobic Gram-negative bacteria be used in aqueous speciation calculations involving a bacterial component. D 2004 Elsevier B.V. All rights reserved.

Keywords: Bacteria; Titration; Shewanella putrefaciens; Surface-complexation; Anaerobic; Growth phase

1. Introduction 2002) or may be chemically modified through direct or indirect microbial processes, such as reduction Bacteria in surface aquatic environments and in (e.g. Lloyd and Lovley, 2001; Lloyd et al., 2002; the subsurface influence the chemical speciation of Nealson et al., 2002; Nevin and Lovley, 2002; Luu solutes in a variety of ways, not least of which and Ramsay, 2003) and oxidation (Olson, 1991; include the coordination of dissolved chemical spe- Sugio et al., 1992; Nordstrom, 2000; Konhauser et cies to bacterial cell walls. Solutes coordinated or al., 2002; Gleisner and Herbert, 2002; Johnson et al., adsorbed onto bacteria may undergo accelerated 2002; Nealson et al., 2002; Hamilton, 2003) to yield movement through porous media (Yee and Fein, solid-phase products. Fe(III)-reducing bacteria appear in some settings to depend on coordination of dis- * Tel.: +1-269-387-2878; fax: +1-269-387-5513. solved or colloidal Fe(III) with the bacterial surface E-mail address: [email protected] (J.R. Haas). as a prerequisite to growth (Nealson et al., 2002; URL: http://homepages.wmich.edu/fjhaas/. Haas and DiChristina, 2002; Luu and Ramsay, 2003).

Thus, understanding the acid–base and ion-coordina- carboxyl and phosphoryl groups as mainly responsi- tive properties of the bacterial cell surface becomes ble for Cd and U coordination at the bacterial surface. an important consideration in efforts to predict the Haas et al. (2001) examined a common Gram- biogeochemical behavior of natural or anthropogenic negative, facultatively aerobic metal-reducing bacteri- solutes in aquatic settings. um (Shewanella putrefaciens), and reported titration Recent efforts to quantify the proton- and cation- and metal-adsorption experimental data that were exchange properties of some common bacteria (e.g. modeled using a discrete multisite approach. In Haas Fein et al., 1997; Fein, 2000; Martinez et al., 2002) et al. (2001) S. putrefaciens surfaces were modeled as have yielded results broadly consistent with expect- having three distinct populations of sites; carboxyl ations based on well-known aspects of bacterial (pKa = 5.2), phosphoryl (7.1) and amine (9.4). These membrane chemistry; i.e. that the outer membrane results are broadly consistent with findings for aerobic of a bacterium, albeit complex, comprises a limited Gram-positive bacteria (e.g. Bacillus, Escherichia), and discrete set of biomolecular identities (functional although site densities and precise values of pKa vary group types) arranged upon an organized matrix of significantly among the selected bacteria. Further repeating structural elements (the phospholipid bilay- variations in cell surface properties are reported by er and associated extracellular lipopolysaccharide or Daughney et al. (2001), who provide titration data for peptidoglycan). Fein et al. (1997), Daughney and B. subtilis at different stages of growth (exponential- Fein (1998) and Daughney et al. (1998) have dem- phase, stationary-phase, sporulating). Those authors onstrated that the surface properties of a common measured significant differences in surface site densi- aerobic Gram-positive bacterial group, Bacillus, may ty and site-population pKa values at different stages of be approximated using a discrete multisite surface- B. subtilis growth in laboratory culture. These varia- complexation model accounting for both autoioniza- tions in apparent surface properties among common tion and solute coordination across a range of pH and bacterial types at different stages of growth are ionic strength values. Cox et al. (1999) and Martinez problematic to constructing an overall generic ‘bacte- et al. (2002) report similar results for B. subtilis and rial component’ in generalized models of solute spe- Escherichia coli. ciation under natural conditions. Reactive transport These studies indicate that the surfaces of Gram- models designed to incorporate bacterial influences positive and Gram-negative bacteria may be approx- into a quantitative and predictive depiction of solute imated as possessing at least three discrete site types fate and transport could most usefully employ a (carboxyl, phosphoryl, hydroxyl or amine), and pos- bacterial component that is broadly representative of sibly also sulfhydryl or phosphodiester groups, each as many different microbial species as possible. Con- population of sites exhibiting an overall average acid sidering the diversity of microbial species found in a dissociation constant (pKa) and site density. Most typical aquatic setting (Nealson and Stahl, 1997), one studies have shown that Gram-positive surfaces are must acknowledge that modeling each individual best modeled using a three-site approach (e.g. Daugh- bacterial strain in a natural consortium is an imprac- ney et al., 1998), however, some studies have reported tical task. However, it remains yet unclear to what up to five discrete site populations on Gram-positive degree the surface properties of bacteria vary. Further- and Gram-negative bacteria (e.g. Cox et al., 1999; more, most studies to date have focused on bacteria Martinez et al., 2002). Further work by Fein et al. grown under nutrient-rich conditions in the laboratory, (2002) has extended the B. subtilis surface model to although natural conditions are normally much more lower pH conditions than previously examined (Fein nutrient-limited. et al., 1997), and recommends modeling B. subtilis as In this study the proton-exchange properties of the a four-site surface having acidic functionalities with common Gram-negative, facultatively aerobic, Fe(III) pKa values ranging from f 3to f 9. The identities and Mn(IV)-reducing bacterium S. putrefaciens are of modeled functional groups are not detectable with examined as a function of pH, media composition, titration or metal uptake data, however, X-ray absorp- culture age, strain type, and aerobic/anaerobic con- tion spectroscopic (XAS) work (Kelly et al., 2001) is ditions during growth. Potentiometric acid–base titra- consistent with bulk metal-adsorption data supporting tions were conducted using viable bacteria grown J.R. Haas / Chemical Geology 209 (2004) 67–81 81

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Adsorption of ferrous ions onto Bacillus subtilis cells

Xavier Chaˆtelliera,*, Danielle Fortinb

aCNRS, UMR 6118, Ge´osciences Rennes, Universite´ de Rennes 1, Campus de Beaulieu, F-35042 Rennes Cedex, France bDepartment of Earth Sciences, University of Ottawa, Ottawa, ON, Canada, K1N 6N5 Received 28 January 2004; accepted 16 August 2004

Abstract

Bacteria are known to have reactive surfaces that can affect the cycling of dissolved elements through sorption reactions. The properties of adsorption of various metallic ions (Cd2+,Pb2+, etc.) onto various types of Gram-positive or -negative bacterial surfaces have been measured in several previous studies, but the behavior of ferrous ions has been somewhat left out, possibly because ferrous ions are easily oxidized under standard laboratory conditions. In this paper, the adsorption of ferrous ions onto Bacillus subtilis (a Gram-positive bacterium) was measured under anoxic conditions as a function of pH for various Fe(II)/bacteria ratios. Our results first indicate that Fe(II) adsorption increased with pH and that the adsorption was reversible. An equilibration time of one hour was found sufficient to reach equilibrium between the bacterial surfaces and the solution. Acid–base titrations of the bacterial suspensions were also performed. The titration data could be fitted using a simple three-site model. Neglecting the electrostatic component, the three sites had pK values of 4.45 (pK1), 6.74 (pK2), and 9.08 (pK3) with corresponding concentrations of 5.6, 5.3, and 6.1 (104 mol/dry g of bacteria), respectively. The high density of sites available for proton adsorption suggests that protons can easily diffuse within the cell walls. The adsorption data for Fe(II) was fitted with a single-site model. A Langmuir adsorption isotherm suggested that the site concentration available for the adsorption of the ferrous ions was equal to 3.5104 mol/g which as far less than for protons. The modeling analysis of the adsorption isotherms also indicated that the adsorption took place onto the type 2 sites (pK2=6.74), or type 3 sites (pK3=9.08) and that the adsorption constant of the Fe(II) ions onto the protonated sites was equal to log(TV)=1.22F0.10. A two-site model did not improve the quality of the fit, but indicated that the type 1 sites (pK1=4.45) could contribute only to a limited extent to the adsorption process. D 2004 Elsevier B.V. All rights reserved.

Keywords: Ferrous ions; Adsorption; Adsorption isotherms; Desorption; FITEQL 2.0; Constant capacitance model; Bacillus subtilis; Bacteria

1. Introduction

* Corresponding author. Tel.: +33 223 5269; fax: +33 223 23 6090. Bacteria are ubiquitous in low-temperature surface E-mail address: [email protected] environments (Madigan et al., 2003). Because of their (X. Chaˆtellier). small size, they have a high surface area per unit

0009-2541/$ - see front matter D 2004 Elsevier B.V. All rights reserved. doi:10.1016/j.chemgeo.2004.08.013 210 X. Chaˆtellier, D. Fortin / Chemical Geology 212 (2004) 209–228 weight. Considering that they tend to colonize mineral we investigated experimentally the thermodynamics surfaces as well as travel in the aqueous phases, they and kinetics of adsorption of ferrous ions onto B. can represent a significant fraction of the reactive subtilis cells under anoxic conditions at various pH surface area available for the adsorption of dissolved and various bacteria/Fe(II) ratios. An acid–base molecules or the adhesion of nanominerals, provided titration was also performed to relate the reactivity that their concentration is large enough. Bacteria are, of the bacterial surfaces towards protons and towards for instance, believed to play a significant role in the ferrous ions, and the titration data and the adsorption cycling of multivalent inorganic cations, such as lead isotherms were modeled using a surface complexation or cadmium ions, which can sorb onto their surfaces model (SCM). (Boyanov et al., 2003; Daughney and Fein, 1998; Daughney et al., 2001; Fein et al., 1997, 2001; Kelly et al., 2001, 2002; Kulczycki et al., 2002; Martinez 2. Materials and methods and Ferris, 2001; Ngwenya et al., 2003; Yee and Fein, 2001). They are also known to affect the dissolution 2.1. Anaerobic solutions of mineral phases (Lovley, 2000; Roden and Urrutia, 2002; Shelobolina et al., 2003; Southam, 2000), or the Anaerobic water was prepared by boiling ultra- precipitation of secondary solid phases, such as pure water for 10 minutes, transferring it into an amorphous silica, and of iron or manganese oxides. anaerobic chamber with a 5% H2/95% N2 atmosphere, In precipitation reactions, bacteria are thought to play and bubbling it with the same gas mixture while a nonmetabolic role, by offering their reactive surfaces stirring vigorously for another 5 h. All anaerobic for nucleation or adhesion and growth of crystal solutions, such as anaerobic electrolyte solutions nuclei, and, in some instances, a more active role, by (NaCl), were prepared using anaerobic water and taking part metabolically in the chemical reactions were stored under anaerobic conditions. A palladium leading to the precipitation of the mineral phases catalyst was replaced daily in the chamber in order to (Chaˆtellier et al., 2001, 2004; Emerson, 2000; Ferris et continuously remove the traces of oxygen, which al., 1986; Fortin and Ferris, 1998; Fortin et al., 1997, might have penetrated it. The Fe(II) stock solution 3 1998; Francis and Tebo, 2002; Kennedy et al., 2003; was prepared by dissolving 1.0010 mol of FeCl2 Schultze-Lam et al., 1996; Yee et al., 2003). The exact in 1 ml of 1 M HCl, to which 999 ml of an anaerobic mechanisms through which bacteria can contribute to 103 M NaCl solution were added. The pH of the the dissolution of iron minerals or to their precip- stock solution was kept low (pH 3) in order to prevent itation are still under investigation. For instance, it is the oxidation of the ferrous ions in case of oxygen still not clear when iron-oxide nanocrystals nucleate contamination. As the equilibrium concentration of directly onto the bacterial cell walls and when they oxygen in aqueous solutions in contact with the nucleate in the bulk (Chaˆtellier et al., 2001, 2004). It atmosphere is on the order of 2.65104 M(Lang- is also believed that adsorption of ferrous ions onto muir, 1997), and because the 1 M HCl solution used bacterial cells can inhibit their ability to reduce ferric to dissolve the iron(II) salt was diluted 1000 times, the iron minerals (Liu et al., 2001; Roden and Urrutia, initial concentration of oxygen in the FeCl2 stock 2002; Urrutia and Roden, 1998). To investigate these solution was likely on the order of 2.65107 M. questions, it is important to understand how ferrous Given that 1 mol of oxygen can oxidize 4 mol of and ferric ions adsorb onto and desorb from bacterial ferrous ions (Cornell and Schwertmann, 2003), we surfaces. The adsorption of ferric ions onto Bacillus estimated that the maximal production of iron(III) in subtilis has recently been investigated by Daughney et the stock solution was on the order of 106 M, which al. (2001) at low pH, i.e., in the pH range where the we deemed to be negligible. Anaerobic solutions of ferric ions are still relatively soluble. The adsorption NaOH (102 and 103 M) and HCl (102 and 103 of ferrous ions onto bacterial cells has been quanti- M) were prepared by diluting 1.0 N standard NaOH tatively studied only in the form of Langmuir and HCl solutions in anaerobic 103 M NaCl. The isotherms at neutral pH (Liu et al., 2001; Roden and acid solutions were then stirred overnight in the Urrutia, 2002; Urrutia and Roden, 1998). As a result, anaerobic chamber, in order to remove the trace 226 X. Chaˆtellier, D. Fortin / Chemical Geology 212 (2004) 209–228

