Diss. ETH No. 15492

Formation and Reactions of Green under Geochemical Conditions found in Natural Soils and Sediments

A dissertation submitted to the

SWISS FEDERAL INSTITUTE OF TECHNOLOGY

for the degree of

DOCTOR OF NATURAL SCIENCES

presented by

MARIANNE ERBS

M.Sc. in environmental chemistry

born January 13, 1973

in Haderslev, Denmark

Accepted on recommendation of

Prof. Dr. Rene P. Schwarzenbach, examiner

Prof. Dr. Stefan B. Haderlein, co-examiner

Prof. Dr. Hans C.B. Hansen, co-examiner

Zürich 2004

In fond memory of my mother

Esther Kristine Erbs (1949-2002)

who taught me how to be strong, feel joy and bear compassion. I dedicate this work to her. Without her support, care and love, I would never have been the person I am today.

To dare is to lose one's footing momentarily. Not to dare is to lose oneself.

Søren Kierkegaard

Acknowledgements

I would like to thank Stefan Haderlein, Hans Christian B. Hansen and Rene Schwarzenbach for their supervision of this work. Without the encouragement and confidence of H.C.B. Hansen and former colleagues at the Royal Veterinary and Agricultural University in Copenhagen, I would never have pursued a Ph.D. and without the understanding of Rene Schwarzenbach after the tragic death of my mother, I would not have had the time necessary to finish it.

I thank Christian Bender Koch, Hanne Nancke-Krogh, Susanne Guldberg and Henrik T. Andersen (Royal Veterinary and Agricultural University, Denmark) for their valuable contribution to my work. I would also like to express my gratitude to former and present members of the Contaminant Hydrology Group from whom I have received many benefits. I mourn the loss of Denis Mavrocordatos (EAWAG), who provided technical assistance in the electron microscopy lab, and I will always keep the sunny hours in his company in fond memory. Finally, I would like to thank Kristina Straub and Bernhard Schink (University of Constance, Germany) who welcomed me in their lab for a week and taught me how to work with strict anaerobic bacteria.

I gratefully acknowledge the grant which I received from the Danish Research Agency.

Table of Contents

Table of Contents

Zusammenfassung I Summary V

1 General Introduction 1

1.1 cycling in the subsurface 1 1.2 Green rusts 3 1.3 Microbial formation of green rusts 7 1.4 Redox reactions of green rusts 8 1.5 Outline of the thesis 10 References 11

2 Solid State Oxidation of Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 17

Abstract 17 2.1 Introduction 17 2.2 Materials and methods 22 2.2.1 Microorganisms and media 22 2.2.2 Characterisation of precipitates 23 2.2.3 Biooxidation experiments 24 2.2.4 Analytical methods 25 2.3 Results and discussion 25 2.3.1 Identification of solid iron-containing phases 25 2.3.2 Factors controlling the rate and extent of Fe(II) biooxidation 34 2.3.3 Morphology of solid iron phases 37 2.4 Conclusions 38 References 39

3 Formation of Layered Iron by Microbial Fe(III) Reduction 43

Abstract 43 3.1 Introduction 44 3.2 Materials and methods 47 3.2.1 Preparation of iron 47 3.2.2 Mineral characterisation 48 3.2.3 Culture conditions and cell preparation 48 3.2.4 Bioreduction experiments 49 Table of Contents 3.2.5 Analytical methods 50 3.3 Results and discussion 50 3.3.1 Fe(II) production and suspension colour changes 50 3.3.2 Identification of solid iron phases 55 3.3.3 Factors controlling the identity of the secondary iron minerals 58 3.3.4 Factors controlling the rate and extent of Fe(III) bioreduction 59 3.4 Conclusions 60 References 61

4 Reduction of Nitroaromatic Probe Compounds by Sulphate Green : The Effect of Probe Compound Charge 65

Abstract 65 4.1 Introduction 66 4.2 Materials and methods 71

4.2.1 Synthesis of GR-SO4 71 4.2.2 Mineral characterisation 72 4.2.3 Lyophilization and determination of specific surface area 72 4.2.4 Estimation of the one-electron reduction potential for 4-NPA 73 4.2.5 Kinetic experiments 74 4.2.6 Analytical methods 74 4.3 Results and discussion 75 4.3.1 Product formation and reaction kinetics 75 4.3.2 Comparison of rate constants for the different NACs 79 4.3.3 Factors influencing the reaction rate 82 4.3.4 Comparison with rate constants obtained for other Fe(II) containing mineral systems 83 4.3.5 Depletion of reactive sites 85 4.3.6 The role of external and internal reactive sites 86 4.4 Conclusions 89 References 91

5 Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 97

Abstract 97 5.1 Introduction 98 5.2 Materials and methods 101 5.2.1 Synthesis of GRs and 102 5.2.2 Preparation of coatings 102 5.2.3 Mineral characterisation 103 5.2.4 Kinetic experiments 103 Table of Contents 5.2.5 Analytical methods 104 5.3 Results and discussion 105 5.3.1 Product formation and reaction kinetics 105 5.3.2 Comparing rate constants obtained for the various Fe(II)-Fe(III) mineral systems 109 5.3.3 Comparing with rate constants obtained for other chlorinated aliphatic compound 112 5.3.4 Factors controlling the reactivity of surface-bound Fe(II) 114 5.3.5 Comparison with biotic and other abiotic systems 118 5.4 Conclusions 119 References 120

6 Conclusions and Outlook 125 References 128

7 Supporting Information I 7.1 Estimation of the one-electron reduction potential for 4-NPA I 7.2 The rate-limiting step IV 7.2.1 Mass transfer (diffusion) limited kinetics V 7.2.2 Surface saturation limited kinetics IX

7.3 External surface area of GR-SO4 and GR-CO3 XI 7.4 Van der Waals radii XIV 7.5 Adsorption of Fe(II) onto Fe(III) XVI References XVIII

Curriculum Vitae

Zusammenfassung I

Zusammenfassung

Geschichtete Fe(II)-Fe(III)- (Grüner Rost) gehören zur Gruppe der

Fe(II)-haltigen Mineralsysteme (z.B. Magnetit (Fe3O4), Siderit (FeCO3), Vivianit

(Fe2(PO4)2⋅8H2O), Fe(II)-Sulfide sowie an die Oberfläche von Fe(III)-Oxiden und Tonmineralien gebundenes, zweiwertiges Eisen), die die Aktivität von Fe(II) in suboxischen und anoxischen Böden und Sedimenten kontrollieren. Grüner Rost Phasen (GRs) bestehen aus planaren, positiv geladenen, trioktaedrischen Fe(II)- Fe(III)-Hydroxidschichten, die durch hydratisierte Anionen in den Zwischenschichten ausgeglichen werden. Ihre generelle Zusammensetzung ist II III x+ x- [Fe (6-x)Fe x(OH)12] [(A)x/n·yH2O] , wobei x = 0.9 - 4.2 ist, A entspricht einem n- 2- – 2- valenten Anion (z.B. CO3 , Cl oder SO4 ) und y repräsentiert die Anzahl Wassermoleküle in der Zwischenschicht. GRs sind wichtige intermediäre Phasen, die durch unvollständige Oxidation von Fe(II) oder teilweise Reduktion von Fe(III) gebildet werden können. Sie können in suboxischen, nicht-sauren, eisenhaltigen natürlichen, wie auch technischen Systemen auftreten, so wie in Wasser gesättigten Böden und interstitiellen Sedimenten, Rohrleitungen in der Trinkwasserversorgung, Stahlpfosten in marinen Sedimenten, Stahlbeton, und in reaktiven durchlässigen Wänden aus nullwertigem Eisen zur in-situ Sanierung von Altlasten und Aquiferen. Aufgrund ihrer Schichtstruktur, den anionischen Zwischenschichten und der hohen spezifischen Oberflächen sind GRs reaktive Ionentauscher und Sorbentien von Anionen. Des Weiteren wurde gezeigt, dass GRs eine Reihe anorganischer und organischer Schadstoffe reduzieren können. Durch Immobilisierung und Transformation können GRs somit eine wichtige Rolle für das Abbauverhalten und den Transport solcher Schadstoffe in suboxischen Böden und Sedimenten spielen. Die Resultate dieser Dissertation tragen zum Verständnis über die Bildung und Reaktivität von Fe(II)-haltigen Mineralsystemen, wie GRs, Vivianit, Magnetit und an Goethit (α-FeOOH)- und Lepidokrozit (γ- FeOOH)-Oberflächen gebundenes Fe(II), in der Natur bei.