1 sites were dominant at low and intermediate pH ions onto B. subtilis cells. The sites with a pK¯ value conditions. However, for the adsorption isotherm at of 4.45 likely played a small role in the adsorption 500 mg/l, the contribution of the type 1 sites was process, even at low pH conditions. The sites with a dominant only at low pH values, up to about pH 4.5– pK¯ of 6.74 were the best candidates as dominant 5.0 at the most. For the adsorption isotherm at 5 g/l, reactive sites for adsorption. In most previous no dominant contribution of the sites of type 1 could studies, it has been proposed that the sites with a be introduced, even at low pH. Modeling all three low pK (pK=4.45 in our case) were carboxylic sites adsorption isotherms together, it was also found that and that they were often the dominant sites for the no significant contribution of the sites of type 1 could adsorption of metallic cations. We have shown that be introduced above pH 4.0–4.5, even for the such conclusions may have to be reconsidered in adsorption isotherm at 50 mg/l (Fig. 6b). It is some instances. Studies examining the adsorption of important to keep in mind that Fig. 6 represents the metallic cations onto bacterial cells using biochem- maximal potential contribution of the sites of type 1 in ical or spectroscopic techniques have suggested that the two-site model, and that their actual contribution either phosphodiester or carboxylic groups are the might be much smaller. Altogether, the modeling of dominant reactive groups, depending on the pH or the adsorption isotherms with a two-site model on the nature of the cation considered (Beveridge indicated that the sites of type 1 were not needed to and Murray, 1980; Boyanov et al., 2003; Hennig et describe the experimental data and that they might al., 2001; Kelly et al., 2001; Panak et al., 2000, have contributed only to a minor fraction of the 2002a,b; Sarret et al., 1998). These sites are believed adsorption except, possibly, in the low pH range to have pK values lower or equal to 5. Hence, there (pHb4.5). It also indicated that an adsorption isotherm is an apparent contradiction between the experimen- at a single bacterial concentration is insufficient to tal data obtained by titration and batch adsorption make conclusions regarding the nature of the reactive experiments, and the data obtained using other sites. techniques. Because the realization of adsorption isotherms is the best way to obtain quantitative information on the values of the adsorption constants 4. Conclusion of the reactive sites, we believe that identifying the chemical nature of the reactive sites, as detected by In this paper, we have demonstrated that ferrous the titration/adsorption data, is a critical point, which ions do adsorb reversibly onto the surface of B. still needs to be resolved. We have also shown that subtilis, and that the adsorption equilibrium is determining a bbest fitQ for a single adsorption reached within minutes. The adsorption isotherms isotherm is a somewhat meaningless task, as many are similar to those obtained in previous studies for different sets of parameters can fit the data almost other metallic divalent cations and in particular to equivalently well. Rather, acceptable families of fits those obtained with cadmium ions. The concentra- should be determined. Fitting a set of adsorption tion of detected reactive sites was on the order of 16 isotherms realized at different bacteria/metal ratios and 3.5104 mol per dry g of bacteria for the can also restrict efficiently the number of parameters protons and for the ferrous ions, respectively. The leading to a good fit of the data. We have proposed titration data revealed the presence of at least three here that all of our data can be best and most simply types of proton-reactive sites, with pK¯ values of fitted when either the sites of type 2 or the sites of 4.45, 6.74, and 9.08, respectively. Several previous type 3 are the dominant reactive sites, with a studies have obtained similar titration curves, but the concentration of 3.5104 mol/g and an adsorption relationship between the proton-reactive sites constant pT¯ Vequal to about 1.2. However, our best detected by the titrations and those detected by fits were still insufficiently good with respect to the biochemical analyses of the cells enveloppes is still quality of the experimental data. In particular, we unresolved. Our adsorption data and modeling were unable to properly describe the adsorption analysis indicated that only one kind of sites was isotherm at 5g/l, even by taking into account two needed to predict adsorption isotherms of ferrous kinds of sites or by trying to use various available X. Chaˆtellier, D. Fortin / Chemical Geology 212 (2004) 209–228 227 electrostatic models (not shown here). The develop- Beveridge, T.J., Murray, R.G.E., 1980. Sites of metal deposition in ment of better models, which will take into account the cell wall of Bacillus subtilis. J. Bacteriol. 141 (2), 876–887. Boyanov, M.I., Kelly, S.D., Kemner, K.M., Bunker, B.A., Fein, the complex and three dimensional structure of the J.B., Fowle, D.A., 2003. Adsorption of cadmium to Bacillus bacterial enveloppes may be needed to achieve a subtilis bacterial cell walls: a pH-dependent X-ray absorption good description of the adsorption isotherms and fine structure spectroscopy study. Geochim. Cosmochim. Acta to resolve the apparent discrepancies between the 67 (18), 3299–3311. data obtained by spectroscopic and biochemical Chaˆtellier, X., Fortin, D., West, M.M., Leppard, G.G., Ferris, F.G., 2001. Effect of the presence of bacterial surfaces during the techniques and those obtained by batch adsorption synthesis of Fe oxides by oxidation of ferrous ions. Eur. J. experiments. Mineral. 13, 705–714. Meanwhile, our results are already of importance Chaˆtellier, X., West, M.M., Rose, J., Fortin, D., Leppard, G.G., with respect to the study of the cycle of iron and Ferris, F.G., 2004. Characterization of iron-oxides formed by associated elements at redox boundaries. They indi- oxidation of ferrous ions in the presence of various bacterial species and inorganic ligands. Geomicrobiol. J. 21 (3), 99–112. cate for instance that bacterial cell surfaces can Cornell, R.M., Schwertmann, U., 2003. The Iron Oxides. Wiley- significantly reduce the dissolved concentration of VCH. Fe(II), thereby reducing the kinetics of the oxidation/ Daughney, C.J., Fein, J.B., 1998. The effect of ionic strength on the hydrolysis reaction in the bulk of the precipitation adsorption of H+,Cd2+,Pb2+, and Cu2+ by Bacillus subtilis and reaction of a Fe(III) solid phase. However, the Bacillus licheniformis: a surface complexation model. J. Colloid Interface Sci. 198, 53–77. adsorption is reversible and the exchange between Daughney, C.J., Fowle, D.A., Fortin, D., 2001. The effect of growth the adsorbed and desorbed Fe(II) should not be phase on proton and metal adsorption by Bacillus subtilis. neglected in a system where the chemical conditions Geochim. Cosmochim. Acta 65 (7), 1025–1035. vary over time. In an oxic system, this exchange Emerson, D., 2000. Microbial oxidation of Fe(II) and Mn(II) at should ultimately lead to the oxidation of most of the circumneutral pH. In: Lovley, D.R. (Ed.), Environmental Microbe–Metal Interactions. ASM Press, pp. 31–52. Fe(II) either as an adsorbed species on the cells Fein, J.B., Daughney, C.J., Yee, N., Davis, T.A., 1997. A chemical surface or, at least, in the bulk phase. Various equilibrium model for metal adsorption onto bacterial surfaces. questions still need to be investigated quantitatively. Geochim. Cosmochim. Acta 61 (16), 3319–3328. For instance, the kinetic of oxidation of the adsorbed Fein, J.B., Martin, A.M., Wightman, P.G., 2001. Metal adsorption Fe(II) in oxygenated conditions have not yet been onto bacterial surfaces: development of a predictive approach. Geochim. Cosmochim. Acta 65 (23), 4267–4273. measured, as well as the potential role of adsorbed Ferris, F.G., Beveridge, T.J., Fyfe, W.S., 1986. Iron–silica crystallite Fe(II) on the activity of iron-reducing bacteria in nucleation by bacteria in a geothermal sediment. Nature 320, anoxic conditions. 609–611. Fortin, D., Ferris, F.G., 1998. Precipitation of iron, silica, and sulfate on bacterial cell surfaces. Geomicrobiol. J. 15, 309–324. Acknowledgments Fortin, D., Ferris, F.G., Beveridge, T.J., 1997. Surface- mediated mineral development by bacteria. In: Banfield, J.F., Nealson, K.H. (Eds.), Geomicrobiology: Interactions We thank Paul Kenward for helping in the Between Microbes and Minerals, vol. 35. Mineralogical Society laboratory and Chris Daughney for useful discussions of America, pp. 161–180. and for fitting our experimental data with FITEQL Fortin, D., Ferris, F.G., Scott, S.D., 1998. Formation of Fe-silicates 2.0R and with various electrostatic models. This and Fe-oxides on bacterial surfaces in samples collected near project was partially funded by an NSERC grant to hydrothermal vents on the Southern Explorer Ridge in the northeast Pacific Ocean. Am. Mineral. 83, 1399–1408. D. Fortin. [LW] Francis, C.A., Tebo, B.M., 2002. Enzymatic manganese(II) oxidation by metabolically dormant spores of diverse Bacillus species. Appl. Environ. Microbiol. 68 (2), 874–880. References Hennig, C., Panak, P.J., Reich, T., Rossberg, A., Raff, J., Selenska- Pobell, S., Matz, W., Bucher, J.J., Bernhard, G., Nitsche, H., Achouak, W., Thomas, F., Heulin, T., 1994. Physico-chemical 2001. EXAFS investigation of uranium (VI) complexes formed surface properties of rhizobacteria and their adhesion to rice at Bacillus cereus and Bacillus sphaericus surfaces. Radiochim. roots. Colloids Surf., B Biointerfaces 3, 131–137. Acta 89, 625–631. Archibald, A.R., 1989. The Bacillus cell enveloppe. 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Characterization of U(VI)-Acidithiobacillus ferrooxidans complexes using EXAFS, transmission electron microscopy, and energy-dispersive X-ray analysis

By M. Merroun∗, C. Hennig, A. Rossberg, T. Reich1 and S. Selenska-Pobell

Institute of Radiochemistry, Forschungszentrum Rossendorf, D-01314 Dresden, Germany

(Received August 9, 2002; accepted in ﬁnal form May 20, 2003)

Acidithiobacillus ferrooxidans eco-types / tion of heavy metals and radionuclides to biomass. It en- Uranium complexation / EXAFS / TEM / EDX compasses both adsorption (accumulation of substances at a surface or interface) and absorption (the almost uniform penetration of atoms or molecules of one phase forming Summary. We used a combination of Extended X-ray a solution with a second phase) [8]. Bacteria express a wide Absorption Fine Structure (EXAFS) spectroscopy, Trans- variety of complex molecules on their surfaces, which at mission Electron Microscopy (TEM) and Energy-Dispersive physiological pH values contain numerous charged chemical X-ray (EDX) analysis to conduct molecular scale studies on groups (such as phosphoryl, carboxyl, and amino groups) U(VI) interaction with three recently described eco-types of that usually give the cell surface a net anionic (negative) Acidithiobacillus ferrooxidans. On the basis of the information obtained by using these methods, we concluded that uranyl charge density. Since the cell surface is in direct contact phosphate complexes were formed by the cells of the three with the environment, the charged groups within the surface eco-types studied. The uranium accumulated by A. ferro- layers are able to interact with ions or charged molecules oxidans cells was located mainly within the extracellular present in the external milieu. As a result, metal cations polysaccharides, and on the cell wall. Smaller amounts were can become electrostatically attracted and bound to the cell also observed in the cytoplasm. surface [7]. Uranium immobilization can take place due to processes Introduction that include transport and intracellular sequestration by complexation with specific metal binding components such Radioactive wastes are classified as high-level wastes (spent as polyphosphate bodies [10], or by precipitation as in- fuel etc.), transuranic wastes (wastes from spent fuel repro- soluble organic and inorganic compounds, e.g, sulfides or cessing, nuclear weapons), and low- and intermediate-level phosphates [6]. wastes which are generated from a variety of activities. The Recently, three eco-types of A. ferrooxidans were re- presence of the actinides U, Th, Np, Pu, and Am in the covered from different sites and depths of two uranium radioactive wastes is of major concern because of their po- mining waste piles [11–13]. These types interact and tol- tential for migration from the waste repositories and for erate uranium in a different way [14]. The objective of the long-term contamination of the environment [1]. Effective present work was to obtain information about the coordina- remediation and stewardship require a better understanding tion surrounding and the cellular localization of the uranium of the environmental chemistry of these actinides and also of accumulated by the mentioned A. ferrooxidans eco-types. their interactions with bacteria and other biological compo- To address this aim, U LIII-edge Extended X-ray Absorp- nents present in the above mentioned environments [2]. tion Fine Structure (EXAFS) spectroscopy, Transmission Results of experimental investigations of uranium accu- Electron Microscopy (TEM), and Energy-Dispersive X-ray mulation and studies of microbial ecology in uranium-rich (EDX) analysis were performed. environments strongly imply that microorganisms can play significant role in the uranium immobilization [3]. Biolog- ical processes for uranium immobilization include reduc- Experimental tion [4, 5] and precipitation [6], biosorption [1, 7], bioaccu- Bacterial strains mulation [8], and biomineralization [1, 9]. The term biosorption is used to describe the metabolism-independent sorp- The bacterial strains used in this work were: A. ferrooxidans W1 (type I), A. ferrooxidans ATCC 33020 (type II), and *Author for correspondence A. ferrooxidans D2 (type III). The strains were cultured in (E-mail: [email protected]). 1 9 K liquid medium containing 3 g of (NH4)2SO4,0.5g of Present address: Johannes Gutenberg-Universität, Institute of . · . · Nuclear Chemistry, Fritz-Strassmann-Weg 2, D-55128 Mainz, K2HPO4,442 g of FeSO4 7H2O, 0 5g ofMgSO4 7H2O, Germany 0.1gofKCl,and0.014 g of Ca(NO3)2 ·4H2O per liter.

Brought to you by | University of Notre Dame Authenticated | 10.248.254.158 Download Date | 8/23/14 9:53 PM 588 M. Merroun et al. tion numbers for oxygen and phosphorous for the EXAFS Figs. 5a and 5b) are very similar to the results obtained by analysis of uranium in bacteria [28] and plants [29]. the difference technique. Small differences in N and σ 2 be- In order to isolate the minor components from the whole tween the ﬁts of raw and ﬁltered residual EXAFS spectra

EXAFS spectrum, the difference technique was used [30]. could be due to nonseparable U−O1,2 contributions, which This method can increase the sensitivity in the determination superimpose onto the U−P1 contribution. of the EXAFS structural parameters of more distant coordi- The radial U−Oax distance of 1.77±0.02 Å and the aver- 2 2 nation shells of the uranium complexes. Briefly, the differ- age σ for Oax of 0.0016 Å are the same within the experi- ence technique consists of the following steps. First, the raw mental uncertainties for all samples. The coordination num- 3 k -weighted EXAFS spectrum was fitted using a structural bers for Oeq1 lie between five and six, and are in good agree- model that contained only the SS and MS paths of the major ment with previous studies [31–33]. The average U−Oeq1 components, i.e.,Oax and Oeq1. The difference between this bond distance of 2.36 ± 0.02 Å is similar to the value of fit and the experimental EXAFS spectrum, i.e., the residual 2.38±0.02 Å measured by EXAFS for Bacillus subtilis ura- spectrum, contains the contributions of the minor compo- nium complexes [34]. But it is significantly longer than nents and experimental noise. Second, the residual spectrum the U−Oeq1 bond distances of 2.28 ± 0.02 Å found for ura- was Fourier filtered in the range of 1.9 to 3.9Å tosepa- nium(VI) in vegetative cells and spores of Bacillus cereus rate the minor scattering contributions from the experimen- and Bacillus sphaericus [28]. tal noise. Tables 2 and 4 summarize the structural parameters All FTs contain a peak at about 2.3 Å as a minor com- of the minor components determined according to models ponent. After correcting for the scattering phase shift, this A and B, respectively. Figs. 6a and 6b show the best fit to the distance is typical for carbonate groups coordinated to U(VI) Fourier-filtered residual EXAFS spectra. in a bidentate fashion [35, 36]. Indeed, carbon atoms at 2.87 In addition to the difference technique, we also fit the to 2.94 Å provide a good fit to the 2.3 Å FT peak. However, raw EXAFS data with a five-shell model. It included both infrared spectroscopy indicated that the carboxyl groups are the major and minor components according to cases A not implicated in the interaction of A. ferrooxidans cells and B. The results for the scattering paths U−Oax,U−Oeq1, with uranium [37]. Moreover, if this FT peak originates U−Oeq2,U−P1,U−Oeq1−P1,andU−P2 (see Tables 1 and 3, from bidentate carboxyl ligands, the U−Oeq1 bond distance

Table 1. Structural parameters of a [ ] b σ 2 [ 2] c ∆ [ ] the A. ferrooxidans uranium com- Sample Shell N R Å Å E eV Error plexes. d a(W1,dry) U−Oax 2 1.77 0.0016 −19.00.24 U−Oeq1 4.4(4) 2.36 0.0071 d U−Oeq2 1 2.88 0.0024 U−P 3.8(7) 3.64 0.0048 U−Oeq1−P MS 7.6 3.78 0.0031

d b(W1,wet) U−Oax 2 1.77 0.0016 −20.20.18 U−Oeq1 4.8(4) 2.36 0.0084 d U−Oeq2 1 2.87 0.0021 U−P 5.0(9) 3.62 0.0067 U−Oeq1−P MS 10 3.76 0.0058

d c (33 020, dry) U−Oax 2 1.77 0.0020 −19.40.21 U−Oeq1 3.5(3) 2.36 0.0060 d U−Oeq2 1 2.88 0.0028 U−P 4.0(7) 3.63 0.0060 U−Oeq1−P MS 7.9 3.76 0.0054

d d (33 020, wet) U−Oax 2 1.77 0.0019 −19.50.23 U−Oeq1 4.4(4) 2.36 0.0074 d U−Oeq2 1 2.86 0.0015 U−P 3.7(7) 3.63 0.0054 U−Oeq1−P MS 7.4 3.77 0.0041

d e(D2,dry) U−Oax 2 1.77 0.0011 −19.90.29 U−Oeq1 4.2(4) 2.36 0.0067 d U−Oeq2 1 2.86 0.0036 U−P 4.1(6) 3.63 0.0033 U−Oeq1−P MS 8.2 3.77 0.0019

d f(D2,wet) U−Oax 2 1.77 0.0016 −20.10.16 U−Oeq1 5.0(4) 2.36 0.0093 d U−Oeq2 1 2.86 0.0032 U−P 3.5(8) 3.64 0.0034 U−Oeq1−P MS 7.0 3.77 0.0020

a: Errors in coordination numbers are ±25% and standard deviations as estimated by EXAFSPAK are given in brackets; b: errors in distance are ±0.02 Å; c: Debye–Waller factor; d: value ﬁxed for calculation

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M. Merroun, C. Hennig, A. Rossberg, H. Funke, S. Selenska-Pobell, T. Reich* Institute of Radiochemistry Forschungszentrum Rossendorf, P.O. Box 51 01 19, 01314 Dresden, Germany * present address: Institut für Kernchemie Johannes Gutenberg-Universität Mainz, Fritz-Straßmann-Weg 2, 55128 Mainz

1. Introduction

Contamination of the environment with radionuclides, toxic metals, metalloids and organo-metals is of considerable economic and environmental significance. There is considerable interest in the microbiological processes which affect the behaviour of contaminants in natural and engineered environments and also in their potential to bioremediate the contaminants. The extent to which these processes can affect the behaviour of contaminants is dependent on the identity and chemical form of the radionuclide released and the physical and chemical nature of the contaminated site or substance. Microbial processes which solubilise radionuclides increase their bioavailability and potential toxicity, whereas those that immobilise them, reduce their bioavailability. Microorganisms can mobilise radionuclides/metals through autotrophic and heterotrophic leaching, chelation by microbial metabolites and siderophores, and methylation which can result in volatilisation. Conversely, immobilisation can result from sorption to cell components or exopolymers, transport into cells and intracellular sequestration or precipitation as insoluble organic and inorganic compounds [1].

Biosorption processes essentially are chemical ones whereby the biomass acts as a surface upon which metals bind by ligand interactions or by ion exchange. Living and dead microorganisms possess abundant functional groups, such as carboxyl, hydroxyl and phosphate on their surface that bind metal ions [2]. In this paper, we present results of U LIII- edge Extended X-ray Absorption Fine Structure (EXAFS) spectroscopy of the uranium complexes formed by different bacteria strains of Acidithiobacillus ferrooxidans, Stenotrophomonas maltophilia and Bacillus sphaericus which were isolated from uranium mining wastes. Also some reference strains, namely Pseudomonas stutzeri ATCC 17588 and Pseudomonas migulae CIP 105470 were involved in the study because they represent predominant indigenous bacterial populations of the uranium wastes.