II Zusammenfassung Um die Rolle von Bakterien bei der Bildung von GRs in natürlichen Böden und Sedimenten aufzuklären, wurden Eisenminerale untersucht die als Folge der Aktivität von eisenrespirierenden Bakterien gebildet wurden. Kapitel 2 beschreibt die Untersuchungen von eisenhaltigen Produkten, die von anaeroben, autotrophen, denitrifizierenden, Fe(II)-oxidierenden Bakterien (FeOB) gebildet wurden. Ein Bikarbonat- und Phosphat-reiches Kulturmedium bot den nitratreduzierenden FeOB optimale Bedingungen. Fe(II) lag zu Anfang der Reaktion als weisses Fe(II)-Hydroxyphosphat (Vivianit) und als gelöstes Fe(II) vor. Die Ergebnisse zeigten, dass die denitrifizierenden FeOB amorphen Goethit via ein grünes Fe(III)- angereichertes Vivianit-Zwischenprodukt bildeten. Die Analyse mit Mössbauer Spektroskopie deutet nicht auf eine Bildung von GR hin.

In Kapitel 3 werden jene Eisenmineralien beschrieben, die während der Reduktion verbreiteter Fe(III)-Oxide durch anaerobe, dissimilative, Fe(III)-reduzierende Mikroorganismen, Shewanella algae BrY, gebildet wurden. Um natürliche Zustände zu simulieren wurden Fe(III)-Oxide als Beschichtungen auf

Silikatpartikel (Modellsystem für Sandböden) oder Calcitpartikel (CaCO3; Modellsystem für kalkhaltige Böden) aufgetragen, sowie synthetische Elektronencarrier und hochkonzentrierte, künstliche pH-Puffer ausgeschlossen. Die erforschten Mineralsysteme umfassten Goethit/Calcit-, Lepidokrozit/Calcit- und Ferrihydrit/Sand-Suspensionen. S. algae BrY reduzierte beachtliche Mengen des eingesetzten Fe(III), und es bildeten sich grüne und schwarze Festphasen innerhalb von 1-2 Wochen nach der Animpfung. Mössbauer Spektroskopie der grünen und schwarzen Präzipitate zeigte, dass sich diese aus GR und Vivianit zusammensetzen.

Die Reaktivität synthetischer GRs gegenüber reduzierbaren organischen Schadstoffen wurde erkundet, um die potentielle Bedeutung von GR-Phasen für das Schicksal solcher Verbindungen abzuschätzen. Zu diesem Zweck wurden Nitroaromaten (NACs) und Chloracetate als Modellverbindungen benutzt, um Zusammenfassung III umweltrelevante Redoxreaktionen zu studieren. In Kapitel 4 wurde die relative Reaktivität von äusseren und inneren reaktiven Stellen in synthetischem Sulfat-

Grünem Rost (GR-SO4) anhand von strukturähnlichen “reaktiven Sondenmolekülen” mit unterschiedlichen Ladungen untersucht. Als reaktive Sondenmoleküle wurden Nitrobenzen, 2-Nitrophenol, 4-Nitrotoluen, 4- Chlornitrobenzen und 4-Nitrophenylessigsäure verwendet. Die Ergebnisse zeigen, dass GR-SO4 die NACs vollständig zu den entsprechenden Anilinen reduzierte. Die Reaktionen folgten einer pseudo 1. Ordnungs Kinetik bezüglich NAC und die auf Oberfläche normalisierten pseudo 1. Ordnungs Geschwindigkeitskonstanten -4 -1 -2 (Anfangsraten) waren 0.16–4.65·10 s ·m ·L für [Fe(II)GR]0 = 1.03-12.60 mM,

[NAC]0 = 20-102 µM und pH 8.4-8.6. Weder durch Einbezug von Massentransferlimitierung noch von Oberflächensättigungskinetik war es möglich, die ähnlichen Oberflächennormalisierten pseudo 1. Ordnungs Geschwindigkeitskonstanten für die Reduktion der neutralen und anionischen

NACs durch GR-SO4 zu erklären. Dieser Umstand lässt vermuten, dass die

Reaktion zwischen NAC und GR-SO4 an den externen reaktiven Fe(II)-Stellen stattfindet. Bei niedrigen Fe(II)GR-Anfangskonzentrationen wurden die externen reaktiven Fe(II)-Stellen aufgebraucht, und die Regenerierung von neuen externen reaktiven Stellen haben schliesslich die Geschwindigkeit der Reduktion von NACs durch GR-SO4 kontrolliert.

In Kapitel 5 wurde die Reaktivität von verschiedenen umweltrelevanten Fe(II)- Fe(III)-Mineralsystemen gegenüber Trichloressigsäure (TCA) und Dichloressigsäure (DCA) in Batchexperimenten, die natürliche Bedingungen imitierten, untersucht. Die Fe(II)-Fe(III)-Systeme umfassten Sulfat-Grüner Rost, Carbonat-Grüner Rost, Magnetit, Fe(II)/Goethit und Fe(II)/Lepidokrozit. TCA wurde von allen Fe(II)-haltigen Mineralien zu DCA reduziert. Die Reaktionen folgten einer pseudo 1. Ordnungs Kinetik bezüglich TCA und die auf Oberfläche normalisierten pseudo 1. Ordnungs Geschwindigkeitskonstanten betrugen 0.33– -5 -1 -2 7.6·10 min ·m ·L bei [Fe(II)]0 = 0.25–11.6 mM, [TCA]0 = 15–1000 µM und pH IV Zusammenfassung 7.0–8.7. Die Ergebnisse zeigen keine signifikanten Unterschiede zwischen den verschiedenen Fe(II)-Fe(III)-Systemen bezüglich Produkteverteilung und oberflächen-normalisierten pseudo 1. Ordnungs Geschwindigkeits-konstanten. In keinem der Systeme wurde DCA innerhalb des experimentellen Zeitraums zu Monochloressigsäure oder Essigsäure weiter reduziert.

Die Ergebnisse die in dieser Dissertation präsentiert werden, zeigen, dass mikrobiologische Prozesse für die Oxidation von Vivianit-Phasen im Untergrund verantwortlich sein können. Zudem wurde nachgewiesen, dass GRs bei tiefen Kohlenstoff- und Fe(III)-Konzentrationen, sowie durch Ausschluss von künstlichen Elektronencarriern und pH-Pufferung, mikrobiell gebildet werden können. Ferner zeigten Befunde, dass GRs eine bedeutende Rolle für die reduktive Transformation von NACs und TCA in natürlichen Böden und Sedimenten spielen können. Summary V

Summary

Layered iron(II)-iron(III)-hydroxides (green rusts) belong to the group of Fe(II)- bearing mineral systems, e.g. magnetite (Fe3O4), siderite (FeCO3), vivianite

(Fe2(PO4)2⋅8H2O), Fe(II) sulfides as well as Fe(II) associated with Fe(III) oxide and clay mineral surfaces, that control the Fe(II) activity in suboxic and anoxic soils and sediments. Green rusts (GRs) consist of plane, positively charged, trioctahedral Fe(II)-Fe(III) hydroxide sheets balanced by hydrated anions in the II III x+ x- interlayers and hold the general formula [Fe (6-x)Fe x(OH)12] [(A)x/n·yH2O] , 2- – 2- where x = 0.9 - 4.2, A is an n-valent anion, e.g. CO3 , Cl or SO4 , and y is the number of molecules in the interlayer. GRs are important intermediate phases formed by partial oxidation of Fe(II) or partial reduction of Fe(III) and they have been found in suboxic, non-acid iron-rich natural environments such as hydromorphic soils and intertidal sediments and in engineering systems including pipeline distribution systems for drinking water, sheet piles in marine sediments, reinforced and permeable reactive barriers of zero-valent iron implemented for on-site remediation of contaminants. Due to their layered structures, anionic interlayers and high specific surface areas, GRs represent reactive exchangers and sorbents of anions. In addition, GRs have been shown to reduce a range of inorganic and organic pollutants. Thus, through sequestration and reductive transformation, GRs may play an important role in the fate and transport of contaminants in suboxic soils and sediments. The work presented in this dissertation adds to the understanding of how Fe(II)-bearing minerals like GRs, vivianite, magnetite and Fe(II) associated with (α-FeOOH) and (γ-FeOOH) may form and react in nature.

In order to elucidate the role of bacteria in the formation of GRs in natural soils and sediments, we studied the iron mineral phases forming as a result of the activity of iron-respiring bacteria. In the study described in chapter 2, the Fe- containing products formed by anaerobic, autotrophic denitrifying Fe(II)-oxidizing VI Summary bacteria (FeOB) were examined. The culture medium applied contained high levels of bicarbonate and phosphate and is typically used in this kind of studies as it provides excellent conditions for the nitrate-reducing FeOB. Fe(II) was present initially as a whitish solid Fe(II) hydroxy phosphate (vivianite) and as soluble Fe(II). The results obtained demonstrate that the denitrifying FeOB produce poorly crystalline goethite via a greenish Fe(III)-enriched vivianite intermediate. Mössbauer spectroscopic analyses provided no significant evidence of formation.