2. Material and methods

2.1 Bacterial strains used

The bacterial strains used in this work are: P. stutzeri ATCC 17588, P. migulae CIP 105470, A. ferrooxidans D2, B. sphaericus JG-A12, S. maltophilia JG-2. A. ferrooxidans D2 was recovered from the depths of a Canadian uranium mine; S. maltophilia JG-2 and B. sphaericus JG-A12 were isolated from a uranium mining waste pile near the town of Johanngeorgenstadt in Germany. The two Pseudomonas strains, S. maltophilia JG-2 and B. sphaericus JG-A12 were cultured in nutrient medium containing 5 g of peptone and 3 g of meat extract per liter. The A. ferrooxidans strain was cultured in 9K liquid medium containing 3 . . g of (NH4)2SO4, 0.5 g of K2HPO4, 44.2 g of FeSO4 7H2O, 0.5 g of MgSO4 7H2O and 0.1 g of . KCl and 0.014 g of Ca(NO3)2 4H2O per liter.

26 bidentate refers to an inner-sphere complex where two oxygen atoms of a single carboxyl functional groups are shared with uranyl ion.

B. sphaericus JG-A12 data U-O : 2.26 Å eq1 fit U-P1: 3.59 Å A. ferrooxidans D2

6 U-Oeq1: 2.36 Å

U-P1: 3.63 Å 20 chi(k)] P. stutzeri ATCC 17588 3* U-Oeq1: 2.29 Å

U-P1: 3.64 Å P. migulae CIP 105470 U-O : 2.27 Å FT Magnitude eq1 EXAFS [k U-P1: 3.63 Å S. maltophilia JG-2 U-O : 2.27 Å 0 eq1

U-P1: 3.59 Å 0 2468101214 02468 k [ Å-1] R + ∆ [Å]

3 Fig. 1: Uranium LIII-edge k –weighted EXAFS spectra (left) and corresponding FT (right) of the bacterial uranium complexes.

The FT spectra of the uranium complexes formed by the other bacteria studied in this work (P. stutzeri, P. migulae, S. maltophilia, A. ferrooxidans) contain the same FT peak as found in B. sphaericus (interpreted as a contribution of carbon atom) at about 2.3 Å. After correcting for the scattering phase shift, this distance is typical for carbonate groups coordinated to U(VI) in a bidentate fashion [9]. Indeed, carbon atoms at 2.86 to 2.91 Å provide a good fit to the 2.3 Å FT peak. However, infrared spectroscopy of the same samples indicated that the carboxyl groups are not implicated in the interaction of these bacteria with uranium [10, 11]. Moreover, if this FT peak originates from bidentate carboxyl ligands, the U-Oeq1 bond distance should be approximately 2.46-2.49 Å [12], which is not the case in this work (2.27 to 2.36 ± 0.02 Å). For these two reasons, we exclude the possibility of interpreting this shell as a contribution of carbon atoms. Moreover, this shell is too short to be attributed to MS pathways. Oxygen neighbours (Oeq2) provide a good fit to the residual EXAFS spectrum corresponding to this shell. Thus, we interpret this peak in the FT as oxygen neighbors.

3.2. Phosphorous neighbors

For all samples, the distance between uranium and phosphorus (U-P) is within the range of previously reported values for phosphate bound to uranyl in microorganisms [6, 13].

The broad FT peak between 2.7 and 3.9 Å could according to Kelly et al. [5] arise from:

• a phosphorus and the twofold degenerated 3-legged MS path U-Oeq1-P1 • or from a phosphorus and an additional SS path U-P2 instead of the MS path U- Oeq1-P1

References

[1] G.M. Gadd, in: Interactions of microorganisms with radionuclides, Eds. M.J. Keith-Roach, F.R. Livens (Elsevier, Oxford, 2002) pp.179-203 [2] A.J. Francis, Biotransformation of uranium and other actinides in radioactive wastes, J. Alloys Compounds 271-273 (1998) 78-84 [3] G.N. George, I.J. Pickering, EXAFSPAK: A suite of computer programs for analysis of x-ray absorption spectra, Stanford Synchrotron Radiation Laboratory, Stanford, CA, USA 1995 [4] A.L. Ankudinov, B. Ravel, J.J. Rehr, S.D. Conradson, Multiple-scattering calculations of x-ray- absorption spectra, Phys. Rev. B 52(4) (1998) 7565 [5] S.D. Kelly, K.M. Kemmer, J.B. Fein, D.A. Fowle, M.I. Boyanov, B.A. Bunker, N. Yee, X-ray absorption fine structure determination of pH-dependent U-bacterial cell wall interactions, Geochim. Cosmochim. Acta 65 (2002) 3855-3871 [6] C. Hennig, P.J. Panak, T. Reich, A. Roßberg, J. Raff, S. Selenska-Pobell, W. Matz, J.J. Bucher, G. Bernhard, H. Nitsche, EXAFS investigation of uranium (VI) complexes formed at bacillus cereus and bacillus sphaericus surfaces, Radiochim. Acta 89 (2001) 625-631 [7] J.R. Bargar, R. Reitmeyer, J.J. Lenhart, J.A. Davis, Characterization of U(VI)-carbonato ternary complexes on hematite: EXAFS and electrophoretic mobility measurements, Geochim. Cosmochim. Acta 64 (2000) 2737-2749 [8] J. Howatson, D.M. Grev, B. Morosin, Crystal and molecular structure of uranyl acetate dihydrate, J. Inorg. Nucl. Chem. 37 (1975) 1933-1935 [9] A. Coda, A.D. Giusta, V. Tazzoli, The structure of synthetic andersonite, Na2Ca[UO2(CO3)3]. H2O, Acta Cryst. B37 (1981) 1496 [10] M.L. Merroun, C. Hennig, A. Roßberg, T. Reich, R. Nicolai, K.-H. Heise, S. Selenska-Pobell, Characterization of uranium (VI) complexes formed by different bacteria relevant to uranium mining waste piles, in: Uranium in the aquatic environment, Eds. B. Merkel, Planer-Friedrich, C. Wolkersdorfer (Springer, 2002) p. 505-511 [11] M.L. Merroun, G. Geipel, R. Nicolai, K.-H. Heise, S. Selenska-Pobell, Complexation of uranium (VI) by three eco-types of Acidithiobacillus ferrooxidans studied using time-resolved laser-induced fluorescence spectroscopy and infrared spectroscopy, Biometals 16 (2003) 331-339 [12] M.A. Denecke, T. Reich, S. Pompe, M. Bubner, K.-H. Heise, H. Nitsche, P.G. Allen, J.J. Bucher, N.M. Edelstein, D.K. Shuh, Differentiating between monodentate and bidentate carboxylate ligands coordinated to uranyl ions using EXAFS, J. Phys. IV France 7 C2 (1997) 637 [13] S.D. Kelly, M.I. Boyanov, B.A. Bunker, J.B. Fein, D.A. Fowle, N. Yee, K.M. Kemmer, XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH, J. Synchrotron Rad. 8 (2001) 946-948

30 Journal of Colloid and Interface Science 262 (2003) 351–361 www.elsevier.com/locate/jcis

Study of the interaction between europium (III) and Bacillus subtilis: ﬁxation sites, biosorption modeling and reversibility

S. Markai,a Y. Andrès,b,∗ G. Montavon,a and B. Grambow a

a SUBATECH UMR 6457, Université de Nantes, Ecole des Mines, IN2P3/CNRS, Nantes, France b Ecole des Mines de Nantes, GEPEA, 4, Rue Alfred Kastler, BP 20722, 44307 Nantes Cédex 03, France Received 17 July 2002; accepted 16 January 2003

Abstract In order to elucidate the underlying mechanisms involved in the biosorption of metal ions, potentiometric titrations, complexation studies, and time-resolved laser-induced fluorescence spectroscopy (TRLFS) measurements were used to characterize the interaction between Eu(III) and Bacillus subtilis. The reversibility of the interaction between Eu(III) and Bacillus subtilis was studied by a cation-exchange technique using the Chelex resin. For complexation studies in the presence of 0.15 mol/l of NaCl, the metal ion, the biomass, concentrations and the pH were varied. The adsorption data were quantified by a surface complexation model without electrostatic term. The data on the Eu(III)/B.subtilis system at pH 5 were satisfactorily described by one site at which Eu(III) was bound through one carboxylic function of the bacteria. With increasing pH, another site should be considered, involving a phosphate-bound environment. This was partially confirmed by time-resolved laser-induced fluorescence spectroscopy. In addition to this, it was evidenced that the site availability was dependent on the nature of the cation, i.e., a proton or Eu(III). Finally, it was shown that, at pH 5, the Eu(III)/Bacillus subtilis equilibrium was reversible.  2003 Elsevier Science (USA). All rights reserved.

Keywords: Complexation; Distribution; Reversibility; Bioavailability; Europium; Bacillus subtilis; Cation exchange resin; TRLFS

1. Introduction out consideration of electrostatic effects [7,8]. It has been reported that biosorption data involving Gram-negative bac- Heavy metal ions in natural aquatic systems are known teria could be conﬁrmed at the molecular level by spec- to strongly interact with humic substances and microorgan- troscopic analysis [4,9]. But it has been shown that biotic isms, in particular with bacteria [1,2]. The transport proper- variables such as the growth phase can affect the metal ad- ties as well as the bioavailability of these ions in association sorption properties of bacteria [3,10,11]. The spectroscopic with the different constituents of the soil are of particular im- analysis presented in this paper was thus realized on Gram- portance in the context of assessment of toxic metal ion mo- positive bacteria. In addition to this, a lack of information ex- bility in contaminated soils as well as for potential radionu- ists concerning the reversibility of the Eu(III)/Bacillus sub- clide migration from nuclear waste repositories [3,4]. Even tilis interaction. This work tries to bring some elements to thought numerous papers have been devoted to the study of clarify this point. Finally, the effect from electrolyte on com- metal–microorganism interaction [5–9], large uncertainties plexation data was taken into consideration. Among the di- persist, particularly for low metal concentrations [7]. The verse available biomass, Bacillus subtilis was chosen for this present study was focused on the heavy metal biosorption work since the genus Bacillus was found in a large variety of equilibrium in the absence of metabolic processes. The ex- natural habitats. For example, Bacillus strains were isolated periments were performed for a wide range of metal ion con- from a uranium mining waste pile in Saxony (Germany). It −10 −4 centrations (10 to 10 mol/l) and particularly for very was shown that this species could accumulate high amounts −6 low metal concentrations (< 10 mol/l). To quantify these of uranium (VI) [4]. Bacillus subtilis is a Gram-positive, aer- interactions, a surface complexation model was used with- obic, rod-shaped bacterium and ubiquitous in soils and waters. Its parietal structure is well known and is composed * Corresponding author. primarily of peptidoglycan and teichoic acid [5]. Peptidogly- E-mail address: [email protected] (Y. Andrès). can is a polymer of acetylglucosamine and acetylmuramic

0021-9797/03/$ – see front matter  2003 Elsevier Science (USA). All rights reserved. doi:10.1016/S0021-9797(03)00096-1 S. Markai et al. / Journal of Colloid and Interface Science 262 (2003) 351–361 357

In contrast to this, the long lifetime τ2 = 730 ± 30 µs, corresponded to a site in which eight water molecules were excluded from the first coordination shell, which was a very surprising result. The same surprising number of excluded water molecules was obtained also on the Eu(III)-glycerol-2-phosphate reference system in which the metal ion interacted only with phosphate groups. However, glycerol-2-phosphate constituted one of the components of the cell wall and offered only one ≡P–OH interaction site at the bacteria surface, while all other P–OH functions were esterified to form the membrane. Consequently, the long lifetime on Eu(III)/Bacillus subtilis could be attributed to a Eu-phosphate bound. The involvement of phosphate groups of Bacillus subtilis in the metal uptake has been showed by EXAFS measurements on Bacillus sp./U(VI) systems [32,33] for U(VI) biosorption or on Pb(II)/Penicillium chrysogenum [34]. This was strengthened by the fact that the Fig. 3. Sorption isotherms of Eu(III) on Bacillus subtilis as a function of pH observed contribution of the long lifetime branch on the de- values, I = 0.15 mol/l. The dashed line was generated by FITEQL for a − 1:1 Eu:R–COO complex. The dotted line is the fit obtained based on the cay curve increased with pH whereas the relative contribu- − tion of the short lifetime branch decreased with pH (see Ta- assumption that a 1:2 Eu:R–COO complex is present at the cell wall. The solid line is the better fit obtained assuming the assumption that there are a ble 2). Indeed, based on potentiometric titration results, the − − 1:1 Eu:R–COO complex and a 1:1 Eu:R–PO4 complex at the surface. deprotonation rate of phosphate groups increased with pH in contrast to the carboxylic groups. This strengthened the as- Table 3 sumption that the first binding site would involve carboxy- Quantitative description of adsorption of Eu(III)/Bacillus subtilis system late functions whereas, for the second binding site, Eu might as modeled by FITEQL (the site concentrations are determined for 1 g of be bound to the bacteria by phosphate functional group. dried bacteria and are compared to the one found from acid/base titration; Finally, TRLFS measurements indicated that two types see text) of sites were present at the bacterial surface involving a Experiment Site 1 attributed to Site 2 attributed to 1:2 Eu:COO− environment and a site in which Eu was in carboxylic groups phosphate groups = ± = ± a phosphate environment. These results will be taken into Eu(III) adsorption Log K1 7.13 0.4LogK2 8.14 0.5 =− ± =− ± consideration to fit the complexation data. LogS1 5.70 0.15 Log S2 5.83 0.15 Acid/base titration Log S1 =−4.02 ± 0.08 Log S2 =−4.29 ± 0.16

6.3. Biosorption experiments and modeling Si are expressed in mol/g.

The adsorption data are displayed as a function of pH in ment with results of Fein et al. [7] on Al(III)/Bacillus sub- Fig. 3, for constant biomass and the metal ion concentra- tilis systems for low pH values. Thus, by considering a 1:1 tions. In the other hand, adsorption isotherms were deter- Eu:COO− site, the experimental data (see Fig. 3, dashed mined for three different pH by varying the bacterial and line) could be explained. Nevertheless, the fit was even im- the metal concentrations. Experimental data are plotted in proved for pH values > 5.5, by considering an additional − − Figs. 4A and 4B. The stronger Eu(III) adsorption with in- site to the 1:1 Eu:COO site, and notably a 1:1 Eu:PO4 creasing pH (see Fig. 3) may be explained by competi- site (see Fig. 3, solid line). This was supported by poten- tion to protonation reactions. At low pH values, all organic tiometric titration results and by the TRLFS results showing functional groups sites are almost fully protonated. Hence, that a phosphate group of the bacteria cell walls could be a small amount of metal ion can be adsorbed by the bac- involved in the interaction with Eu. It must be noticed that terial surface. As pH increases, the cell wall surface sites other stoichiometries remained possible for the phosphate deprotonate sequentially, resulting in increasing adsorption. environment but as the simplest one was well describing the To quantify this competition, a surface complexation model experimental data, this last one would be considered in the without electrostatic term was used, by taking into account following. The first conclusion was that a 1:1 R–COO−:Eu − the parameters of the cell wall surface functional groups de- anda1:1R–PO4 :Eu site appeared to satisfactorily account termined from the acid/base titrations (see Table 3). for the observed adsorption behavior as a function of pH. As deduced from TRLFS results, the existence of only a The consideration of two sites was in agreement with 1:2 Eu:COO− site was first considered to describe the de- TRLFS results; however, the existence of a 1:2 Eu:COO− pendence of Eu sorption with pH (see Fig. 3). Nevertheless, site was not confirmed by surface complexation modeling this 1:2 Eu:COO− site could not explain the dependence results. Consequently, for TRLFS data, the approach devel- of Eu sorption with pH (see Fig. 3, dotted line). Then, the oped to identify the nature and the number of binding sites existence of a 1:1 Eu:COO− site was considered in agree- reached its limits. The hypothesis that a single nature of S. Markai et al. / Journal of Colloid and Interface Science 262 (2003) 351–361 361

[31] S. Lis, Z. Whang, G.R. Choppin, Inorg. Chim. Acta 239 (1995) 139– [33] S.D. Kelly, M.I. Boyanov, B.A. Bunker, J.B. Fein, D.A. Fowle, N. 143. Yee, K.M. Kemmer, J. Synchrotron Radiat. 8 (2001) 946–948. [32] C. Henning, P.J. Panak, T. Reich, A. Rossberg, J. Raff, S. Selenska- [34] G. Sarret, A. Manceau, L. Spadini, J.C. Roux, J.L. Hazemann, Y. Pobell, W. Matz, J.J. Bucher, G. Bernhard, H. Nitsche, Radiochim. Soldo, L. Eybert-Bérard, J.J. Menthonnex, Environ. Sci. Technol. 32 Acta 89 (2001) 625–631. (1998) 1648–1655. Planta (2003) 216: 939–950 DOI 10.1007/s00425-002-0947-6

ORIGINAL ARTICLE

Patrick Carrier Æ Aurore Baryla Æ Michel Havaux Cadmium distribution and microlocalization in oilseed rape (Brassica napus) after long-term growth on cadmium-contaminated soil

Received: 4 June 2002 / Accepted: 24 September 2002 / Published online: 26 November 2002 Springer-Verlag 2002

Abstract Brassica napus (L.) was grown from seeds on a oles and the cell walls. This phenomenon diverted Cd reconstituted soil contaminated with 100 mg Cd kg)1. ions from metabolically active compartments (cytosol, Compared with roots and stems, leaves accumulated chloroplasts, mitochondria), resulting in a reduction of high amounts of Cd. Although the Cd concentration in Cd toxicity in the leaves. the leaves remained high throughout plant growth and no appreciable change was noticed in the total, ex- Keywords Brassica Æ Cadmium Æ Electron tractable or soluble Cd in the soil adhering to the roots, spectroscopic imaging Æ Metal sequestration Æ the symptoms of Cd toxicity (leaf chlorosis, growth re- Photosynthesis Æ Phytochelatin tardation) decreased with time. Cd induced a noticeable accumulation of phytochelatins in young plants (aged Abbreviation ESI: electron spectroscopic imaging 22 days), which decreased in parallel to the disappear- ance of the symptoms of Cd intoxication. The subcellular distribution of Cd in leaves of Cd-acclimated plants Introduction was determined using biochemical, microscopic and metal-imaging techniques. Leaf fractionation by differ- Cadmium is a major environmental contaminant ema- ential centrifugations showed that Cd was present pre- nating from mining, industrial usage, anthropogenic dominantly in the ‘soluble’ fraction corresponding to the activity and use of agro-chemicals (Prasad 1995; Sanita vacuoles and the cytoplasm. Transmission electron mi- di Toppi and Gabbrielli 1999). Cd is also recognized to croscopic analyses revealed that those cell compartments be one of the most phytotoxic metal pollutants. Excess contained electron-dense granules associated with nee- Cd typically causes a number of toxic symptoms in dle-like structures. Cd, and also high amounts of sulfur, plants such as growth retardation, leaf chlorosis, in- was detected in those structures by electron-spectro- duction/inhibition of enzymes and altered stomatal scopic imaging. This technique also showed Cd binding function. In spite of intensive research, the molecular to cell walls by a mechanism that does not involve sulfur mechanisms of Cd toxicity are not known with any atoms. In contrast, very little Cd was found in chlo- certainty. Binding to essential sulfhydryl groups of en- roplasts, and this is consistent with the preservation of zymes or structural proteins and competition with Ca photosynthesis in plants grown on Cd-polluted soil. The ions for binding to Cd-binding proteins are two possible microanalytical results presented here confirm that long- targets of Cd ions in cells. term growth of B. napus on Cd-contaminated soil is It is clear that in order to survive and grow in accompanied by preferential storage of Cd in the vacu- Cd-contaminated environments plants must keep the intracellular concentration of Cd ions at a low level, and this can be achieved by a number of different mecha- P. Carrier Æ A. Baryla Æ M. Havaux (&) nisms. Decreased metal uptake and active metal efflux CEA/Cadarache, DSV, DEVM, have been proposed as the principal mechanisms of Laboratoire d’Ecophysiologie de la Photosynthe` se, metal tolerance, particularly in microorganisms (Silver UMR 163 CNRS CEA, Universite´-Me´diterraneé CEA 1000, and Phung 1996). In higher plants, heavy-metal detox- 13108 Saint-Paul-lez-Durance, France E-mail: [email protected] ification seems to rely mainly on metal chelation by Fax: +33-4-42256265 specific organic molecules and compartmentalization Present address: A. Baryla (Salt et al. 1995a; Rauser 1999; Cobbett 2000; Clemens Cerexagri, 1 rue des Fre` res Lumie` re, 2001). In a number of studies, Cd was localized pre- BP 9, 78373 Plaisir Cedex, France dominantly in the cell vacuoles (Rauser and Ackerley 948

Fig. 9a–d ESI pictures of chloroplasts (Chl) in leaf cells in oilseed rape plants grown on a soil contaminated with 100 mg Cd kg–1. Pictures taken at zero loss (a, c) and the corresponding ESI pictures of Cd distribution (b, d). Plant age = 47 days

induced by the fixation/dehydration treatments of the Interestingly, chloroplasts remained dark in the Cd leaf sections. Nevertheless, the presence of these Cd- and distributional maps of leaf cells obtained by ESI, indi- S-enriched structures confirms the central role of the cating that they contained very little Cd. A relatively low vacuoles in Cd storage in plants. Also, the cellular dis- concentration of Cd in chloroplasts compared with the tribution of Cd measured on fresh material by bioaccumulation level in whole leaves was also concluded chemical methods agrees with the Cd microlocalization from atomic absorption spectroscopic measurements on determined by ESI, suggesting that no or little relocal- isolated chloroplasts. The photosynthetic machinery is ization of Cd occurred during sample preparation for known to be extremely sensitive to Cd (Van Duijvendijk- ESI. Matteoli and Desmet 1975; Atal et al. 1991). Conse- Besides the vacuole, the cell wall seems to be another quently, the fact that photosynthesis was preserved in site of preferential Cd binding in leaf cells of B. napus. leaves of Cd-exposed B. napus plants is another indica- However, in this case, Cd appears to bind through a tion that Cd did not enter massively into the chlorop- mechanism different from that occurring in the vacuoles lasts. As shown previously (Baryla et al. 2001), since S was not associated with Cd. In bacterial cell Cd-induced leaf chlorosis is due to a strong reduction of walls, X-ray absorption fine-structure measurements chloroplast density in leaf cells. As light absorption by have shown that carboxyl functional groups are re- leaves is not linearly related to the chlorophyll content sponsible for complexation of Cd (Kelly et al. 2001). The (Terry 1980), a reduction in the number of photosyn- role of cell walls in Cd binding and storage in plants is thetically competent chloroplasts will have less impact controversial. A number of studies have failed to show on the light absorption and the photosynthetic activity substantial Cd binding in cell walls (Weigel and Ja¨ ger of the leaves than on the chlorophyll concentration. 1980; Rauser and Ackerley 1987; Vo¨ geli-Lange and We have no indication about the Cd concentration in Wagner 1990; Vazquez et al. 1992). In contrast, in other mitochondria but a previous study of oilseed rape plants studies, cell walls were assumed to play the most im- grown on Cd-polluted soil did not reveal any significant portant role in metal accumulation (Khan et al. 1984; inhibition of mitochondrial respiration (Baryla et al. Lindsey and Lineberger 1981). Although the wall of B. 2001) although Cd is known to be a potent inhibitor of napus leaf cells was enriched in Cd, the fractionation the respiratory electron transport chain in vitro (Miller experiment of Fig. 5a seems to indicate that the amount et al. 1973; Carreau and Mazliak 1977). Also, the al- of Cd bound to cell walls represents a relatively small terations usually observed in Cd-stressed chloroplasts fraction of the total Cd storage in B. napus. However, we and mitochondria (Silverberg 1976; Baszynski et al. cannot exclude loss of Cd during leaf fractionation with 1980; Barcelo et al. 1988; Neumann et al. 1994) were not the biochemical method used in this study. observed in the transmission electron micrographs of 949 B. napus leaf cells. Consequently, one can consider that microscopy. In: Mayer F (ed) Methods in microbiology, vol 20. little Cd was translocated to mitochondria, as was the Academic Press, New York, pp 113–146 Blake-Kalff MMA, Harrison KR, Hawkesford MJ, Zhao FJ, case for chloroplasts. Thus, acclimation of B. napus to McGrath SP (1998) Distribution of sulfur within oilseed rape Cd toxicity is correlated with a heterogeneous distribu- leaves in response to sulfur deficiency during vegetative growth. tion of Cd at the cell level, with Cd being diverted from Plant Physiol 118:1337–1344 Cd-sensitive organelles (chloroplasts, mitochondria) to Carreau JP, Mazliak P (1977) Effects of some heavy metals (lead, cadmium, mercury) on animal or vegetable mitochondria res- less metabolically active cellular components such as the piration. Cah Nutr Diet 12:45–51 vacuole and the cell wall. In this way, functions as vital Chardonnens AN, ten Bookum WM, Kuijper LDJ, Verkleij JAC, as photosynthesis and respiration are preserved, pre- Ernst WHO (1998) Distribution of cadmium in leaves of cad- sumably allowing plant acclimation to metal toxicity. In mium tolerant and sensitive ecotypes of Silene vulgaris. Physiol contrast to other plant species (Chardonnens et al. Plant 104:75–80 Choi Y-E, Harada E, Wada M, Tsuboi H, Morita Y, Kusano T, 1998), Cd was not stored in the epidermal cells of Sano H (2001) Detoxification of cadmium in tobacco plants: B. napus leaves. formation and active excretion of crystals containing cadmium Because of its high phytotoxicity, Cd is usually pre- and calcium through trichomes. Planta 213:45–50 sent in relatively small amounts in plant leaves Clemens S (2001) Molecular mechanisms of plant metal tolerance –1 and homeostasis. Planta 212:475–486 (<260 lgg in this study), with the notable exception Cobbett CS (2000) Phytochelatin biosynthesis and function in of some metal hyperaccumulators (see, e.g. Knight et al. heavy-metal detoxification. Curr Opin Plant Biol 3:211–216 1997; Ku¨ pper et al. 2000). Although the Cd concentra- Cotelle V, Forestier C, Vavasseur A (1996) A reassessment of the tion can be locally higher (as actually shown here), this intervention of calmodulin in the regulation of stomatal movement. Physiol Plant 98:619–628 makes difficult the study of Cd distribution in plant cells Coulomb C, Pollan C, Lizzi Y, Coulomb P-J (1996) Ultrastructural by conventional techniques such as X-ray microanalysis. changes concerning the photosynthetic apparatus during Phy- Cellular compartmentation of Cd was achieved recently tophthora capsici infection of susceptible and induced pepper by X-ray microanalysis in Arabidopsis halleri (Ku¨ pper at leaves. C R Acad Sci Paris Life Sci 319:893–899 al. 2000) but this species is a Zn and Cd hyperaccumu- De Knecht JA, van Baren N, Ten Bookum WM, Wong Fong Sang HW, Koevoets PLM, Schat H, Verkleij JAC (1995) Synthesis lator in which the Cd concentration of the shoots was at and degradation of phytochelatins in cadmium-sensitive and least 10 to 20 times higher than that found in B. napus. cadmium-tolerant Silene vulgaris. Plant Sci 106:9–18 This study is one of the first studies where Cd is imaged Ebbs S, Lau I, Ahner B, Kochian L (2001) Phytochelatin synthesis together with S in plant leaf cells (see also Neumann et al. is not responsible for Cd tolerance in the Zn/Cd hyperaccumulator Thlaspi caerulescens (J. & C. Presl). Planta 214:635–640 1994). The Cd distribution images presented here illus- Gerard E, Echevarria G, Sterckeman T, Morel JL (2000) Cadmium trate the potential of ESI for metal micro-mapping in availability to three plant species varying in cadmium accu- plants with a very good spatial resolution (ca. 5 nm) and mulation pattern. J Environ Qual 29:1117–1123 a high sensitivity. Haag-Kerwer A, Scha¨ fer HJ, Heiss S, Walter C, Rausch T (1999) Cadmium exposure in Brassica juncea causes a decline in Acknowledgements We thank C. Sahut for allowing us to use her transpiration rate and leaf expansion without effect on photo- atomic absorption spectrophotometer, B. Dimon for help with synthesis. J Exp Bot 50:1827–1835 image analyses, S. Sauge-Merle and J. Balesdent for helpful dis- Havaux M, Tardy F (1996) Temperature-dependent adjustment cussions and S. Cuine´for help in phytochelatin determination (all of the thermal stability of photosystem II in vivo: possible at CEA/Cadarache). The electron-microscopic studies were done involvement of xanthophyll-cycle pigments. Planta 198:324– with help from P. and C. Coulomb (Universite´d’Avignon, France), 333 B. Perrat-Mabilon (UniversiteĆlaude Bernard-Lyon I, France) Kelly SD, Boyanov MI, Bunker BA, Fein JB, Fowle DA, Yee N, and B. Vacher (Laboratoire de Tribologie et Dynamique des Sys- Kemner KM (2001) XAFS determination of the bacteria cell te` mes, Ecole Centrale de Lyon, Ecully, France) who are gratefully wall functional groups responsible for complexation of Cd and acknowledged. U as a function of pH. J Synch Rad 8:946–948 Khan DH, Duckett JG, Frankland B, Kirkham JB (1984) An X-ray microanalytical study of the distribution of cadmium in roots of Zea mays L. J Plant Physiol 115:19–28 References Knight B, Zhao FJ, Mcgrath SP, Shen ZG (1997) Zinc and cadmium uptake by the hyperaccumulator Thlaspi caerulescens in Atal N, Saradhi PP, Mohanty P (1991) Inhibition of the chlorop- contaminated soils and its effects on the concentration and last photochemical reactions by treatment of wheat seedlings chemical speciation of metals in soil solution. Plant Soil 197:71– with low concentrations of cadmium: analysis of electron 78 transport activities and changes in fluorescence yield. Plant Cell Kumar PBAN, Dushenkov V, Motto H, Raskin I (1995) Phy- Physiol 32:943–951 toextraction: the use of plants to remove heavy metals from Barcelo J, Vazquez MD, Poscenrieder C (1988) Structural and ul- soils. Environ Sci Technol 29:1232–1238 trastructural disorders in cadmium-treated bush bean plants Ku¨ pper H, Lombi E, Zhao F-J, McGrath SP (2000) Cellular (Phaseolus vulgaris L.). New Phytol 108:37–49 compartmentation of cadmium and zinc in relation to other Baryla A, Carrier P, Franck F, Coulomb C, Sahut C, Havaux M elements in the hyperaccumulator Arabidopsis halleri. Planta (2001) Leaf chlorosis in oilseed rape plants (Brassica napus) 212:75–84 grown on cadmium-polluted soil: causes and consequences for Lichtenberger O, Neumann D (1997) Analytical electron micros- photosynthesis and growth. Planta 212:696–709 copy as a powerful tool in plant cell biology: examples using Baszynski T, Wajda L, Krol M, Wolinska D, Krupa Z, Tukendorf electron energy loss spectroscopy and x-ray microanalysis. Eur A (1980) Photosynthetic activities of cadmium-treated tomato J Cell Biol 73:378–386 plants. Physiol Plant 48:365–370 Lichtenthaler HK, Wellburn AR (1983) Determination of total Bauer R (1988) Electron spectroscopic imaging: an advanced carotenoids and chlorophylls a and b of leaf extracts in different technique for imaging and analysis in transmission electron solvents. Biochem Soc Trans 603:591–592 Geochimica et Cosmochimica Acta, Vol. 67, No. 21, pp. 4057–4066, 2003 Copyright © 2003 Elsevier Ltd Pergamon Printed in the USA. All rights reserved 0016-7037/03 $30.00 ϩ .00 doi:10.1016/S0016-7037(03)00214-X Association of uranyl with the cell wall of Pseudomonas fluorescens inhibits metabolism

RIZLAN BENCHEIKH-LATMANI* and JAMES O. LECKIE Department of Civil and Environmental Engineering, Stanford University, Stanford, CA 94305

(Received September 10, 2002; accepted in revised form March 4, 2003)

Abstract—Citric acid is found along with uranyl in the subsurface of former nuclear facilities because of its use as a decontamination agent in the nuclear industry. Citrate’s metal chelating properties affect the mobility of uranyl in the subsurface and consequently, citrate biodegradation may significantly impact uranyl fate and transport. Under the non-growth conditions considered, low (micromolar) uranyl concentrations inhibit the biodegradation of citrate by Pseudomonas fluorescens, a common subsurface denitrifying bacterium. Addi- tionally, uranyl is found readily associated with the cell envelope of P. fluorescens. The observed inhibition appears to be linked to the binding of uranyl to the cell surface and is reversible by desorbing cell-bound uranyl. This study establishes a link between uranyl association with the cell surface and the observed inhibitory effect of uranyl on cell metabolism. Copyright © 2003 Elsevier Ltd