In chapter 3, the Fe-containing products formed during reduction of common Fe(III) oxides by the anaerobic dissimilatory Fe(III)-reducing microorganism, Shewanella algae BrY, are discussed. In order to simulate natural conditions, Fe(III) oxides were applied as coatings on silica (model system for sandy soils) or calcite (CaCO3) particles (model system for calcareous soils) and synthetic electron shuttles as well as highly concentrated artificial pH buffers were excluded. The mineral systems studied include goethite/calcite, lepidocrocite/calcite and hydrous ferric oxide/sand suspensions. S. algae BrY reduced substantial amounts of the initial Fe(III) and green and blackish mineral phases were produced within 1-2 weeks after inoculation. Mössbauer spectroscopic analyses showed that the green and black precipitates consisted of GR and vivianite.

The reactivity of synthetic GRs towards reducible organic pollutants was investigated in order to asses the potential significance of GR phases for the fate of such compounds. To this end, we used nitroaromatic compounds (NACs) and chlorinated acetates as suitable model compounds for studying environmentally relevant redox reactions. In the work described in chapter 4, the relative reactivity of outer and inner Fe(II) reactive sites in synthetic green rust (GR-SO4) was studied using a series of structurally closely related compounds with different charge properties as “reactive probes”. The probe compounds included nitrobenzene, 2-nitrophenol, 4-nitrotoluene, 4-chloronitrobenzene and 4- Summary VII nitrophenylacetic acid. The results show that NACs are completely reduced to their corresponding anilines by GR-SO4. The reactions followed pseudo 1. order kinetics with respect to NAC and the surface area-normalised pseudo 1. order rate -4 -1 -2 constants (initial rates) obtained were 0.16–4.65·10 s ·m ·L at [Fe(II)GR]0 = 1.03-

12.60 mM, [NAC]0 = 20-102 µM and pH 8.4-8.6. Neither mass transfer control nor surface saturation kinetics could explain the similarity of the surface-normalised pseudo 1. order rate constants obtained for the reduction of the neutral and anionic

NACs by GR-SO4. These observations suggest that the reaction between NAC and

GR-SO4 takes place at the external reactive Fe(II) sites. At low initial Fe(II)GR concentrations, the external reactive Fe(II) sites were depleted and the regeneration of new external reactive sites eventually controlled the reduction of the NACs by

GR-SO4.

Finally, the reactivity of various Fe(II)-Fe(III) mineral systems towards trichloroacetic acid (TCA) and dichloroacetate (DCA) has been investigated in laboratory batch experiments imitating natural conditions (Chapter 5). The Fe(II)-

Fe(III)-systems investigated included GR-SO4, carbonate green rust, magnetite, Fe(II)/goethite and Fe(II)/lepidocrocite. TCA was readily reduced to DCA by all Fe(II)-containing minerals. The reactions followed pseudo 1. order kinetics with respect to TCA and the surface area-normalised pseudo 1. order rate constants -5 -1 -2 obtained were 0.33–7.6·10 min ·m ·L at [Fe(II)]0 = 0.25–11.6 mM, [TCA]0 = 15–1000 µM and pH 7.0–8.7. Our results showed no significant differences regarding product distribution and surface area-normalised reaction rate constants between the Fe(II)-Fe(III)-systems. DCA was not further reduced to monochloroacetate (MCA) or acetate in any of the systems within the time frame in our experiments.

The results presented in chapter 2 indicate that microbiological processes may be responsible for the oxidation of vivianite phases in natural subsurface environments. In chapter 3, we demonstrated that GRs may be produced VIII Summary microbially at conditions including low carbon and Fe(III) concentrations as well as the exclusion of synthetic electron shuttles and pH buffers. The results obtained in chapter 4 and 5 show that GRs transform NACs and TCA readily. The reductive transformation of NACs and TCA by GRs is relevant to understanding the processes responsible for their degradation in the subsurface and the development of innovative technologies for their remediation.

General Introduction 1

1 General Introduction

1.1 Iron cycling in the subsurface Iron is the fourth most abundant element (4-5 mass%) and the most abundant redox sensitive element in the Earth’s crust. It is found as Fe(II) and Fe(III) in a number of minerals in rocks, soils and sediments. Under anoxic conditions, solid Fe(III)- containing minerals can be reduced to soluble Fe(II) once the more energetically favoured electron donors - nitrate and manganese(IV) oxides - have been consumed. Dissolved Fe(II) can be reoxidized to insoluble Fe(III) microbially, or abiotically upon exposure to . Due to this ready alternation between the Fe(II) and Fe(III) redox states, iron plays a major role in controlling the redox potential and the carbon cycling in subsurface environments (Nealson & Saffarini, 1994).

Nonenzymatic processes were previously considered to account for most of the Fe(III) reduction in subsurface environments. The significance of bacteria in the biogeochemical cycling of iron has been broadly recognized over the past two decades. Dissimilatory Fe(III)-reducing bacteria (DIRB) that gain energy by coupling the oxidation of hydrogen or organic compounds to the reduction of Fe(III) oxides have been known for many years, but their biogeochemical importance was only widely acknowledged about a decade ago (reviewed by Lovley, 1997). Fe(III) bioreduction accounts for a major fraction of the carbon oxidation in many different anoxic environments and in the presence of sufficient amounts of reactive Fe(III), microbial Fe(III) reduction may even inhibit sulphate reduction and methanogenesis (King, 1990; Lovley & Phillips, 1986). In fact, most of the Fe(III) reduction in the Fe(III) reduction zone of aquatic sediments and aquifers is enzymatically catalyzed by microorganisms (Lovley et al., 1991). A wide diversity of DIRB, distributed among several different phylogenetic groups, 2 Chapter 1 is known today. The two most studied DIRB are the obligate anaerobic Geobacter spp. and the facultatively anaerobic Shewanella spp. (Figure 1.1).

Aerobic oxidation of Fe(II)-containing minerals by lithotrophic acidophilic and neutrophilic bacteria has been known for many years, but their broad significance in the biogeochemical cycling of iron has only been recognized over the past two decades. Both acidophilic (Thiobacillus ferrooxidans) and neutrophilic (Gallionella ferruginea, Leptothrix ochracea, Sphaerotilus natans) aerobic Fe(II)- oxidizing bacteria (FeOB) have been isolated (Hanert, 1992; Kuenen et al., 1992; Mulder & Deinema, 1992).

Figure 1.1. The microbial iron cycle.

Anaerobic Fe(II) oxidation by phototrophic, purple, non-sulfur bacteria utilizing Fe(II) as an electron donor in the light was recognized only a decade ago (Widdel et al., 1993). Subsequently, it was demonstrated that the biological oxidation of Fe(II) in the absence of oxygen is possible by light-independent, chemotrophic microorganisms using nitrate as the electron acceptor (Straub et al., 1996). Thus, the microbial iron cycle includes anaerobic Fe(III)-reducing microorganisms and aerobic as well as anaerobic Fe(II)-oxidizing bacteria (Figure 1.1).

General Introduction 3 1.2 Green rusts Iron oxides, iron hydroxides and iron oxyhydroxides (collectively termed iron oxides or Fe(III) oxides) are ubiquitous in the pedosphere where they originate from aerobic weathering of surface magmatic rocks such as ferromagnesium silicates and pyrite (Cornell & Schwertmann, 1996). Goethite (α-FeOOH), lepidocrocite (γ-FeOOH), (Fe5HO8⋅4H2O), hematite (α-Fe2O3), magnetite (Fe3O4), maghemite (γ-Fe2O3) and akageneite (β-FeOOH) constitute the most important iron oxides in soils and sediments (Schwertmann & Cornell, 1991). The formation and transformation of iron oxides depend on pH, solution composition, redox potential, temperature, rate of oxidation/reduction and degree and rate of hydration/dehydration. Iron oxides are important to many soil properties such as colour, pH and redox buffer capacity, aggregation with other soil particles as well as retention of anions and cations (Cornell & Schwertmann, 1996). A number of Fe(II)-bearing minerals including Fe(II)-containing clays (e.g. smectites, vermiculites, and micas), magnetite, siderite (FeCO3), vivianite

(Fe2(PO4)2⋅8H2O), Fe(II) sulphides and green rusts (layered Fe(II)-Fe(III) hydroxides) may be present in soils and sediments under suboxic and anoxic conditions. Green rusts are believed to play a central role as metastable intermediates in the redox cycling of iron at circumneutral pH in aquatic and terrestrial environments.