1. INTRODUCTION aquifers at circumneutral pH rarely exceed 30 ␮mol/L (Depart- ment of Energy, 1997). Although DOE sites are heavily con- Uranium is a contaminant of concern in the subsurface taminated, uranyl sorbs strongly at circumneutral pHs and the environment of a majority of former nuclear fuel and weapons majority of uranyl is associated with the sediments at those manufacturing facilities of the Department of Energy (DOE) pHs. (2) Non-growth conditions were used and are representa- (Riley et al., 1992). In the 1950s and 1960s, wastes containing tive of often oligotrophic subsurface environments. Previous a mixture of radionuclides and chelators such as citric acid, studies were conducted under growth conditions in the pres- oxalic acid, ethylenediaminetetraacetic acid (EDTA) or nitrilo- ence of ample supply of a carbon source. In this study, we triacetic acid (NTA) were disposed of in shallow trenches examine the metabolism of P. fluorescens under non-growth (Department of Energy, 1997, 1998; Hartman et al., 2000). In conditions, in the presence of low, environmentally relevant, several cases, the radionuclides have reached the water table uranyl and citrate concentrations. (3) Lastly, and most impor- and are threatening water supply wells or surface waters. The tantly, an additional novelty of this study resides in the fact that chelators are not contaminants in their own right (with the it focuses on the relationship between uranyl biosorption and notable exception of NTA) but are important because they have the inhibition of citrate degradation by uranyl and establishes a the potential to significantly alter uranyl mobility in the sub- link between the two. surface. Several researchers have studied citrate degradation by bac- Specifically, citric acid is known to affect uranyl adsorption teria in the presence of uranyl but exclusively under growth to mineral surfaces. Citrate has been shown to enhance uranyl conditions. Joshi-Tope and Francis (1995) and Francis et al. adsorption on goethite and to reduce the adsorption of uranyl (1992) showed that, in the presence of equimolar uranyl (0.52 on kaolinite (Redden et al., 1998). Thus, the mobility of uranyl mmol/L), there was no degradation of citrate by P. fluorescens. in the subsurface is crucially linked to the persistence of citric However, in cell-free extract, citrate was fully biodegraded acid and other chelators in the subsurface. under similar conditions. The authors concluded that the per- However, citric acid is readily metabolized by a variety of sistence of citrate stemmed from the inability of the cells to microorganisms, both aerobically and anaerobically (Hugen- transport the uranyl-citrate complex across the cell membrane. holtz, 1993; Fuller and Scow, 1997; Coates et al., 1999). An Another study by Banaszak et al. (1999) suggests that the understanding of the propensity of citrate to be biodegraded in uranyl-citrate complex itself is toxic to P. fluorescens. Finally, the presence of uranyl would be useful in assessing its persis- Huang et al. (1998) show that there is incomplete citrate (0.55 tence in uranyl-contaminated environments and thus its role in mmol/L) degradation by a mixed cell culture in the presence of uranyl fate and transport. uranyl (0.42 mmol/L) at pH 6. The goal of this study is to investigate the inhibitory effect of This study shows that uranyl inhibits cell metabolism at low, uranyl on the biodegradation of citrate by Pseudomonas fluo- environmentally relevant concentrations and that the inhibition rescens, a common subsurface denitrifier, at environmentally is associated with the binding of uranyl to the cell envelope. relevant concentrations and under non-growth conditions. The This inhibitory effect is reversible by desorbing cell-bound present study differs from others in several ways: (1) The uranyl. uranyl concentrations used are considerably lower than those in other studies (Ͻ10 ␮mol/L vs. Ͼ400 ␮mol/L). This is significant because measured uranyl concentrations in contaminated 2. MATERIALS AND METHODS 2.1. Thermodynamic Calculations

* Author to whom correspondence should be addressed, at the Scripps Equilibrium speciation calculations were conducted using the chem- Institution of Oceanography, Marine Biology Research Division, 9500 ical equilibrium program HYDRAQL (Papelis et al., 1988). The ther- Gilman Drive, La Jolla, CA 92093-0202, USA ([email protected]). modynamic data used are shown in Table 1. Brieﬂy, the citrate com- 4057 4064 R. Bencheikh-Latmani and J. O. Leckie

complexation was described by Haas et al. (2001) for the association of uranyl with another gram-negative bacterium Shewanella putrefaciens. The precise nature of the uranyl-cell complex is unknown. Researchers have suggested the importance of carboxyl and phosphato functional groups in the biosorption of uranyl by both gram-positive (Fowle et al., 2000; Kelly et al., 2001) and gram-negative bacteria (Haas et al., 2001; Kelly et al., 2002). We are currently investigating the molecular details of uranyl binding to the surface of P. fluorescens. Under our experimental conditions, a large fraction of uranyl present is biosorbed and, consequently, a large fraction of citrate is uncomplexed (Fig. 1). Therefore, citrate is expected to be readily degraded by cells in the presence of uranyl because citrate is present to a large extent as either free citrate or Fig. 12. Energy dispersive X-ray spectrum if the membrane of a cell Kcit2Ϫ, both biodegradable species. However, we observe little exposed to uranyl. X-rays from the uranium L and M shells are clearly citrate degradation in the presence of some concentrations of detectable, while no X-rays from Fe appear. The Ni signal comes from Ն ␮ the grid. uranyl ( 5 mol/L). Similarly, there is little glucose degradation in the presence of 3 ␮mol/L uranyl or higher. These results reveal the inhibitory effect of uranyl on cell metabolism of both citrate and glucose even at micromolar concentrations. 4. DISCUSSION The inhibition appears to be associated with the binding of This study shows that P. fluorescens, a common subsurface uranyl to the cell envelope. A cell wash with bicarbonate denitrifier, can serve as a ligand for uranyl. Under the condi- recovers ϳ90% of uranyl and leads to a recovery (30%) in tions investigated (Fig. 2), cells rapidly scavenge solution ura- metabolic activity of cells. The link between uranyl binding to nyl and even outcompete citrate for uranyl, leading to the the cell surface and the inhibitory effect of uranyl is important dissociation of preformed uranyl-citrate complex(es). Metabol- because it suggests that biosorption inhibits metabolism. The ically inactive cells also cause the dissociation of uranyl-citrate ease of reversibility of the inhibition by washing with a com- complex(es) and scavenge uranyl. Thus, the dissociation of plexing agent strengthens the argument that the inhibitory uranyl-citrate complexes can be attributed to the depletion of effect of uranyl is not associated with intracellular intake of uncomplexed uranyl by binding to the cell envelope, rather than uranyl but rather with binding to the cell surface. We hypoth- to the depletion of uncomplexed citrate by biodegradation. esize that the binding of U may alter the conformation of The ability of P. fluorescens to outcompete citrate for uranyl membrane proteins crucial for the metabolism of citrate and under the conditions studied is probably due to the low con- glucose or that uranyl biosorption reduces membrane fluidity centration of citrate present and the absence of other uranyl- and permeability, thus limiting citrate and glucose uptake. The binding ligands (principally carbonate). As with all ligands, the cells recover only 30% of their activity after washing with amount of uranyl complexed by cells will vary according to the bicarbonate even as 90% of uranyl is in solution, suggesting a relative concentration of uranyl, cells, other ligands and pH. hysteresis effect in the recovery from uranyl biosorption. One situation in which cells may complex a significant amount Several studies have described the toxic effect of uranyl on of uranyl is during the bioremediation of a low-carbonate site, bacterial metabolism. An early study (Barron et al., 1948) where high cell concentrations are expected due to the bioaug- documented the toxic effect of uranyl on yeast and bacteria, mentation or biostimulation strategies. including Pseudomonas aeruginosa. Tuovinen and Kelly Electron micrographs (Figs. 10 and 11) show the association (1974a, 1974b) also recorded the toxic effect of uranyl on the of uranyl with both the outer and the cytoplasmic membranes iron oxidizer Thiobacillus ferrooxidans at acidic pH. They as well as with the periplasm. A study of uranyl biosorption on found that the presence of a chelator such as EDTA or a high another strain of P. fluorescens shows the association of uranyl concentration of competing cations reduced the inhibitory ef- with the cell envelope with the exception of ϳ10% of the cells fect of uranyl on T. ferrooxidans. that accumulate uranyl intracellularly (Krueger et al., 1993). In Huang et al. (1998) studied the degradation of citrate in the that study, uranyl precipitates as platy crystals aligned in the presence of uranyl by a growing mixed culture. At pH 6, they periplasm so that their flat surfaces associate with the periplasm found that 73% of citrate was degraded within 100 h when side of the outer and cytoplasmic membranes. It may be that present in threefold excess but no further reduction took place those crystals are a uranyl phosphate phase formed during a cell in the following 140 h. At that pH, uranyl was sorbed to cells wash with a phosphate buffer. extensively (ϳ50%). At higher citrate:uranyl ratios (8:1), com- The present study scrupulously excluded phosphate. Spec- plete citrate degradation was observed within 130 h at pH 6 but troscopic and isotherm data both show no evidence of uranyl less uranyl was associated with cells. For pHs at which solution precipitation onto the cell surface (data not shown). Moreover, carbonate limits U sorption to cells (e.g., pH 8, 9), citrate a rapid wash of uranyl-laden cells with bicarbonate released degradation was unhindered and proceeded to 100% within ϳ90% of uranyl suggesting the presence of a labile uranyl-cell 10 h even at low citrate:uranyl ratios. The authors attributed the “complex” rather than a precipitate. Thus, we characterize the difference in citrate degradation at pH 6 and 8 to changes in association of uranyl with cells as complexation. A similar uranyl speciation (carbonate vs. citrate complexes). Nonethe- Uranyl inhibits metabolism in P. fluorescens 4065 less, there appears to be a correlation between incomplete from a hydrocarbon-contaminated aquifer. Int. J. Sys. Bact. 49, citrate degradation and sorption of uranyl onto cells in their 1615–1622. data. The study (Huang et al., 1998) was conducted under Department of Energy (1997) Integrated Data Base Report—1996: U.S. Spent Nuclear Fuel and Radioactive Waste Inventories, Pro- growth conditions, with a mixed bacterial community and with jections, and Characteristics. Revision 13, December 1997. http:// considerably higher uranyl and citrate concentrations ([uranyl] www.em.doe.gov/idb97/contents.html. ϭ 0.42 mmol/L and [citrate] ϭ 0.55–3.5 mmol/L). Despite the Department of Energy (1998) Fiscal Year 1997 Integrated Water major differences with conditions in our study, the effect of Quality Program Annual Report for the U.S. Department of Energy Oak Ridge National Reservation, Oak Ridge, Tennessee, BJC/OR- biosorbed uranyl on citrate biodegradation was also observed. 32. U.S. Department of Energy, Washington, DC. Prepared by Other studies show different results. Joshi-Tope and Francis Science Applications International Corp. (1995) and Francis et al. (1992) found that P. fluorescens was Dodge C. J. and Francis A. J. (1997) Biotransformation of binary and unable to degrade citrate in the presence of equimolar uranyl. ternary citric acid complexes of iron and uranium. Environ. Sci. Cell free extract of the cells readily degraded 100% of citrate in Technol. 31, 3062–3067. Feldman I., Havill J. R., and Newman W. F. (1954) Polymerization of the presence of equimolar uranyl. The authors suggested that uranyl-citrate, -malate, -tartrate and -lactate complexes. J. Am. the uranyl-citrate complexes had limited lability, rendering Chem. Soc. 76, 4726–4732. citrate non-bioavailable. Unfortunately, no uranyl measure- Fowle D. A., Fein J. B., and Martin A. M. (2000) Experimental study ments were made and thus was not possible to assess the of uranyl adsorption onto Bacillus subtilis. Environ. Sci. Technol. 34, 17, 3737–3741. fraction of uranyl sorbed to cells in those experiments. Francis A. J., Dodge C. J., and Gillow J. B. (1992) Biodegradation of The toxic species responsible for binding to cells and inhib- metal citrate complexes and implications for toxic-metal mobility. iting metabolism was not identified in this study. However, due Nature 356, 140–142. to the chemical simplicity of the system studied, only four Fuller M. E. and Scow K. M. (1997) Lack of capsular exopolymer 2ϩ effects on the biodegradation of organic compounds by Pseudomo- species can be considered as possible inhibitors: UO2 , ϩ ϩ 2ϩ nas sp. strains JS1 and JS150. Microbiol. Ecol. 34, 248–253. UO2Cl ,UO2OH and (UO2)2(OH)2 . Only two hydroxide Grenthe I., Fuger J., Konings R. J. M., Lemire R. J., Muller A. B., species are considered because, the inhibitory species is likely Nguyen-Trung C., and Wanner H. (1992) Chemical Thermodynam- to be a cation due to its binding to the negatively charged ics of Uranium. Nuclear Energy Agency, OECD, Paris. surface of cells. Banaszak et al. (1999) concluded that a uranyl- Haas J. R., DiChristina T. J., and Wade R., Jr. (2001) Thermodynamics of U(VI) sorption onto Shewanella putrefaciens. Chem. Geol. 180, citrate complex was the toxic species. The design of our ex- 33–54. periments does not allow a definitive conclusion on the toxicity Hartman M. J., Morasch L. F., and Webber W. D. (2000) Hanford Site of the uranyl-citrate complex. However, it is clear that there is Groundwater Monitoring for Fiscal Year 1999. http://hanford.pnl. at least one additional inhibitory species as inhibition was also gov/groundwater/reports/gwrep99/html/start1.htm. observed in the absence of citrate. Huang F. Y. C., Brady P. V., Lindgren E. R., and Guerra P. (1998) Biodegradation of uranium-citrate complexes: Implications for extraction of uranium from soils. Environ. Sci. Technol. 32, 379–382. 5. CONCLUSIONS Hugenholtz J. (1993) Citrate metabolism in lactic-acid bacteria. FEMS Microbiol. Rev. 12, 165–178. This study shows that, under the conditions studied, the Joshi-Tope G. and Francis A. J. (1995) Mechanisms of biodegradation cellular envelope of P. fluorescens serves as a ligand for uranyl. of metal-citrate complexes by Pseudomonas fluorescens. J. Bacte- Additionally, the binding of uranyl to the cell envelope results riol. 177, 1989–1993. in the inhibition of cell metabolism. The inhibition is reversible Kelly S. D., Boyanov M. I., Bunker B. A., Fein J. B., Fowle D. A., Yee by desorbing uranyl. N., and Kemner K. M. (2001) XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J. Synchrotron Rad. 8, 946–948. Acknowledgments—This work was supported by a grant from the Kelly S. D., Kemner K. M., Fein J. B., Fowle D. A., Boyanov M. I., Environmental Management Science Program in the Department of Bunker B. A., and Yee N. (2002) X-ray absorption fine structure Energy (DE-FG07-96ER14698). We are indebted to Jonathan Mulhol- determination of pH-dependent U-bacterial cell wall interactions. land for his generous help with sample preparation, microtomy and Geochim. Cosmochim. Acta 66, 3855–3871. electron microscopy and for the use of the Biology Electron Micros- Krueger S., Olson G. J., Johnsonbaugh D., and Beveridge T. J. (1993) copy Laboratory at Stanford University. Also, we would like to thank Characterization of the binding of gallium, platinum, and uranium to three anonymous reviewers for their constructive input. Pseudomonas fluorescens by small-angle X-ray scattering and transmission electron microscopy. Appl. Environ. Microbiol. 59, 4056–4064. Associate editor: J. B. Fein. Langmuir D. (1997) Aqueous Environmental Geochemistry. Prentice Hall, New York. REFERENCES Lenhart J. J., Cabaniss S. E., MacCarthy P., and Honeyman B. D. (2000) Uranium(VI) complexation with citric, humic and fulvic Allen P. G., Shuh D. K., Bucher J. J., Edelstein N. M., Reich T., acids. Radiochim. Acta 88, 345–353. Denecke M. A., and Nitsche H. (1996) EXAFS determinations of Markovits G., Klotz P., and Newman L. (1972) Formation constants for uranium structures: The uranyl ion complexed with tartaric, citric, the mixed-metal complexes between indium(III) and uranium(VI) and malic acids. Inorg. Chem. 35, 784–787. with malic, citric and tartaric acids. Inorg. Chem. 11, 2405–2408. Banaszak J. E., Rittmann B. E., and Reed D. T. (1999) Subsurface Newman W. F., Havill J. R., and Feldman I. (1951) The uranyl: citrate interactions of actinide species and microorganisms: Implications for system. II. Polarographic studies of the 1:1 complex. J. Am. Chem. the bioremediation of actinide-organic mixtures. J. Radioanal. Nucl. Soc. 73, 3593–3595. Chem. 241, 385–435. Nunes M. T. and Gil V. M. S. (1987) New NMR evidence on the Barron E. S. G., Muntz J. A., and Gasvoda B. (1948) Regulatory uranyl-citrate complexes. Inorg. Chim. Acta 129, 283–287. mechanisms of cellular respiration I. The role of cell membranes: Papelis C., Hayes K. F., and Leckie J. O. (1988) HYDRAQL: A uranium inhibition of cellular respiration. J. Gen. Physiol. 32, 163– Program for the Computation of Chemical Equilibrium Composition 178. of Aqueous Batch Systems Including Surface-Complexation Model- Coates J. D., Ellis D. J., Gaw C. V., and Lovley D. R. (1999) Geothrix ing of Ion Adsorption at the Oxide/Solution Interface. Stanford fermentans gen. nov., sp nov., a novel Fe(III)-reducing bacterium University, Stanford, CA.