Green rusts (GRs) are layered iron(II)-iron(III)-hydroxides consisting of plane, positively charged, trioctahedral Fe(II)-Fe(III) hydroxide sheets balanced by hydrated anions in the interlayers (cf. Figure 4.1, this work). GRs belong structually to the pyroaurite-sjögrenite group of layered hydroxides and they hold II III x+ x- the general formula [Fe (6-x)Fe x(OH)12] [(A)x/n·yH2O] , where x = 0.9 - 4.2, A is 2- – 2- an n-valent anion, e.g. CO3 , Cl or SO4 , and y is the number of water molecules in the interlayer. The three most common and investigated green rust forms include GR (GR-Cl), sulphate GR (GR-SO4) and carbonate GR (GR-CO3). Generally, GRs are crystallographically classified into the GRI (rhombohedral; 4 Chapter 1

GR-Cl and GR-CO3) and GRII (hexagonal; GR-SO4) crystal systems. The GR interlayer thickness is a function of both the size and the charge of the interlayer anion. Tetrahedrally coordinated anions like sulphate lead to larger interlayer distances than smaller monoatomic anions like chloride or planar like carbonate (Mendiboure & Schöllhorn, 1986). Not only size but also charge density plays a role for the interlayer spacing. That is, for anions having the same number of valence electrons, anions with smaller ionic radii (higher electron density) are bound more strongly and therefore result in smaller interlayer spacings. The interlayer in GR-SO4 is composed of two consecutive planes of anions and water whereas GR-Cl and GR-CO3 interlayers consist of only one single plane (Simon et al., 2003).

GRs are important intermediate phases formed by partial oxidation of Fe(II) or partial reduction of Fe(III). In neutral and weakly alkaline solutions, the oxidation of dissolved Fe(II) always passes through solid GR phases (Bernal et al., 1959). GRs may also form during oxidation of zero-valent iron and as a result of the combination of Fe(II) and Fe(III) at circumneutral pH (Figure 1.2).

Figure 1.2. Formation and transformation of GRs. Fe3O4 = magnetite, γ-Fe2O3 = maghemite, α- FeOOH = goethite, γ-FeOOH = lepidocrocite, akageneite = β-FeOOH. General Introduction 5

Oxidation of GR-CO3 usually produces goethite and magnetite-maghemite whereas GR-Cl and GR-SO4 transform into lepidocrocite and magnetite- maghemite depending on pH and oxidation rate (Bernal et al., 1959; Taylor, 1980; Carlson & Schwertmann, 1990). The brown δ-FeOOH is formed by vigorous oxidation of GR using air or a 30% aqueous solution of hydrogen peroxide (Bernal et al., 1959; Misawa et al., 1974). Black ferromagnetic magnetite forms by slow oxidation of GR whereas lepidocrocite forms at high oxidation rates (Misawa et al., 1974). The presence of chloride is a prerequisite for the formation of akageneite (Bernal et al., 1959; Refait & Genin, 1997).

A substantial amount of work has been conducted in order to estimate the free energies of formation of green rusts. The free energies of formation reported for the carbonate and sulphate GRs fall in the range 4234–4384 kJ⋅mol-1 as determined from solution data monitored during anoxic alkalimetric titrations and from reduction potential (Eh) and pH recordings monitored during oxidation of GRs in aqueous solution (Hansen et al., 1994; Drissi et al., 1995; Genin et al., 1996). The free energies of formation provided allow for estimation of the stability domains of

GRs in Eh-pH phase diagrams (Drissi et al., 1995; Genin et al., 1996). As evidenced from such diagrams (Figure 1.3), the stability domain of GR-SO4 lies within pH 6-8 and Eh -700 – -400 mV depending on the activities of Fe(II) and sulphate (compare Figures 1.3a&b). This agrees with the natural GR occurrences found in suboxic non-acid iron-rich environments such as hydromorphic soils and intertidal sediments (Al-Agha et al., 1995; Trolard et al., 1996; Genin et al., 1998). In addition, GRs have been found as products in numerous engineering systems including a pipeline distribution system for drinking water, steel sheet piles in marine sediments, (ferro-concrete) and permeable reactive barriers of zero-valent iron implemented for on-site remediation of organic and inorganic contaminants (Tuovinen et al., 1980; Nielsen, 1976; Genin et al., 1991; Roh et al., 2000). 6 Chapter 1

-3 -3 -2 Figure 1.3. Eh-pH phase diagrams of GR-SO4. a) a 2+ = 10 , a 2− = 10 and b) a 2+ = 10 , Fe SO4 Fe

-1 a 2− = 10 . SO4

The stability domains of GR-Cl and GR-CO3 are similar to the stability domain of

GR-SO4. At Fe(II) and sulphate activities lower than depicted in Figure 1.3b, the stability domain of GR-SO4 will be situated at higher pH and lower Eh. Other dissolved species present at anoxic conditions such as phosphate, sulphide, carbonate and organic ligands may exert considerable effects on the availability of Fe(II) and Fe(III). At anoxic and circumneutral conditions, vivianite

(Fe2(PO4)2⋅8H2O) controls the Fe(II) activity even at very low phosphate concentrations (Nriagu & Dell, 1974). The formation of solid Fe(II) sulphides and siderite (FeCO3) as well as the complexation of Fe(II) and Fe(III) by organic ligands may also control the activity of Fe(II) in the subsurface and thereby interfere with the formation of GRs.

Due to their layered structures, anionic interlayers and high specific surface areas, GRs represent reactive ion exchangers and sorbents of environmentally concerning anions, e.g. arsenate and selenate (Myneni et al., 1997; Randall et al., 2001). In addition, GRs may incorporate divalent transition cations like Ni2+, Zn2+, Cd2+, Co2+ and Mg2+ by isomorphic substitution for Fe2+ in the hydroxide layers General Introduction 7 (Tamaura, 1985; Tamaura, 1986; Refait et al., 1994; Parmar et al., 2001; Refait et al., 2001). Furthermore, GRs have been shown to reduce a range of inorganic contaminants such as nitrite, nitrate, selenate, chromate, uranyl, pertechnetate and the transition AgI, AuIII, CuII and HgII as well as organic pollutants including halogenated ethanes, ethenes and methanes (Hansen et al., 1994; Hansen et al., 1996; Myneni et al., 1997; Erbs et al., 1999; Loyaux-Lawniczak et al., 1999; Cui & Spahiu, 2002; Lee & Batchelor, 2002b; Heasman et al., 2003; O’Loughlin et al., 2003a&b; Pepper et al., 2003; Elsner et al., 2004; O’Loughlin & Burris, 2004). Thus, through sequestration and reductive transformation, GRs may play an important role in the fate and transport of contaminants in suboxic soils and sediments. It should be noted that the rate constants reported for the reduction of these inorganic and organic pollutants by GRs cannot be directly compared as the various studies were conducted at very different experimental conditions.

1.3 Microbial formation of green rusts Generally, one would expect that biogenic minerals have chemical compositions and crystal habits similar to those produced by nonenzymatic processes as they are governed by the same equilibrium principles. In fact, since the latter stages of mineralization are abiotically driven and since the secondary Fe(II)-containing minerals are formed indirectly by electron transfer outside the bacterial cell and not directly inside the bacterial cell, the type of iron mineral formed is a function of the environmental conditions in which the bacteria live, i.e. the same microorganism form different minerals in different environments.

The microbial formation of GRs resulting from bioreduction of various Fe(III) oxides, including ferrihydrite, goethite and lepidocrocite, by strains of the anaerobic dissimilatory DIRB, Shewanella putrefaciens, has been reported repeatedly over the last years (Fredrickson et al., 1998; Kukkadapu et al., 2001; Parmar et al., 2001; Ona-Nguema et al., 2002a&b; Glasauer et al., 2003). However, no evidence of biogenic formation of GRs at natural geochemical 8 Chapter 1 conditions have been offered and it is still unknown whether this process may take place at natural conditions comprising low nutrient levels, low iron concentrations and the absence of synthetic electron shuttles and highly concentrated artificial pH buffers. Moreover, the biotic formation of GRs by anaerobic, denitrifying Fe(II)- oxidizing bacteria has been suggested but the phases still need to be properly identified (Chaudhuri et al., 2001). In order to elucidate the role of bacteria in the formation of GRs in natural soils and sediments, we studied the iron mineral phases forming as a result of the activity of iron-respiring bacteria (Chapters 2 and 3).

1.4 Redox reactions of green rusts Fe(II) is one of the most abundant reductants present in aquatic and terrestrial environments under suboxic and anoxic conditions (Lyngkilde & Christensen, 1992; Rügge et al., 1998). In these environments, Fe(II) may be present as soluble organic and inorganic complexes, as surface complexes and as a host of Fe(II)- bearing minerals. Although aqueous Fe(II) complexes may reduce a number of contaminants, Fe(II) associated with mineral surfaces and structural Fe(II) present in the mineral lattice in Fe(II)-containing minerals are often more powerful reductants. Fe(II)-bearing minerals including GRs, magnetite, siderite, Fe(II) sulphides as well as Fe(II)-carrying Fe(III) oxide and clay mineral surfaces have been shown to reduce a number of organic and inorganic contaminants such as nitroaromatic compounds, chlorinated aliphatics, chromate, uranyl, pertechnetate, nitrate, monochloramine and carbamate pesticides (Klausen et al., 1995; Cui & Eriksen, 1996; Butler & Hayes, 1998&1999; Erbs et al., 1999; Liger et al., 1999; Loyaux-Lawniczak et al., 1999; Amonette et al., 2000; Hwang & Batchelor, 2000; Hansen et al., 2001; Gander et al., 2002; Lee & Batchelor, 2002a&b; Pecher et al., 2002; Vikesland & Valentine, 2002; Hofstetter et al., 2003; O’Loughlin et al., 2003a&b; Strathmann & Stone, 2003; Elsner et al., 2004; O’Loughlin & Burris, 2004). However, only few comparative studies on the reactivity of Fe(II)-bearing minerals exist (Lee & Batchelor, 2002b; Elsner et al., 2004). When examining the General Introduction 9 reaction rates of the reductive transformation of NACs and chlorinated aliphatics by GRs and other Fe(II)-bearing minerals reported in these studies, the rate constants for GRs are mostly among the highest rates reported and in some cases even higher than the rate constants for Fe(II) sulphides. Thus, GRs may play an important role in the transformation of reducible contaminants in the subsurface.

Nitroaromatic compounds (NACs) are widely applied as explosives, herbicides, insecticides, solvents and intermediates in the synthesis of dyes and pesticides (Hartter, 1985; Rosenblatt et al., 1991). NACs are ubiquitous in the subsurface environment and pose a health risk due to their toxicity (Rickert, 1985). In anoxic environments, reduction of the nitro group is generally the first step during abiotic or microbial transformation of the NACs (Macalady et al., 1986). The transformation reaction generally produces the corresponding aromatic amines and minor amounts of intermediates (hydroxylamines and nitroso compounds) as well as coupling products (azo and azoxy compounds). These products may be of similar or even greater environmental concern.

Trichloroacetic acid (TCA) is ubiquitous in soils and the concentrations reported range from <0.05 to 380 µg/kg (Euro Chlor, 2001; McCulloch, 2002; Ahlers et al., 2003). On account of its phytotoxicity, suspected human carcinogenicity and widespread occurrence, TCA is of considerable environmental concern especially in the terrestrial compartment (Ahlers et al., 2003). Moreover, the daughter compounds of TCA - dichloroacetic acid (DCA) and monochloroacetic acid (MCA) - are also toxins and suspected human carcinogens as well as widespread in the environment (Reimann et al., 1996; Berg et al., 2000; Ahlers et al., 2003 and references therein). In this work, the reactivity of synthetic green rusts towards nitroaromatic compounds (NACs) and the reactivity of various Fe(II)-Fe(III) mineral systems including synthetic GRs towards chlorinated acetates have been studied (Chapters 4 and 5).

10 Chapter 1 1.5 Outline of the thesis An examination of the Fe-containing products produced during solid state oxidation of vivianite by anaerobic, autotrophic, denitrifying Fe(II)-oxidizing bacteria is presented in chapter 2. The Fe(II)-oxidizing bacteria were cultured in a mineral medium containing high levels of bicarbonate and phosphate which is typically used in this kind of studies as it provides excellent conditions for the nitrate-reducing FeOB. The solid iron phases forming were investigated by transmission Mössbauer , infrared spectroscopy and scanning electron microscopy.

Chapter 3 includes a study on the Fe-containing products formed during reduction of common Fe(III) oxides by the anaerobic dissimilatory Fe(III)-reducing microorganism, Shewanella algae BrY. In order to simulate natural conditions, Fe(III) oxides were applied as coatings on silica (model system for sandy soils) or calcite particles (model system for calcareous soils) and synthetic electron shuttles as well as highly concentrated artificial pH buffers were excluded. The mineral systems studied include goethite/calcite, lepidocrocite/calcite and hydrous ferric oxide/sand suspensions. The solid iron phases produced were examined by transmission Mössbauer spectroscopy.

A study on the relative reactivity of outer and inner Fe(II) sites in synthetic GR-

SO4 by using a series of structurally closely related compounds with different charge properties as “reactive probes” is presented in chapter 4. The probe compounds included nitrobenzene, 2-nitrophenol, 4-nitrotoluene, 4- chloronitrobenzene and 4-nitrophenylacetic acid.

In chapter 5, an investigation of the reactivity of various Fe(II)-Fe(III) mineral systems towards TCA and DCA is presented. The study included laboratory batch experiments imitating natural conditions. The Fe(II)-Fe(III)-systems investigated included GR-SO4, carbonate green rust, magnetite, Fe(II)/goethite and General Introduction 11 Fe(II)/lepidocrocite. The reactivities of the Fe(II)-Fe(III) mineral systems were examined by comparing their surface-normalized rate constants.

The results and environmental implications of this work are summarized in chapter 6.

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2 Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria

Abstract This work investigated the Fe-containing products formed by anaerobic, autotrophic denitrifying Fe(II)-oxidizing bacteria in a specific bicarbonate buffered - (30 mM HCO3 , pH 7.0) culture media containing 10 mM Fe(II), 4 mM nitrate and 4 mM phosphate. Fe(II) was present initially as a whitish vivianite-like

(Fe3(PO4)2·8H2O) precipitate and as soluble Fe(II). The initial phase of the II III oxidation produced a greenish metavivianite-like ((Fe 3-x,Fe x)(PO4)2(OH)x·(8- x)H2O, x > 1.2) phase. In the late oxidation phase, a reddish precipitate of poorly crystalline goethite (α-FeOOH) dominated the colour of the media in coexistence with Fe(II)-containing siderite (FeCO3). The increasing amounts of Fe(III) present in the “vivianite” and “metavivianite” structures were accompanied by an increasing intensity in the green colour as the Fe(II) biooxidation progressed. This colour development has produced the idea of biogenic green rusts (layered Fe(II)- Fe(III) hydroxides) in several studies on nitrate-dependent Fe(II) biooxidation. However, in this work no evidence of green rust formation mediated by anaerobic denitrifying Fe(II)-oxidizing bacteria was obtained.

2.1. Introduction Aerobic oxidation of Fe(II)-containing minerals by lithotrophic acidophilic and neutrophilic bacteria has been known for many years, but their broad significance in the biogeochemical cycling of iron has only been recognized over the past two decades. Aerobic Fe(II)-oxidizing bacteria (FeOB) have been isolated from acidic ecosystems (Thiobacillus ferrooxidans), neutral eutrophic systems (Sphaerotilus natans, Leptothrix ochracea) and neutral oligotrophic systems (Gallionella ferruginea) (Hanert, 1992; Kuenen et al., 1992; Mulder & Deinema, 1992). At neutral pH, Fe(II) is unstable in the presence of oxygen and is rapidly oxidized to the insoluble Fe(III). Hence, the only pH neutral environments where soluble

18 Chapter 2 Fe(II) is available for aerobic FeOB are at interfaces between oxic and anoxic conditions. The aerobic, neutrophilic FeOB (Leptothrix ochracea, Gallionella ferruginea and Sphaerotilus natans) live at such interfaces and are usually associated with the yellowish/reddish ferric deposits formed there.

Over the past several years there has been a growing recognition that other, less readily detectable types of bacteria are involved in Fe(II) oxidation in ecosystems at circumneutral pH. For example, it has been reported that neutrophilic FeOB are abundant at the Loihi seamount hydrothermal vents and play a major role in the Fe(III) oxide deposition (Emerson & Moyer, 2002). Similarly, unidentified neutrophilic, obligate lithotrophic FeOB have been isolated from the rhizosphere of wetlands plants where they are closely associated with deposits of amorphous Fe(III) oxides (Emerson et al., 1999). It was previously believed that Fe(III) oxide deposits associated with sheaths were produced biologically whereas Fe(III) oxide deposits not associated with cells were produced abiotically. Recently, the formation of amorphous Fe(III) oxide in gradient tubes has been attributed to the action of FeOB (Sobolev & Roden, 2001). The authors attribute 90% of the oxidation to biological processes and indicated that the organisms seem to produce a mobile form of Fe(III) that diffuses away from the cells before being precipitated, thereby avoiding encrustation of the cells. They suggest that such soluble Fe(III) complexes might be substrates for closely associated Fe(III)- reducing bacteria. Such an arrangement might allow close coupling between microbial Fe(II) oxidation and Fe(III) reduction within millimeters of the oxic- anoxic interface.