Biometals 2002: Third International Biometals Symposium

Focused Meeting Organized by S. Andrews (School of Animal and Microbial Sciences, University of Reading), R. Cammack, R. W. Evans, R. Hider and J. M. Wrigglesworth (Division of Life Sciences, King’s College London), C. C. Cooper, P. S. Dobbin and M. T. Wilson (Department of Biological Sciences, University of Essex), N. J. Robinson (Department of Biochemistry and Genetics, University of Newcastle) and R. J. Simpson (Department of Endocrinology, Diabetes and Internal Medicine, King’s College London), and Edited by R. W. Evans, R. Hider and J. M. Wrigglesworth (Division of Life Sciences, King’s College London), and Sponsored by The Wellcome Trust, British Technology Group, Apotex, Italfarmaco, and the Inorganic Biochemistry Discussion Group, held at King’s College London, London, I 1-13 April 2002. Bioremediation

Molecular and atomic analysis of uranium complexes formed by three eco-types of Acidithiobacillus ferrooxidans M. Merroun’, C. Hennig, A. Rossberg, G. Geipel, T. Reich and S. Selenska-Pobell Forschungszentrurn Rossendorf, Institute of Radiochemistry, D-0 I3 I4 Dresden, Germany

Abstract region of Germany (Saxony and Thuringia). At A combination of EXAFS, transmission electron present a large number of uranium mining waste microscopy and energy-dispersive X-ray was used piles and mill-tailings continue to represent haz- to conduct a molecular and atomic analysis of the ardous sources of environmental pollution in this uranium complexes formed by Acidithiobacillus region [l]. These contaminated sites require long- ferrooxiduns. The results demonstrate that this term stewardship in addition to remediation [2]. bacterium accumulates uranium as phosphate For effective remediation and stewardship, a bet- compounds. We suggest that at toxic levels when ter understanding is needed of uranium environ- the uranium enters the bacterial cells, A. ferro- mental chemistry, and also of the interactions of oxidans can detoxify and efflux this metal by a this metal with bacteria and other biological process in which its polyphosphate bodies are components that are present in these environ- involved. ments. The main scientific issues concerning the chemical forms or speciation of contaminants such Introduction as uranium depend ultimately on the molecular- The intensive uranium mining and milling that scale structure. Basic understanding at this level is was carried out over a period of 40 years (up to essential for management and removal of the en- 1990) have caused significant pollution with vironmental contaminants. Information at the uranium and other toxic metals in the Southeast molecular level on complex environmental systems, including complex structures such as Key words EXAFS, time-resolved laser-induced fluorescence bacteria/uranium, for instance, can be obtained spectroscopy (TRLFS), transmission electron microscopy (TEM)/ from powerful molecular-scale analyses that energy-dispersive X-ray (EDX) utilize electromagnetic radiation, such as X-ray or Abbreviations used EDX, energy-dispersive X-ray, FT. Founer IR radiation. The main objective of the present transform, TEM, transmission electron microscopy, TRLFS, time- work was to conduct studies at the atomic and resolved laser-induced fluorescence spectroscopy ‘To whom correspondence should be addressed (e-mail molecular scales on uranium(V1) complexes m rnerroun@fz-rossendorfde) formed by three eco-types of Acidithiobacillus

669 0 2002 Biochemical Society Biochemical Society Transactions (2002) Volume 30, part 4

ferrooxidans. For this reason, a combination of uranium. They are identified as axial oxygens EXAFS, time-resolved laser-induced fluorescence (OaJ of the linear uranyl group. Backscatters from spectroscopy (TRLFS), transmission electron oxygen atoms lying in the equatorial plane of the microscopy (TEM) and energy-dispersive X-ray uranyl ion appear as a single broad peak at 1.8 A. (EDX) analysis was used. Three different eco- The average distance between uranium and the types of A.ferrooxidans have been recovered from equatorial oxygen atoms (Oe,) is 2.36 f 0.02 A, different sites and depths of two uranium mining with a co-ordination number of 4-5. The U-O,, waste piles [3-51. These eco-types differ in bond distance of 2.36k0.02 A is similar to the their capability to accumulate and tolerate ura- value of 2.38 A measured by EXAFS for Bacillus nium [6]. subtilis uranium complexes [8], but significantly greater than the U-0,, bond distances (2.28- EXAFS 2.29 A) found in the complexes of vegetative cells and spores of uranium-treated Bacillus cereus and X-ray absorption spectroscopy can be used for Bacillus sphaericus [9]. As is evident from Figure 1, characterization of the valence and the local struc- a weak peak appears at approx. i-AR = 3 A (not ture (< 10 A) of elements in a solid matrix [7]. By R corrected from phase shift) in the FT, which was the use of EXAFS, molecular structural infor- modelled to the interaction of uranium with mation, such as co-ordination number, bond phosphorus, giving a distance of 3.63 As shown distance and oxidation state of elements, can be A. in Figure 1, no structural differences were ob- determined. served between the uranium complexes formed by Uranium L,,,-edge EXAFS spectra of the the three eco-types of A.ferrooxidans. Therefore uranium species formed by the A. ferrooxidans we suggest that the same groups are implicated cells at pH 2 and their corresponding Fourier in the complexation of uranium by the three transforms (FTs) are shown in Figure 1. It is ferrooxidans eco-types. established that when X(k)is Fourier-transformed A. over a finite k range, the result is a radial structure TRLFS function exhibiting a series of peaks whose pos- TRLFS is a very sensitive method which can give itions and amplitudes are related to the interatomic insight into complexation reactions, even with distances and the number of atoms in the different poorly defined polyfunctional ligands such as co-ordination shells. living bacterial cells. Using this method, different The most prominent peak in all samples spectroscopic characteristics (i.e. energy and shape occurs at 1.3 A, and arises from the backscattering of the emission bands, and the lifetime of fluores- caused by the oxygen atoms that are nearest to the cence decay) can be used to identify a particular uranyl species. Figure I The results of the present work demonstrate Uranium L,,,-edge k3-weighted EXAFS spectra (left) that the uranium complexes built by the three eco- and corresponding FTs (right) of the uranium com- types of A. ferrooxidans have different lifetimes. plexes formed by the three eco-types of A. ferrooxidans Eco-type I1 forms very strong uranium complexes (higher formation constants) with longer lifetimes

~ data Ithan those of eco-types I and 111. The amount of -.....fit uranium accumulated by the three bacterial eco- types increases with the increasing lifetimes of the corresponding uranium complexes. The lifetimes of the uranyl ion species bound to eco-types I, I11 and I1 of A. ferrooxidans were measured to be 15.7, 17.1 and 29.8 times greater respectively than those of the aqueous uranyl ion. Correspondingly, the amounts of uranium bound by eco-types I, I11 and I1 were 21.54,30 and 44.62 mg/g dry biomass respectively,

TEM and EDX microanalysis 1.1'1 4 0 12 02468 TEM observations of A.ferrooxidans cells exposed k [ A-'] R + A [A] to uranium solution allowed detection of electron-

0 2002 Biochemical Society 670 Biometals 2002: Third International Biometals Symposium

dense accumulates essentially within the extra- phosphates is a passive protective mechanism cellular polysaccharides (results not shown), against heavy metals. It is not induced by the which consist mainly of sugars and lipids, in presence of the metals, but nonetheless improves addition to small amounts of nitrogen, phosphorus cellular tolerance to them [14]. This mechanism ( z 1 "") and non-esterified fatty acids [lo]. How- involves the transport and efflux of short uranium- ever, by EDX analysis the presence of uranyl phosphorus complexes from the bacterial cyto- phosphate in the electron-dense accumulates was plasm to outside the cells. It has been reported demonstrated. This result is in agreement with that uranyl ions at high concentrations negatively our EXAFS data and with the results published by affect the growth of A.ferrooxidans, based on iron Gonzalez-Mufioz et al. [ll], who found that the oxidation and carbon dioxide fixation [15]. On the uranium accumulated by another bacterial species, basis of our results, we suggest that A.ferrooxidans Myxoccocus xanthus, is located mainly within cells can use polyphosphates to detoxify and extracellular polysaccharides as phosphate com- remove uranium when this metal enters the cells pounds. The possible origin of the phosphorous and is present at toxic levels. residues in the extracellular polysaccharide fraction of A. ferrooxidans treated with uranium is Summary discussed below. EXAFS and TRLFS measurements, as well as Electron-dense granules were also observed TEM and EDX analysis, of the uranium(V1) in the bacterial cell cytoplasm (Figure 2). In some complexes formed by the three eco-types of cases such granules were associated with the A. ferrooxidans demonstrate that phosphorus bacterial cell membrane system. EDX analysis of residues of this bacterium are involved in the these granules showed that they contain high interaction with uranium. The molecular-level amounts of uranium and phosphorous (results not structural information obtained using these tech- shown). We suggest that these granules corre- niques, particularly EXAFS and TRLFS, is spond to polyphosphate bodies. Inorganic poly- essential for understanding the interaction of phosphates are linear polymers of Pi residues uranium with natural A.ferrooxidans populations linked by phosphoanhydride bonds. Chain lengths in the environment. Such information may be use- may vary from three to 1000 Pi residues, depend- ful for modelling and quantifying the microbial ing on the organism, its growth, and other physio- impact on the fate of uranium and other actinides logical conditions [12]. Polyphosphates can be a in Nature. substitute for ATP, sugar and adenylate kinases, an energy source, and a chelator of bivalent metals References ions [13]. It has been suggested that the intra- I Bemhard, G., Geipel, G., Brendler, V. and Nitsche, H. cellular chelation of heavy metals by poly- (I 996) Radiochim. Acta 74, 87-9 I 2 John,S. G., Ruggiero, C.E., Hersman, L. E., Tung, C.-S. and Neu, M. P. (200 I) Environ. Sci. Technol. 35, 2942-2948 Figure 2 3 Selenska-Pobell, S.,Flemming, K. and Radeva, G. (2000) Transmission electron micrograph of a thin section of Report FZR 285,52 A. ferrooxidans cells challenged with uranium 4 Selenska-Pobell, S., Flemming, K., Kampf, G., Radeva, G. and Satchanska, G. (200 I) Antonie van Leewenhuek 79, The accumulated metal is located inside the intracellular polyphos- 149-1 6 I phate body fraction (a). Magnification 38000 x . 5 Selenska-Pobell, S. (2002) in Interaction of Microorganisms with Radionuclides (Keith-Roach, M. and Livens, F., eds), pp. 225-253, Elsevier Sciences, Oxford 6 Merroun. M. L. and Selenska-Pobell. S. (200 I) Biometals 14. 171-179 7 Mansour, A. N. and Melendres, C. A. (I998) J. Phys. Chem. A 102, 65 8 Kelly, S. D., Boyanov, M. I,, Bunker, B. A,, Fein, J. B.. Fowle, D. A,, Yee, N. and Kemmer. K. M. (2001) J. Synchrotron Radiat. 8, 946-948 9 Hennig, C.,Panak, P. J,, Reich, T., Rossberg, A,. RaK J., Selenska-Pobell, S., Matz, W., Bucher, J. J., Bemhard, G. and Nitsche, H. (2001) Radiochim. Ada 89, 625-63 I 10 Gehrke, T., Telegdi, J,, Thierry, D. and Sand, W. (I 998) Appl. Environ. Microbiol. 64, 2743-2747 II Gonzalez-Murioz, M. T., Merroun, M. L., Ben Omar, N. and Arias, J. M. (1997) Int. Biodeterior. Biodegrad. 40, 107-1 14

67 I 0 2002 Biochemical Society Progress in Solid State Chemistry 30 (2002) 1–101 www.elsevier.nl/locate/pssc

Future directions in solid state chemistry: report of the NSF-sponsored workshop Robert J. Cava a,∗, Francis J. DiSalvo b, Louis E. Brus c, Kim R. Dunbar d, Christopher B. Gorman e, Sossina M. Haile f, Leonard V. Interrante g, Janice L. Musfeldt h, Alexandra Navrotsky i, Ralph G. Nuzzo j, Warren E. Pickett k, Angus P. Wilkinson l, Channing Ahn m, James W. Allen n, Peter C. Burns o, Gerdrand Ceder p, Christopher E.D. Chidsey q, William Clegg r, Eugenio Coronado s, Hongjie Dai t, Michael W. Deem u, Bruce S. Dunn v, Giulia Galli w, Allan J. Jacobson x, Mercouri Kanatzidis y, Wenbin Lin z, Arumugam Manthiram aa, Milan Mrksich bb, David J. Norris cc, Arthur J. Nozik dd, Xiaogang Peng ee, Claudia Rawn ff, Debra Rolison gg, David J. Singh hh, Brian H. Toby ii, Sarah Tolbert jj, Ulrich B. Wiesner kk, Patrick M. Woodward ll, Peidong Yang mm

∗ Corresponding author. Tel.: +1-609-258-0016; fax: +1-609-258-6746. E-mail addresses: [email protected] (R.J. Cava); [email protected] (F.J. DiSalvo); brus@- chem.columbia.edu (L. Brus); [email protected] (K.R. Dunbar); [email protected] (C. Gorman); [email protected] (S.M. Haile); [email protected] (L.V. Interrante); [email protected] (J. Musfeldt); [email protected] (A. Navrotsky); [email protected] (R. Nuzzo); pickett@physics.ucdavis.edu (W.E. Pickett); [email protected] (A.P. Wilkinson); cca@cal- tech.edu (C. Ahn); [email protected] (J.W. Allen); [email protected] (P.C. Burns); [email protected] (G. Ceder); [email protected] (C. Chidsey); [email protected] (W. Clegg); [email protected] (E. Coronado); [email protected] (H. Dai); [email protected] (M.W. Deem); [email protected] (B.S. Dunn); [email protected] (G. Galli); [email protected] (A.J. Jacobson); [email protected] (M. Kanatzidis); [email protected] (W. Lin); [email protected] (A. Manthiram); [email protected] icago.edu (M. Mrksich); [email protected] (D. Norris); [email protected] (A. Nozik); [email protected] (X. Peng); [email protected] (C. Rawn); [email protected] (D. Rolison); singh@- dave.nrl.navy.mil (D. Singh); [email protected] (B. Toby); [email protected] (S. Tolbert); [email protected] (U.B. Wiesner); [email protected] (P. Woodward); [email protected] (P. Yang).

0079-6786/03/$ - see front matter  2002 Published by Elsevier Science Ltd. doi:10.1016/S0079-6786(02)00010-9 R.J. Cava et al. / Progress in Solid State Chemistry 30 (2002) 1–101 17

Fig. 4. Development of the wo¬lsendorﬁte sheet anion topology as a stacking sequence of chains. This approach demonstrates that the wo¬lsendorﬁte anion topology is composed of modules of the simpler a-

U3O8 and b-U3O8 anion topologies (Reprinted by permission of the Mineralogical Society of America, USA). for uranyl polyhedra should be established, perhaps using quantum mechanical simulations, although this approach is especially demanding because of relativistic effects involving core electrons on the U atoms. Such potentials may then be used for static energy minimizations and molecular dynamic simulations of uranyl minerals, including crystal surfaces. The relationships between the structural units and interstitial components also need to be modeled, and the role of interstitial components as structure directing agents should be addressed. An understanding of the energetics of complex sheets of uranyl polyhedra is needed, and may provide the basis for development of a method to predict thermodynamic properties of uranyl minerals on the basis of structural connectivity. X-ray absorption spectroscopy (XAS) permits the probing of oxidation states and local coordination environments about U in minerals [34]. Further development of this approach is warranted in the area of analysis of extended X-ray absorption ﬁne structure (EXAFS) spectra for U6+. The situation is complex owing to the occurrence of U6+ in three different coordination environments, and the sharing of uranyl poly- hedral elements with a diverse range of oxyanions, including silicate, phosphate, arsenate, carbonate, vanadate, and sulfate. The spectra of U6+ associated with mineral and bacteria surfaces can also be difﬁcult to interpret [35,36]. Studies of the EXAFS spectra of several model compounds containing different coordination environments about U6+, and multiple U valence states, will substantially improve the applicability of this technique to complex natural specimens. Recent advances in characterization of the structures and chemistries of uranyl minerals provide the foundation for studies of their thermodynamics. However, there 100 R.J. Cava et al. / Progress in Solid State Chemistry 30 (2002) 1Ð101

[34] Thompson HA, Brown GE, Parks GA. XAFS spectroscopic study of uranyl coordination in solids and aqueous solution. Am. Mineral 1997;82:483Ð96. [35] Kelly SD, Boyanov MI, Fein JB, Fowle DA, Yee N, and Kemner KM. XAFS determination of U- bacterial cell wall interaction at low pH, Geochim. Cosmochim. Acta, submitted for publication, 2001. [36] Kelly SD, Boyanov MI, Bunker BA, Fein JB, Fowle DA, Yee N, Kemner KM. XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J. Synchrotron. Radiation 2001;8:946Ð8. [37] Weiss A. Angew. Chem. Int. Ed. Engl 1981;20:850. [38] Sarikaya M, Aksay IA. In: Biomimetics: design and processing of materials. Woodbury, NY: AIP Press; 1995. [39] Mann S, Ozin GA. Nature 1996;382:313. [40] Volkmer D. Chem. unserer Zeit 1999;33:6. [41] Marentette JM, Norwig J, Sto¬ckelmann E, Meyer WH, Wegner G. Adv. Mater 1997;9:647. [42] Templin M, Franck A, Du Chesne A, Leist H, Zhang Y, Ulrich R, Scha¬dler V, Wiesner U. Science 1997;278:1795. [43] Ulrich R, Du Chesne A, Templin M, Wiesner U. Adv. Mater 1999;11:141. [44] Go¬ltner CG, Antonietti M. Adv. Mater 1997;9:431. [45] Zhao D, Feng J, Huo Q, Melosh N, Fredrickson GH, Chmelka BF, Stucky GD. Science 1998;120:6024. [46] Simon PF, Ulrich R, Spiess HW, Wiesner U. Chem. Mater 2001;13:3464. [47] Schmidt-Rohr K, Spiess HW. Multidimensional solid-state NMR and polymers. London: Academic Press, 1994. [48] Soong RK, Bachand GD, Neves HP, Olkhovets AG, Craighead HG, Montemagno CD. Powering an inorganic nanodevice with a biomolecular motor. Science 2000;290:1555Ð8. [49] Lee I, Lee JW, Greenbaum E. Biomolecular electronics: vectorial arrays of photosynthetic reaction centers. Phys. Rev. Lett 1997;79(17):3294Ð7. [50] The US Geological Survey predicts that 2659 billion barrels of oil remain, including both known and yet-to-be-discovered sources, USGS World Petroleum Assessment 2000 in USGS Survey Digital Data Series 60; the US Dept of Energy predicts worldwide oil consumption will be 120 million barrels a day by 2020, US Energy Information Administration, International Energy Outlook 2001, p. 27.