Anaerobic Fe(II) oxidation by phototrophic purple, non-sulfur bacteria utilizing Fe(II) as an electron donor in the light was recognized only a decade ago (Widdel et al., 1993). Subsequently, it was demonstrated that the biological oxidation of Fe(II) in the absence of oxygen is possible by light-independent, chemotrophic microbial activity using nitrate as the electron acceptor (Straub et al., 1996). In

Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 19 addition, studies conducted in gradient cultures revealed that nitrate-reducing strains could also oxidize Fe(II) with molecular oxygen (Benz et al., 1998). Hence, these Fe(II)-oxidizing strains may use nitrate as well as oxygen as electron acceptors. The microbial oxidation of Fe(II) was coupled to stoichiometric reduction of nitrate to N2 and only one strain produced traces of N2O as a by- product (Straub et al., 1996; Benz et al., 1998). The authors proposed the formation of 2-line ferrihydrite as the end product of Fe(II) biooxidation. The chemical reduction of nitrate by Fe(II) requires a catalyst, e.g. at least 10 µM Cu2+, in order to take place at significant rates and may thus be considered insignificant under the conditions applied in our study (Moraghan & Buresh, 1976). The chemical oxidation of Fe(II) with nitrous oxide has not been observed. However, nitrite can oxidize Fe(II) chemically (Moraghan & Buresh, 1977; Straub et al., 1996) but this process is considered insignificant at the conditions applied here. No denitrifying Fe(II)-oxidizing enrichment culture has been found to produce ammonium from nitrate.

Both lithoheterotrophic (depending on organic cosubstrates such as acetate) and strictly lithoautotrophic nitrate-reducing FeOB have been found in various marine and freshwater sediments. However, most isolates depend on organic cosubstrates for cell biosynthesis (Benz et al., 1998). Most probable number estimations showed that denitrifying FeOB accounted for 0.0006-0.8% of the acetate-oxidizing denitrifying microbial population. Lithotrophic FeOB accounted for less than

0.0001% of the total bacterial community. Attempts to isolate CO2-fixing nitrate- dependent FeOB from lithotrophic cultures have failed (Straub & Buchholz- Cleven, 1998). Mixotrophic FeOB accounted for 0.004-0.04% of the total bacterial community. In addition, microbial nitrate-dependent Fe(II) oxidation was demonstrated in a flooded paddy soil as well as in activated sludge from a wastewater treatment plant (Nielsen & Nielsen, 1998; Ratering & Schnell, 2001). Since the activity is not restricted to sunlight exposed habitats, microbial nitrate- dependent Fe(II) oxidation is supposedly more important on a global scale than

20 Chapter 2 anaerobic Fe(II) oxidation by phototrophic bacteria. Furthermore, it has been reported that anaerobic denitrifying FeOB aptly oxidize biogenic Fe(II) minerals formed by bioreduction of synthetic goethite and ferrihydrite and that anaerobic Fe(III)-reducing bacteria readily reduce Fe(III) minerals formed by biooxidation of Fe(II) (Weber et al. 2001; Straub et al., 1998). Hence, autotrophic denitrifying FeOB may play a significant role in the nitrogen and iron cycles in subsurface environments where the nitrate and the Fe(II) zones overlap and organic carbon supply is limited (Figure 2.1).

Figure 2.1. The microbial iron cycle linking the carbon and nitrogen cycles.

Phosphate is released into the environment through natural processes, such as rock weathering and decomposition of dead organic material, and anthropogenic activities, e.g. wastewater effluents and application of manure and fertilizers in horti- and agriculture. In anoxic soils and sediments, phosphate may be sequestered by sorption onto Fe(III) oxides (Williams et al., 1971; Patrick & Khalid, 1974). Phosphate strongly influences the type, morphology and properties of Fe(III) oxides formed by oxidation and hydrolysis of Fe(II) as well as the degree of their transformation (Kandori et al., 1992; Cumplido et al., 2000; Benali et al., 2001). Phosphate may also be retained by precipitation of Fe(II) phosphates such as the monoclinic vivianite (Fe3(PO4)2·8H2O), which is the most important stable Fe(II) orthophosphate solid encountered in the subsurface under most conditions (Nriagu, 1972). At anoxic and circumneutral conditions, the whitish vivianite

Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 21 controls the Fe(II) activity even at very low phosphate concentrations (Nriagu & Dell, 1974). Vivianite occurs as a secondary mineral in the gossans of metallic ore deposits and as a weathering product of primary iron-manganese phosphates in pegmatites (Gaines et al., 1997). Moreover, natural vivianite occurrences have been identified in a number of lake and river sediments (Zwaan & Kortenbout van der Sluys, 1971; Nriagu & Dell, 1974; Postma, 1981; Nembrini et al., 1983; Henderson et al., 1984; Dodd et al., 2003; House, 2003 and references therein). Vivianite is also found in sewage sludge as a result of the wastewater treatment where iron salts are added in order to remove phosphate (Seitz et al., 1973). It is, however, still indefinite how ubiquitous vivianite is in nature. Furthermore, only little is known about the mechanism of vivianite formation and the role played by sedimentary Fe(III) oxides. Anaerobic Fe(III)-reducing microorganisms may reduce Fe(III) oxides thereby releasing the iron as soluble Fe(II) and mobilizing the phosphate adsorbed to the Fe(III) oxides (Lovley, 1997). It has been suggested that vivianite is formed by precipitation following reductive dissolution of Fe(III) oxides (Manning et al., 1981; Manning & Jones, 1982). However, it has also been proposed that the transformation of Fe(III) oxides to vivianite occurs topotactically and not via reductive dissolution (Nembrini et al., 1983). Vivianite was shown to form microbially as a result of the activity of the anaerobic Fe(III)-reducing bacteria, Shewanella putrefaciens, in the presence of high Fe(III)-citrate and phosphate concentrations (Jorand et al., 2000). Moreover, vivianite formation by bioreduction of Fe(III) in hydrous ferric oxide and in smectite has been reported (Fredrickson et al., 1998; Dong et al., 2003).

Only little is known about the oxidation products of vivianite. Metavivianite, a greenish triclinic iron hydroxy phosphate mineral was first described by Ritz et al. (1974) and it was later found to coexist with vivianite in several natural sediment samples (Henderson et al., 1984). Once the Fe(III) content became evident, the true II III composition of metavivianite ((Fe 3-x,Fe x)(PO4)2(OH)x

22 Chapter 2

·(8-x)H2O, x > 1.2)) was established (Rodgers & Johnston, 1985; Rodgers, 1986 and references therein). The formation of intermediate greenish precipitates during oxidation of fluffy, colourless Fe(II) precipitates by anoxic phototrophic microorganisms and nitrate-dependent FeOB have been reported (Ehrenreich & Widdel, 1994; Chaudhuri et al., 2001). Since both studies were conducted in – bicarbonate buffered mineral media (22-30 mM HCO3 , pH 7.0-7.2) containing 3.7-5 mM phosphate, we assume that the initial fluffy, whitish precipitates consisted mainly of vivianite. Chaudhuri et al. (2001) proposed that the intermediate green phases produces by the denitrifying FeOB consist of carbonate green rust (GR-CO3) but no convincing evidence of this biogenic GR-CO3 has been provided yet. The major objective of this work was to examine the Fe-containing products forming during the course of biooxidation of vivianite by non- phototrophic, anaerobic, denitrifying Fe(II)-oxidizing bacteria.

2. 2 Materials and methods

All handling and sampling of solutions and suspensions were carried out under sterile and strict anoxic conditions. All chemicals were p.a. quality.

2.2.1 Microorganisms and media Enrichment cultures of nitrate-reducing FeOB taken from town ditches (Bremen, – Germany) were grown in anoxic, bicarbonate-buffered (30 mM HCO3 , 90%

N2/10% CO2, pH 7.0) mineral media containing 4 mM phosphate as well as essential trace elements and vitamins (Table 2.1; Straub & Buchholz-Cleven, 1998). Ammonium was omitted from the media in order to facilitate detection of ammonium possibly produced by reduction of nitrate. The techniques used for preparation of media and cultivation of bacteria under anoxic conditions have been described by Widdel & Bak (1992). 0.5 M aqueous stock solutions of FeCl2 or

FeSO4 were prepared in 100 mL glass flasks by reacting 65 mmol of iron powder (particle size 10 µm; Merck) with 100 mL deoxygenated 1.0 M HCl or 0.5 M

Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 23

H2SO4, respectively. The solutions were magnetically stirred and heated (~80°C) during reaction until the H2(g) production had ceased (≥ 1 hour). The FeCl2 and

FeSO4 stock solutions were stored under a small Ar overpressure at 5°C.

Table 2.1. Composition of the mineral medium (adopted from Straub & Buchholz-Cleven (1998)).

Components Concentration (M) -3 KH2PO4 1.5·10 -3 K2HPO4 2.5·10 -3 MgSO4·7H2O 1.0·10 -4 CaCl2·2H2O 5.0·10 -5 H3BO3 5.6·10 -6 ZnSO4·7H2O 1.0·10 -6 Na2MoO4·2H2O 4.0·10 -7 CuSO4·5H2O 2.0·10 -6 MnSO4·H2O 1.0·10 -5 Na2SeO4 1.2·10 -6 CoCl2·6H2O 5.0·10 -6 NiCl2·6H2O 8.0·10 NaCl 1.0·10-5 – -2 NaHCO3 3.0·10 -8 Cyanocobalamine (vitamin B12) 3.7·10 p-aminobenzoic acid (vitamin H’) 3.6·10-7 D(+)-biotin (vitamin H) 4.1·10-8 Nicotinic acid (Niacin) 8.1·10-7 -8 Ca-D(+)-pantothenate (vitamin B5) 5.2·10 Pyridoxamine dihydrochloride 9.6·10-7 -7 Thiaminechloridehydrochloride (vitamin B1) 1.5·10 -3 NaNO3 4.0·10

FeSO4 or FeCl2 0.010

2.2.2 Characterisation of precipitates In order to optimize the characterization and distinction between the spectral components, transmission Mössbauer spectra were obtained at temperatures between 5 K and 250 K and in external magnetic fields of 4 T (parallel to the γ-ray direction) using a conventional constant acceleration spectrometer and a source of 57Co in Rh. The spectrometer was calibrated using a 12.5 µm foil of α-Fe at room temperature and isomer shifts are given relative to the centroid of the spectrum of this absorber. The spectra were fitted using simple Lorentzian line shape. Infrared (IR) spectra were obtained using a Perkin Elmer FT-IR 2000 spectrometer and the

24 Chapter 2 KBr pellet technique. Scanning electron microscopy (SEM) was carried out in order to study the morphology and composition of the precipitates. Specimens for SEM were prepared by depositing suspended particles onto an aluminum stub coated with a carbon sticker. The stub was quickly transferred into a sputtering chamber and coated with a thin Pt film (~20 nm). In order to avoid interfering Pt signals in the energy dispersive spectra, the stubs were in some cases not coated with Pt but quickly transferred to the SEM chamber for evacuation. Measurements were performed using a Philips XL30 equipped with a LaB6 source and an accelerating voltage of 20 kV and an EDAX eDXi X-ray dispersive spectrometer.

2.2.3 Biooxidation experiments The biooxidation experiments were conducted in 50-400 mL butyl rubber stoppered bottles with a 90% N2/10% CO2 headspace constituting 10% of the total volume. Prior to inoculation, 4 mM NaNO3 was added as the electron acceptor and 10 mM Fe2+ (as chloride or sulphate) as the electron donor to the mineral media. Control experiments were performed in the same media only they were not inoculated. Addition of ferrous iron to the media induced an immediate precipitation of a solid whitish material. The whitish precipitate was collected on 0.22 µm polyvinylidendifluorid (Durapore, Millipore) filters and stored in an anoxic atmosphere until further measurements. Old outgrown media suspensions that had turned reddish in color due to a precipitate produced by the denitrifying FeOB were used as inocula. Inoculum volume was kept <1% of the culture volume in order to prevent the reddish inoculum from dominating over the initial whitish precipitate. Cultures were incubated in the dark at room temperature and gently agitated once every day. Typical color successions for the media were initial whitish precipitates turning first more and more greenish over time and then finally turning reddish (see Fig 2.2). The color developed uniformly without any indications of multiple phases in the precipitate. At different time intervals suspension samples were withdrawn using 90% N2/10% CO2-flushed polyethylene syringes. The precipitates were collected on 0.22 µm polyvinylidendifluorid

Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 25 (Durapore, Millipore) filters and analyzed by Mössbauer spectroscopy and SEM. Nonfiltered suspension samples were digested in 0.1 M HCl and chemically - + analyzed for Fe(II), NO3 and NH4 .

2.2.4 Analytical methods Fe2+ was determined using a modified phenanthroline method (Fadrus & Maly, 1975). Nitrate was quantified by ion chromatography (Morales et al., 2000) and ammonium was measured photometrically using the indophenol reaction (Rossum & Villarruz, 1963).

2.3 Results and discussion

2.3.1 Identification of solid iron-containing phases In most cases, the mineral media for cultivating denitrifying FeOB contained 10 – - mM FeCl2 or FeSO4, 4 mM NO3 , 4 mM total phosphorus and 30 mM HCO3 at pH - 2- 7.0 ([HCO3 ] = 2138·[CO3 ] at pH 7.0). Whitish flocs precipitated immediately at these initial conditions when Fe(II) was added to the media (Figure 2.2a). Such colourless flocs have been reported to precipitate in similar mineral media (10 mM – Fe(II), 3.7-5 mM phosphate; 22-30 mM HCO3 , pH 7.0-7.2) (Ehrenreich & Widdel, 1994; Chaudhuri et al., 2001). Our Fe(II) measurements showed that 20- 50% of the total Fe(II) added was present in this initial white precipitate.

26 Chapter 2

a

2'/10-"'

Figure 2.2. Colour of suspended material in the growth media during Fe(II) biooxidation. a) Initial whitish precipitate prior to inoculation. b) Inte1mediate greenish phase fo1med within 2-3 days after inoculation. c) reddish precipitate at late stage of biooxidation (>5-6 days).

The precipitates were filtered and investigated by Mossbauer and IR spectroscopy. The transmission Mossbauer spectra obtained for the initial whitish precipitate at temperatures between 20 and 250 K are shown in Figure 2.3. The spectrum measured at 250 K consists of two fairly well-resolved Fe(II) doublets (see parameters in Table 2.2.) The change in line-overlap with decreasing temperature is primarily ascribed to differences in the temperature dependence of the quadrupole splitting of the two components. From the spectra at 10 and 6 K (Figure 2.4), it can be concluded that magnetic ordering takes place between these two temperatures and that only one transition occurs (indicating the presence of only one phase). The parameters of one of the Fe(II) doublets at 250 K (designated B in Table 2.2) are in very good agreement with previously published values for the vivianite Fe(Il)8 site at room temperature (e.g. McCammon & Burns, 1980) and the ordering temperature also agrees well with an assignment as vivianite (Forsyth et al., 1970). However, the second Fe(II) doublet in the initial colourless precipitate (Table 2.2) has parameters that deviate from previously reported values by having a smaller quadrupole splitting (indicating a less distorted coordination) and a significantly higher relative intensity and line width. These effects might be due to the presence of numerous defects in the vivianite crystal lattice, particularly Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 27 affecting the Fe(II)A sites. It should be noted that further components may be added to the fit in order to improve its statistics. Nevertheless, we decided to include no further components as suggested by the finding of one magnetic ordering only. Accordingly, our interpretation of the Mössbauer results for the initial white precipitate suggests a highly defective vivianite, having a distribution of local coordination environments particular in the A site. This assignment is further supported by a major absorption band due to phosphate anions in the infrared spectrum at approximately 1000 cm-1 and the absence of other complex anions (data not shown). Thus, the whitish precipitate is referred to as a vivianite-like (“vivianite”) precipitate.

0,0

0,5

1,0

1,5 20 K 2,0

2,5

3,0

3,5

,0

0

1

2 80 K 3

0,0

0,5

1,0

1,5

Relative absorption (%) Relativeabsorption 2,0 150 K

2,5

3,0

3,5

0,0

0,5

1,0

1,5 250 K

2,0

2,5

-5 -4 - 3 - 2 -1 0 1 2 3 4 5 Velocity (mm/s)

Figure 2.3. Transmission Mössbauer spectra measured between 250 and 20 K of the initial whitish precipitate prior to inoculation (see Fig. 2.2a). Fitting components (and their sum) are shown as full lines.

28 Chapter 2

Table 2.2. Selected Mossbauer parameters of the doublet components in the spectra obtained for different precipitates.

Temperature Isomer shift Quadrupole Line width Area Precipitates 1 1 1 (K) (mms- ) splitting (mms- ) (mms- ) (%) Whitish Fe(II)B 250 1.27 3.09 0.35 38 Fe(II)A 250 1.28 1.81 0.51 62

Dark greenish Fe(II)B 250 1.26 3.05 0.23 17

Fe(II)A 250 1.32 2.38 0.51 44

Fe(III) 250 0.36 0.85 0.40 38 Reddish-orange 40 1.38 2.44 0.94 67 Fe(II) Hyperfine parameters are generally given with uncertainties of 0.03 mms- , the spectral area with an uncertainty of 3%.

1,005 ~ 1,000 .. .;-~ 0,995 ,•• •, 0,990 ' I• • ,. 10 K ..-.. 0,985 •• ~ • e.... • ~. c 0,980 ., 0 • • - ~ 0,975 .E ~ ; c "'nl .b 1,000 g? ~ "It . Qi 0,995 • • • • (J 0:: 0,990 ~ \i~ 6K •. •·f.• 0,985 ~ • 0,980 • -12 -8 -4 0 4 8 12 Velocity (mm.ls)

Figure 2.4. Transmission Mossbauer spectra measured at 10 and 6 K of the initial whitish precipitate prior to inoculation (see Fig. 2.2a).