[51] Three-quarters of anthropogenic CO2 results from fossil fuel combustion and the remainder from changes in land use patterns, US Energy Information Administration, International Energy Outlook 2001, p. 159.

[52] Marland G, Boden T. The increasing concentration of atmospheric CO2: How much, when, and why? Erice International Seminars on Planetary Emergencies, 26th Session, Erice, Sicily, Italy, 19Ð 24 August 2001. [53] Saracco G, Neomagus HWJP, Versteeg GF, van Swaaij WPM. Chem. Eng. Sci 1999;54:1997Ð2017. [54] Hirschenhofer JH, Stauffer DB, Engelman RR, Klett MG. Fuel cell handbook, 4th Edition, Parsons Corp., for US Dept. of Energy. Report no. DOE/FETC-99/1076, 1998. [55] Larmine J, Andrews D. Fuel cell systems explained. Chichester: John Wiley and Sons Ltd, 2000. [56] Sandstede G, Wurster R. Water electrolysis and solar hydrogen demonstration projects. Modern Aspects of Electrochemistry 1995;27:411Ð515. [57] Vining CB. Nature 2001;13:577Ð8. [58] Nolas GS, Sharp J, Goldsmid HJ. Thermoelectrics: basic principles and new materials developments. Berlin: Springer-Verlag, 2001. [59] Venkatasubramanian R, Siivola E, Colpitts T, O’Quinn B. Nature 2001;413:597Ð602. [60] Besenhard JO, editor. Handbook of battery materials. Weinheim: Wiley-VCH; 1999. [61] See the following websites: (a) Tokin America, activated carbon particles [http://www.tokin.com/Tokin—America—Products/43/p43/p43.html]; (b) Tavrima Canada Ltd. [http://www.tavrima.com/]; (c) XX Pty Ltd. and CSIRO, activated carbon [http://www.cap- xx.com/about/default.html]; and (d) Paul Scherer Institute, ABB and Leclanche« SA, glassy carbon [http://ecl.web.psi.ch/supercap/kti1.htm]. 1 An Overview of 11 Synchrotron Radiation Applications to Low Temperature Geochemistry and Environmental Science Gordon E. Brown, Jr.1,2 and Neil C. Sturchio3 1Department of Geological and Environmental Sciences 2Stanford Synchrotron Radiation Laboratory Stanford University Stanford, California, 94305-2115, U.S.A.

3Department of Earth and Environmental Sciences University of Illinois at Chicago 845 West Taylor Street Chicago, Illinois, 60607-7059, U.S.A.

INTRODUCTION The availability of synchrotron radiation (SR) to the scientific community has literally revolutionized the way X-ray science is done in many disciplines, including low temperature geochemistry and environmental science. The key reason is that SR provides continuum vacuum ultraviolet (VUV) and X-ray radiation five to ten orders of magnitude brighter than that from standard sealed or rotating anode X-ray tubes (Winick 1987; Altarelli et al. 1998). Although SR was first observed indirectly by John Blewitt in 1945 (Blewitt 1946) and directly by Floyd Haber in 1946 at the General Electric 100-MeV Betatron in Schenectady, NY (see Elder et al. 1947; Baldwin 1975), it took 10 to 15 years before the first systematic applications of SR, which involved spectroscopic studies of the VUV absorption of selected elements (Tomboulian and Hartman 1956) using the 300- MeV synchrotron at Cornell University and of rare gases (Madden and Codling 1963) using the National Bureau of Standards SURF I synchrotron. As of September 2002, there are about 75 storage ring-based SR sources in operation, in construction, funded, or in advanced planning in 23 countries, with 10 fully dedicated SR storage ring facilities in the U.S.. A listing of these sources can be obtained at the following web site: http://www- ssrl.slac.stanford.edu/sr_sources.html. The first SR experiments relevant to low temperature geochemistry and environmental science, although not performed on earth or environmental materials, were X-ray absorption fine structure (XAFS) spectroscopy measurements on amorphous and crystalline germanium oxide conducted on the SPEAR storage ring at the Stanford Synchrotron Radiation Project in 1971 by Dale Sayers, Farrel Lytle, and Edward Stern (Sayers et al. 1971). Prior to the availability of SR in the hard X-ray energy range (> 5 keV), XAFS spectroscopy measurements were impractical because of the high X-ray flux required and the need for a continuously tunable range of X-ray energies extending up to 1000 eV above the absorption threshold of the element of interest. In the 30 years since these SR-based XAFS measurements, this method has been applied to materials ranging from metalloproteins (e.g., Hasnain and Hodgson, 1999) and catalysts (e.g., Lytle et al. 1980; Sinfelt et al. 1984; Maire et al. 1986; Koningsberger et al. 1999) to silicate liquids at high temperatures (e.g., Brown et al. 1995a), cation complexes in aqueous and hydrothermal solutions (e.g., Fontaine et al. 1978; Hoffman et al. 2001), and complex environmental samples containing heavy metal and metalloid contaminants and pollutants

1529-6466/00/0049-0001$15.00 Overview of Synchrotron Radiation Applications 13 studies by Teng et al. (2001) have provided important details about feldspar dissolution as a function of pH in the presence and absence of oxalate. Further mechanistic information on the effect of organic ligands on mineral dissolution and growth was provided by the X-ray reflectivity study of the adsorption of the growth inhibitor HEDP on barite (Fenter et al. 2001a). Future studies of mineral dissolution and growth reactions using soft X-ray XAFS spectroscopy at the C K-edge in the presence of water vapor may provide some of this information as well (see Myneni, this volume). Natural organic matter has long been thought to form coatings on mineral surfaces that can result in significant changes in mineral surface charge and reactivity to metal ions in solution (e.g., Neihof and Loeb 1972, 1974; Hunter and Liss 1979; Hunter 1980; Balistrieri et al. 1981; Davis and Gloor 1981; Tipping 1981; Tipping and Cooke 1982; Davis 1984; Mayer 1994a,b; Au et al. 1999). More recent evidence, however, suggests that such coatings comprise only a small fraction (< 15%) of an effective monolayer on mineral particles in ocean and estuarine waters (Ransom et al. 1997; Mayer 1999) and in soils (Mayer and Xing 2001). Nonetheless, organic coatings when present in discrete regions on mineral surfaces may change the surface charge and reactivity of mineral surfaces and may affect sorption reactions of aqueous metal ions. X-ray standing wave, X-ray reflectivity, and X-ray spectromicroscopy methods can be used to investigate these phenomena. For example, Fenter and Sturchio (1999) demonstrated how X-ray reflectivity could be used to provide a detailed characterization of the growth and structure of stearate monolayers on calcite. Biota can have a major effect on the speciation and sorption of contaminant and pollutant species Microbial organisms and fungi such as lichens can enhance dissolution of mineral surfaces (Banfield and Hamers 1997; Banfield et al. 1999), thus changing the types and densities of surface functional groups, which in turn can affect the reactivity of the surface. They can also generate nanoparticulate precipitates (Banfield and Zhang 2001). Important questions are whether microbial biofilm coatings on mineral surfaces block reactive sites or provide alternative functional groups that can bind metal ions. X-ray standing wave and XAFS methods can uniquely address these questions (e.g., Templeton et al. 1999, 2001; Hennig et al. 2001; Kelly et al. 2001, 2002; Webb et al. 2001; Panak et al. 2002). Molecular-scale mechanisms of bioremediation and phytoremediation of environmental toxins are typically not known Many microorganisms and plants are capable of transforming toxic chemical species into less toxic forms (e.g., Lytle et al. 1996, 1998; Hunter et al. 1997). Some plants are particularly useful for remediation of contaminants in soils and natural waters because they hyperaccumulate specific toxins (e.g., Van der Lelie et al. 2001; Fuhrmann et al. 2002). In most cases, however, the molecular-scale mechanisms of these transformations or of hyperaccumulation are not known. This is a fertile research field for geochemists and mineralogists that will require multidisciplinary studies and will benefit from SR- based microspectroscopy methods. Scaling of molecular-level observations to mesoscopic and macroscopic phenomena: the “Holy Grail” of low temperature geochemistry and environmental science It would be extremely useful if laboratory measurements at a variety of spatial scales, including the nanometer scale, could be accurately extrapolated to the field scale so that the consequences of a field-scale pollution event could be assessed and predicted in quantitative detail as a function of time and location. Conversely, it would be useful if Overview of Synchrotron Radiation Applications 21 (this volume), and Waychunas (this volume). Examples of XAFS studies on materials of relevance to low temperature geochemistry and environmental science include the following: (1) sorption of aqueous metal ions on high surface area minerals (Hayes et al. 1987; see Table 1 in the Appendix for many additional references) (2) interaction of aqueous metal ions with natural organic matter and organic acids (Allen et al. 1996a; Hesterberg et al. 1997b, 2001; Sarret et al. 1997; Xia et al. 1997a,b, 1999; Denecke et al. 1998a,b, 2002; Rossberg et al. 2000; Alcacio et al. 2001; Liu et al. 2001; Pompe et al. 2001; Schmeide et al. 2001) (see Table 2 in the Appendix) (3) interaction of aqueous metal ions with microbial and fungal cell walls (Sarret et al. 1998a, 1999a, 2001; Templeton et al. 1999, 2002a; Soderholm et al. 2000; Hennig et al. 2001; Kelly et al. 2001, 2002; Webb et al. 2001; Panak et al. 2002) (4) determination of the structures and composition of metal ion complexes in aqueous solutions (Fontaine et al. 1978; Apted et al. 1985; Mosselmans et al. 1999; Frenkel et al. 2001; Sandstrom et al. 2001), hydrothermal solutions (Fulton et al. 1996; Seward et al. 1996; Mayanovic et al. 1999), and natural fluid inclusions (Anderson et al. 1995, 1998a; Mayanovic et al. 1996, 1997) (see section on water and metal ions in aqueous solutions for additional examples) (5) structural studies of iron oxides and hydroxides (Waychunas et al. 1983, 1986; Combes et al. 1986 1989a,b, 1990; Manceau and Combes 1988; Manceau et al. 1990b; Hazemann et al. 1992a,b; Manceau and Drits 1993; Wilke et al. 2001) (6) structural studies of manganese oxides and hydroxides (Manceau and Combes 1988; Manceau et al. 1992a,b,c; Drits et al. 1997a; Manceau and Gates 1997; Silvester et al. 1997; Drits et al. 1998; Lanson et al. 2000) (7) characterization of Mn oxidation reaction products at sediment/water interfaces (Wehrli et al. 1995) (8) selenium transformations in soils and sediments (Tokunaga et al. 1994; Pickering et al. 1995; Tokunaga et al. 1996, 1997) (9) characterization of “neoformed” phyllosilicates (Dähn et al. 2002a) (10) XAFS studies of the crystal chemistry of kaolinites and low temperature iron and manganese hydroxides (Muller et al. 1995) (11) XAFS studies of the crystal chemistry of iron silicate scale deposits in the Salton Sea geothermal field (Manceau et al. 1995) (12) XAFS studies of trace element environments in natural and synthetic goethites (Manceau et al. 2000b) (13) nucleation and growth of iron oxides in the presence of phosphate (Rose et al. 1996) (14) formation of ferric hydroxide-nitrate oligomers in aqueous solutions (Rose et al. 1997) (15) mechanism of formation of iron oxyhydroxide-chloride polymers in solution (Bottero et al. 1994) (16) oxidation-reduction mechanisms for iron in smectites (Drits and Manceau 2000; Manceau et al. 2000c,d; Gates et al. 2002) (17) mechanism of Cr(III) and As(III) oxidation by hydrous manganese oxides (Fendorf 1995; Silvester et al. 1995; Tournassat et al. 2002) 46 Brown & Sturchio section on “comments on the structure and properties of water at solid/aqueous solution interfaces”). Early XAFS studies of metal ion sorption on metal oxides (e.g., Co(II) sorption at the γ-Al2O3/solution interface: Chisholm-Brause et al. 1989a, 1990a, 1991) also showed that divalent transition metal ions tend to form isolated mononuclear, inner-sphere surface complexes at the lowest metal ion sorption densities (< 10% of an effective monolayer), multinuclear complexes with increasing sorption density, and precipitates at sorption densities approaching or above effective monolayer coverages. As shown by TEM examination of the high sorption density samples, these precipitates may or may not be associated with the mineral surface. Another discovery that merits attention is the finding that mineral dissolution in aquatic systems may be followed by precipitation of a new phase that incorporates ions derived from both the dissolution process and the aqueous solution (e.g., Co(II), Ni(II), and Zn(II) ions sorbed on alumina or phyllosilicates: d’Espinose de la Caillerie et al. 1995a,b; 1997; Scheidegger et al. 1997; Towle et al. 1997; Thompson et al. 1999a,b, 2000; Trainor et al. 1999). This process results in the formation of Co, Ni, and Zn layered double hydroxides, which have a much larger capacity for sequestering contaminant ions than a static mineral surface with a limited number of reactive sites per unit surface area. Other EXAFS studies of sorption reaction products found that some contaminant or trace metal ions are incorporated in the structure of the solid phases on which they adsorb by a “neoformation” process. Such studies of divalent transition metal ion sorption on clay minerals found that the adions attach to edge sites of the clay minerals (e.g., Co(II) and Zn(II) neoformation on hectorite: Schlegel et al. 1999a,b, 2001a,b). Another important recent finding is that some of the so-called “indifferent electrolytes” that are used in sorption studies, including nitrate and perchlorate, are not necessarily indifferent. EXAFS evidence suggests that they form ternary complexes with cations such as U(VI) sorbed on hematite (Bargar et al. 2002c) and Pb(II) sorbed on ferrihydrite (Trevedi et al. 2002). Each of these findings is beginning to affect surface complexation modeling of macroscopic uptake of cations and anions from aqueous solutions onto mineral surfaces, where chemical reactions describing surface sites and sorption products must be assumed in the absence of molecular-level information on reactants and products. It is now known that different chemical reactions with different sorption products may fit macroscopic uptake data equally well (e.g., Hayes and Katz 1996; Brown et al. 1999c), confirming the suggestion of Sposito (1986) that microscopic information cannot be extracted from macroscopic observations. EXAFS spectroscopy is also being applied to the interaction of cations and anions with microbial cell walls, as described in the next section, and is showing that assumptions made about reactive functional groups in bacterial cell walls are not always consistent with direct EXAFS spectroscopic observation. EXAFS studies of metal ion coordination sites in bacterial and fungal cell walls Synchrotron-based experimental approaches are well suited for studying the interaction of cations and anions with bacterial cells, although only a few systems have been studied to date. EXAFS spectroscopy can be used in some cases to determine which functional groups (e.g., carboxylic, phosphoryl, phenolic hydroxyl, thiol) within a biofilm or bacterial or fungal cell wall are reactive to trace metals (e.g., Sarret et al. 1998a, 1999a; Kelly et al. 2001, 2002; Boyanov et al. 2002; Webb et al. 2001). For example, Sarret et al. (1999a) used EXAFS spectroscopy and solution chemistry experiments to show that Pb(II) reacts with carboxyl and phosphoryl sites in the cell wall of the fungus Penicillium chrysogenum, with weak binding of Pb(II) to carboxyl groups accounting for 95% of Pb(II) uptake and strong binding of Pb(II) to less abundant phosphoryl groups accounting for 5% of the uptake. In EXAFS studies of the sorption products of Cd(II) and U(VI) on Bacillus subtilis, Kelly et al. (2002) and Boyanov et al. (2002) found that U(VI) and Cd(II), respectively, bind to phosphoryl groups at low pH and that U(VI) and Cd(II) bind Overview of Synchrotron Radiation Applications 87

Kelly SD, Boyanov MI, Bunker BA, Fein JB, Fowle DA, Yee N, Kemner KM (2001) XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J Synchrotron Rad 8:946-948 Kelly SD, Kemner KM, Fein JB, Fowle DA, Boyanov MI, Bunker BA, Yee N (2002) X-ray-absorption fine-structure determination of pH-dependent U-bacterial cell wall interactions. Geochim Cosmochim Acta 66:(in press) Kemner KM, Hunter DB, Bertsch PM, Kirkland JP, Elam WT (1997a) Determination of site specific binding environments of surface sorbed cesium on clay minerals by Cs-EXAFS. J de Physique IV 7(C2, X-Ray Absorption Fine Structure, Vol. 2):777-779 Kemner KM, Hunter DB, Gall EJ, Bertsch PM, Kirkland JP, Elam WT (1997b) Molecular characterization of Cr phases in contaminated soils by Cr and Fe EXAFS: a tool for evaluating chemical remediation strategies. J Phys IV 7 (C2, X-Ray Absorption Fine Structure, Vol. 2): 811-812 Kemner KM, Lai B, Maser J, Schneegurt MA, Cai Z, Ilinski P, Kulpa CF, Legnini DG, Nealson KH, Pratt ST, Rodrigues W, Lee-Tischler M, Yun W (2000) Use of the high-energy X-ray microprobe at the Advanced Photon Source to investigate the interactions between metals and bacteria. AIP Conf Proc 507 (X-ray Microscopy): 319-322 Kemner KM, Yun W, Cai Z, Lai B, Lee H-R, Maser J, Legnini DG, Rodriques W, Jastrow JD, Miller RM, Pratt ST, Schneegurt MA, Kulpa CF Jr (1999) Using zone plates for X-ray microimaging and microspectroscopy in environmental samples. J Synchrotron Rad 6:639-641 Kendelewicz T, Doyle CS, Carrier X, Brown GE Jr (1999) Reaction of water with clean surfaces of MnO(100). Surf Rev Lett 6:1255-1263 Kendelewicz T, Liu P, Brown GE Jr, Nelson EJ (1998a) Atomic geometry of the PbS(100) surface. Surf Sci 395:229-238 Kendelewicz T, Liu P, Brown GE Jr, Nelson EJ (1998b) Interaction of sodium overlayers with cleaved surfaces of galena (PbS(100)): Evidence for exchange reactions. Surf Sci 411:10-21 Kendelewicz T, Liu P, Doyle CS, Brown GE Jr (2000a) Spectroscopic study of the reaction of Cr(VI)aq with Fe3O4 (111) surfaces. Surf Sci 469:144-163 Kendelewicz T, Liu P, Doyle CS, Brown GE Jr, Nelson EJ, Chambers SA (2000b) Reaction of water with the (100) and (111) surfaces of Fe3O4. Surf Sci 453:32-46 Kendelewicz T, Liu P, Labiosa WB, Brown GE Jr (1995) Surface EXAFS and X-ray standing wave study of the cleaved CaO(100) surface. Physica B 208 & 209:441-442 Kersting AB, Efurd DW, Finnegan DL, Rokop DJ, Smith DK, Thompson JL (1999) Migration of plutonium in ground water at the Nevada Test Site. Nature 397:56-59 Kieffer F (1992) Metals as essential trace elements for plants, animals, and humans. In: Merian E (ed) Metals and Their Compounds in the Environment. VCH, Weinheim, Germany, p 481-489 Kikuma J, Tonner BP (1996) XANES spectra of a variety of widely used organic polymers at the C K- edge. J Electron Spectrosc Relat Phenom 82:53-60 Kim CS, Bloom NS, Rytuba JJ, Brown GE Jr (2002a) Mercury speciation by extended X-ray absorption fine structure (EXAFS) spectroscopy and sequential chemical extractions: An intercomparison of speciation methods. Environ Sci Technol (submitted) Kim CS, Brown GE Jr, Rytuba JJ (2000) Characterization and speciation of mercury-bearing mine wastes using X-ray absorption spectroscopy (XAS). Science of the Total Environment 261:157-168 Kim CS, Rytuba JJ, Brown GE Jr (1999) Utility of EXAFS in characterization and speciation of mercury- bearing mine wastes. J Synchrotron Rad 6:648-650 Kim CS, Rytuba JJ, Brown GE Jr (2002b) EXAFS study of mercury(II) sorption on Fe- and Al- (hydr)oxides: II. Effects of chloride and sulfate. J Colloid Interface Sci (submitted) Kim CS, Rytuba JJ, Brown GE Jr (2002c) EXAFS study of mercury(II) sorption on Fe- and Al- (hydr)oxides: I. Effects of pH. J Colloid Interface Sci (submitted) Kim CS, Rytuba JJ, Brown GE Jr (2002d) Geological and anthropogenic factors influencing mercury speciation in mine wastes. Appl Geochem (submitted) Kneebone PE, O’Day PA, Jones N, Hering JG (2002) Deposition and fate of arsenic in iron- and arsenic- enriched reservoir sediments. Environ Sci Technol 36:381-386 Kobayashi K, Kawata H, Mori K (1998) Site specification on normal and magnetic XANES of ferrimagnetic Fe3O4 by means of resonant magnetic Bragg scattering. J Synchrotron Rad 5:972-975 Kolker A, Huggins FE, Palmer CA, Shah N, Crowley SS, Huffman GP, Finkelman RB (2000) Mode of occurrence of arsenic in four US coals. Fuel Process Technol 63:167-178 Koningsberger D, Mojet B, Miller J, Ramaker D (1999) XAFS spectroscopy in catalysis research: AXAFS and shape resonances. J Synchrotron Rad 6:135-141 Kortright J, Warburton W, Bienenstock A (1983) Anomalous X-ray scattering and its relationship to EXAFS. Springer Ser Chem Phys, Vol. 27, Springer-Verlag, Berlin, p 362-372

...f...·.?;.·~.·.·.···· I SORPTION OF TRACE ELEMENTS ON MINERAL SURFACES 1031

TADLE 11. Studies of the Uptake of Selected Metal Cations on Bacterial, Algal, and Fungal Surfaces

CUiion Bllcterillm, ruga, ii. fungus References

Ag(I) Bacillus ,mbtili.~ Beveridge Ilnd Murray, 1976; Mullen eL al., 1989 Bacillll.~ cereus Mullen el Ill., 1989 Escherichia coli Mullen cln!.. 1989 Pse:u1mmmru fU!rllgina.m Mullen e! ul .• 1989 Cd(ll) Sargtu.mmjluitaru Fourest und Volesky, 1996; Davis el ul., 2000 Sarga.'t.mmjIuitam Sehiewer und Volesky, 1997; Figueira el ul., 2000 l'tl:uchena alga Crist eL at., 1999 Bacililu .mbtili., Beveridge nml Murray, 1980; Fein eL ut, 1997 Dauehnev and Fein. 199B; Fowle and Fein. 1999 B(Jcillu.~ lichcniJonni.f Bev;ridg~ eL ul., 1982; Dllughney nnd Fein. 1998 Fowle llnd Fein, 1999 Bacillu.~ .mutilis Kelly eL u1., 2001 E,(cherichiu coli Yee and Fein, 2001 Pscudoawna.( acmginosa Yee and Ftlin, 2001 Bacillu.( megulUrium Yee und Fein, 2001 Streptocuccu.(jaecalis Yee und Fein, 2001 Staph)'lococclu aurelu Yee und Fein. 2001 Spuro.mrcina 1lTl!(le Yee und Fein, 2001 Bucillu.t cercu.( Mullen et uL, 1989; Yee und Fein, 2001 C"(Il) BacilllL.( subtili.t Beverldge und Murruy, 1976 Escherichia coli Beveridge und Koval, 1981; Hoyle und Beveridge, 1983 Bacillu.( lichenifi,nnis Beveridge et uL, 1982 A.fcophyllu11t nodo.mm Kuyucak und Volesky, 1989 Oscillatoria (lTIgui.fti.uima Ahuja et nl., 1999u Cr(IiI;, CrfiI) BucillIL( subtilis Mayers and Beveridge, 1989; Fein tli aI., 2001 Sarg(wumjIuituns Krutuchvil et nl., 1998 Cu(Ifj Bacillus suhtilis Beveridge und Murruy, 1976; Fowle und Fein, 2001 Vauchcriu alga Crist et ul., 1981 KlelMiclio pneumoniae Goncnlves et ul., 1987 Oscillatorir: an6~,li.~ti.<;.~imr: .A.!mjn et ill., 1997 E.(c/terichill coli Beveridgu and Koval, 1981; Hoyle and Beveridge, 1983 Bacillus .mbtili.f Fein et ul., 1997; Fowle and Fein,1999 BacilllL.f licllCniJonnis Beveridge et al., 1982; Fowle and Fein, 1999) Hg(II) BaciIIlL.~ subtili.( Beveridge and Murray, 1976 E!lCherichia coli Beveridge and Kovul. 1981; Hovle and Beveridge. 1983 BacilllLt licheniJonni.( Beveridge et al., 1982' - Fe(llfj Bucillu.( .mbtilis Mayers and Beveridge, 1989 Ni(I1) Bucillu.( subtili.( Beveridge and MUITay,1976 Escherichia coli Beveridge und Koval, 1981; Hoyle and Beveridge, 1983 Bacillu.( lichenifonni.f Beveridge eL a1., 1982 Pb(ll) Bacillu.t suhtilL( Beveridge and Murray, 1976 E,(cherichiu coli Beveridge and Koval, 1981j Hoyle und Beveridge, 1983 P,wudomoau.( atlantica Harvey and Leckie,1985 Sargll.s.mmjluit(lfu Fourest and Volesky, 1996 BacillILf subtili.( Fein el aI., 1997; Fowle and Fein, 1999 BaciillL.f iicilCni./'onni.( Beveridge et ai., 1982; Daughney and Fein, 1998 Fowle and Fein, 1999 Penicillium cltry',wgcnum Sarret el al., 1999 Bacillus .mbtili.f Beveridge und MUITay, 1976 E!lchericltia coli Beveridge and Koval, 1981; Hoyle and Beveridge, 1983 Hwdwna ulga Crist at ul., 1981 U(Vfj E5cherichia cfJli Beveridge and Koval, 1981; Hoyle und Beveridge, 1983 Zn(ll) Bacillll.'l !luhtilLf Beveridge and Murray, 1976 E.~cherichia coli Beveridge and Kuval, 1981 VaucllCria alga Crist elnl., 1981 Bacill!M lich!!niformi.~ Beveridge et nt, 1982 O,~cilluttJria fJngui.(ti.f,fimrt Ahuja et nl., 1999b Shewanelln putrifacicn.~ Cooper el u1., 2000

I. 1032 BROWN AND P,IRKS

2000; Yee and Fein, 2001) have used acid/base sites. No significant differences were observed in titrations to determine deprotonation constants for the uptake of these ions on the two bacteria in mixed the surface fUIictionul groups, metal-hacteria bacteria. upta.ke experiments. One important conclu adsorption experiments, and equilibrium surface sion reached in this study is that adsorption of these complexation modeling (SCM) to estimate stability trace metal ions onto these two bacteria appears constants for the metal-bacterial surface complexes. to occur abiotically, with no preference given to a Tn " ~t11rlv of thp. Rrl;;nrntion of eu(lIt Cdill). and particular metal ion beause it is less toxic to the _ .. - ----J ------~- '''' " Pb(II) onlo Bacillus subtilis, Fein el a1. (1997) found bacteria. that the adsorption dala were best fit by assuming In similar types of uptake experiments. Yee and three different types of surface functional groups in Fein (2001) measured the adsorption of Cd(II) on the B. subtilis cell walls (carboxylic, phosphate, and two gram-negative (Escherichia coli and Pseudomo hydroxyl sites) with pK/J values of 4.8,6.9, and 9.4, nas aemginosa) and five gram-positive bacteria respectively. SCM fits of the cation uptake data indi (Bacillus megaturium, Streptococcus faecalis, Sta cate that carboxyl sites are dominant and highly sta phylococcus aureus, Sporosarcina ureae, Bacillus ble, with log stability constants of 4.3 for Cu(II), 3.4 cereus). They found that Cd(II) adsorption on these for Cd(II), and 4.2 for Pb(I1) for the pH range 2 to 6, bacteria under essentially th~ same conditions indicating that Cu(I1) and Pb(II) bind more strongly showed no significant differences and could be fit to the B. subtilis cell wall than Cd(II). with a single adsorption edge as a function of pH. In a related study, Daughney and Fein (1998) These observations led Yee and Fein to propose that examined the effect of ionic strength on the adsorp metal-bacteria adsorpiion is not dependent on the tion of Cu(II), Cd(II), and Pb(I!) on Bacillus subtiiis type of bacterial species. These and similar uptake and Bacillus licheniformis and found that the uptake experiments have clearly shown that certain metals data are best described by a two-site model which [e.g., Cu(II), Pb(II)] have a high affinity for func- included carboxyl iind phosphate binding sites. tiona! groups in bacterial cell walls, whereas others They also found that increasing ionic strength [e.g., Cd(II)] do not. caused an increase in the metal-bacteria stability constants, which is a trend opposite to that found for EXAFS studies of metal ions coordination sites in metal binding on mineral surfaces (Davis and Kent, bacterial and fungal cell walls 1990). However, the structure of these metal surface Synchrotron-based experimental approaches are complexes and their mode of attachment to the bac well suited for studying the interaction of cations terial cell walls are unknown, so it is not currently with bacterial cells, although only a few systems possible explain this difference. have been studied to date. Extended x-ray absorp In a somewhat more complicated multicompo tion fine structure spectroscopy (EXAFS) can be nent system, competitive adsorption of Ca(II), used to determine which functional groups (e.g., Cu(I1), Cd(I1), and Pb(I1) on B. subtiiis and B. carboxylic, phosphoryl, phenolic) within a biofilm or lichenifonnis has been studied by Fowle and Fein bacterial or fungal cell wall are reactive toward trace (1999). Binding of these ions to carboxyl and phos metals (e.g., Sarret et aI., 1999; Kelley et aI., 2001; phate sites in the cell walls of these bacteria is indi Webb et aI., 2001). For example, Sarret et aI. (1999) cated by a two-site surface complexation model with used EXAFS spectroscopy and solution chemistry stability constants consistent with metal binding to experiments to show that Pb(II) reacts with carboxyl •• • ", '.1 11 11 r .1 .. r. these two functional groups. The uptake of Cu(II) ana pnospnoryl SHes In me cell wau 01 lIIe lLi.ilgUS and Pb(II) on B. subtiiis at high bacteria-to-dis Penicillium chrysogenum. with weak binding of solved metal ratio was found to be essentially the Pb(II) to carboxyl groups accounting for 95% of same as a function of pH (pHntls :=::3.5), and both ions Pb(II) uptake and strong binding of Pb(II) to phos

adsorbed more strongly than Cd(II) (pHnds ~4.5). On phor"-/I groups accounting for 5% of the uptake. In an B. licheniforTnis at high bacteria-to-dissolved metal EXAFS study of the sorption products of Cd(II) and ratio, the same trend was observed, with Cu(Il) and U(VI) on Bacillus subtilis, Kelly et al. (2001) found Pb(I!) sorbing more strongly (pH,d. ",3.8) than Cd(II) that U(V!) binds to phosphoryl groups and that (oH_.l.. :=::5), At low bacteria:dissolved metal ratios. Cd(ll) binds to carboxyl groups at low pH. The inter the pH-mls 'values were found to be about 1 to 2 pH action of Zn(ll) with five microbial species cultured units higher than for the high ratio, indicating more from a lake sediment was studied by Webb et aI. competition among these ions for bacterial cell wall (2001) using EXAFS spectroscopy. This study SORPTION OF TIIACE ELEMENTS ON MINE&lL SURFACES 1057

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