In general, the color of the media suspension changed from whitish into light green color within 2-3 days after inoculation (Figure 2.2b ). This transformation occurred Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 29 without dissolving the initial whitish precipitate or preserving the whitish precipitate as a separate phase, implying some kind of solid state transformation. Two identical media suspensions were inoculated simultaneously, but at the sampling time they had distinctly different intensities of the green color, designated light and dark green, respectively. Mössbauer spectra of the dark green sample are shown in Figure 2.5. The spectra of this sample are all fitted using three doublet components (two Fe(II) and one Fe(III)) and the parameters of the spectrum measured at 250 K are given in Table 2.2.

0. 0

0. 5

1. 0

1. 5 20 K

2. 0

2. 5

3. 0

0. 0

0. 5

1. 0

1. 5

2. 0 80 K

2. 5

3. 0

3. 5

0. 0

0. 5

1. 0 Relative absorption (%)

1. 5 150 K 2. 0

2. 5

3. 0

0. 0

0. 5

1. 0 250 K

1. 5

2. 0

2. 5

-5 -4 -3 -2 -1 0 1 2 3 4 5 Velocity (mm/s)

Figure 2.5. Transmission Mössbauer spectra measured between 250 and 20 K of the dark greenish precipitate formed during biooxidation. Fitting components (and their sum) are shown as full lines.

No magnetic ordering of the dark green precipitate was observed at temperatures above 20 K, but ordering occurred around 10 K (not shown - due to very thin

30 Chapter 2 samples this was not investigated in details). The two greenish samples had very similar parameters only differing in the relative intensity of Fe(III): 26% and 38% in the light greenish and dark greenish samples, respectively. Assuming the spectral area of a component to be proportional to the abundance of the species in the solid, these results indicate a correlation between the intensity of the green color and the content of Fe(III) in the precipitate. The parameters of the Fe(II) doublets in the dark green precipitate (Table 2.2) were in very good agreement with previously published values for vivianite with a non-negligible Fe(III) content (McCammon & Burns, 1980), whereas the Fe(III) component in particular had a higher quadrupole splitting. The observation that magnetic ordering of both Fe(II) and Fe(III) occurred at similar temperature for the light green phase (Figure 2.6) is a strong indication that they both belong to the same phase. The absence of the component with the low quadrupole splitting in the spectra of both green samples might indicate that the initial vivianite-like phase crystallized into a more well- defined vivianite over time. However, freshly prepared and long-term aged (>1 year) suspensions of the initial vivianite-like precipitate did not differ significantly. Hence, we suggest that the recrystallization of the vivianite-like precipitate can be explained by Fe(II) biooxidation.

Solid State Oxidation ofFe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 31

0 ••• hi 6\,..~ 2 ••••• • 3 ••• ,,• 10 K .-. 4 •. ~ 5 c • ~ 6 • • e- • • 0 7 -2 • nl Q) 0.0 > ~ 0.5 Qi ~~ ~~ 0::: 1.0 ' ' 1.5 .. .. .l\f '".. 6K 2.0 • • •• 2.5 • .....• r ·'• • • • 3.0 '• • • •• 3.5 • - 12 -a -4 0 4 8 12 Velocity (mm/s)

Figure 2.6. Transmission Mossbauer spectrn measured at 6 and 10 K of the light greenish precipitate fonned during biooxidation.

The solid state oxidation of monoclinic vivianite to triclinic metavivianite is well- known (McCammon & Bums, 1980; Pratesi et al., 2003 and references therein). The vivianite crystal structure contains Fe(II) ions in both isolated Fe(II)A and paired Fe(Il)8 octahedra. Mossbauer spectroscopic analyses have shown that the

Fe(Il)8 /Fe(II)A ratio increases with increasing Fe(III) concentration suggesting that the remaining Fe(II)A ions are more readily oxidized than the Fe(II)a ion of an

Fe(Il)8 -Fe(III)8 pair (McCammon & Bums, 1980). The mechanism of oxidation of

Fe(II) in vivianite involves conversion of H20 ligands to OH- ions producing a progressive collapse of the vivianite structure due to the elimination of hydrogen bonds (Moore, 1971). The exact oxidation limits between which the triclinic lattice is stable are somewhat disputed as the results obtained for synthetic and natural vivianites oxidized chemically in the laboratory and naturally oxidized natural vivianite specimens do not agree completely (Rodgers, 1986 and references therein). Taking all reports into account, the monoclinic structure of vivianite is supposedly maintained until 40-50% of total iron is oxidized. Further oxidation 32 Chapter 2 leads to the formation of the triclinic metavivianite in which the FeA site is fully oxidized whereas the oxidation of the FeB ranges from 20% to almost 100%. Thus, the triclinic metavivianite structure persists close to complete oxidation of total iron. The Mössbauer results obtained in this study are consistent with the vivianite solid state oxidation mechanism reported by McCammon & Burns (1980). Thus, we propose that the intermediate greenish precipitate is a metavivianite-like (“metavivianite”) phase. It should be noted that a minor oxidation of dissolved Fe(II) may have occurred even though the solid state oxidation of Fe(II) was predominant.

Within 5-6 days after inoculation, the greenish intermediate was transformed into a reddish product (Figure 2.2c). The magnetically ordered sextet in the spectrum of the red phase (Figure 2.7) measured at 40 K was due to goethite (α-FeOOH) (magnetic hyperfine field of 47.0 T and a quadrupole shift of -0.1 mms-1; cp. Mørup et al., 1983). The sextet deviated from ideal goethite by its asymmetric line shape and its low ordering temperature (around 100 K – data not shown) and, thus, the goethite was poorly crystalline. It is very likely that the presence of phosphate in the media retarded the crystal growth of goethite. The unusual reddish colour of the goethite might also be explained by the presence of phosphate. The spectrum at 40 K was, however, dominated by a Fe(II) doublet (Table 2.2) that ordered magnetically between 40 and 20 K (Figure 2.7). The hyperfine parameters and the magnetic ordering temperature indicated that this component was due to siderite

(FeCO3) having a magnetic ordering temperature of 38 K (Jacobs, 1963). The siderite component may have formed as a result of the microbial activity changing the chemistry of the solution and precipitating a major part of the remaining dissolved Fe(II) at this stage. The characteristic vivianite Fe(II) doublet with large quadrupole splitting was not detected in this sample. The reddish precipitate contained considerably less Fe(III) than the greenish precipitate (only 33 % as estimated from the spectral area). None of the components in the reddish sample

Solid State Oxidation ofFe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 33 could be detected in freshly inoculated samples indicating that carryover of mineral precipitates by inoculation of the culture media was negligible.

0

2

.-. 3 40K ~4 c .Q 5 6 e.0 ~"' 7 g? ""'..!3l 0 Q) a::: 1

2

3

4 '•• 5 (

-12 .a 4 0 4 8 12 Velocity (mm/s)

Figure 2.7. Transmission Mossbauer spectra measured at 20 and 40 K of the reddish precipitate fo1med during the late biooxidation stage. Fitting components (and their sum) are shown as full lines.

Strengite (FeP04·2H20) was not detected at any time during oxidation. Santabarbaraite, a new amorphous F e(III) hydroxy phosphate mineral

(Fe3(P04)i(OH)3·5H20) was reported in a recent study (Pratesi et al., 2003). The brownish mineral was a result of the solid state oxidation of vivianite through metavivianite. However, no Mossbauer data have been provided for this new mineral yet and, therefore, we cannot give an account of whether santabarbaraite forms in our system or not. Thus, goethite was the dominating end product and we propose the reaction path depicted in Figure 2.8 for the nitrate-dependent biooxidation of Fe(II) in our systems. 34 Chapter 2

- - NO3 N2 NO3 N2 Fe3(PO4)2·8H2O II III (Fe 3-x,Fe x)(PO4)2(OH)x·(8-x)H2O α-FeOOH “Vivianite” “Metavivianite” Goethite

Figure 2.8. Proposed reaction path and iron-containing minerals forming during solid state oxidation of vivianite by denitrifying FeOB at the experimental conditions applied in this study.

The biotic formation of layered Fe(II)-Fe(III) hydroxides (green rusts) by anaerobic denitrifying Fe(II)-oxidizing bacteria has been suggested but proper identification of these phases still lacks (Chaudhuri et al., 2001). We cannot rule out that small amounts of green rusts (GRs), perhaps a phosphate intercalated GR (Hansen & Poulsen, 1999), might have been present here during the greenish intermediate “metavivianite” oxidation stage. When present in low concentrations, especially in mixtures including other iron minerals, it is very difficult to identify GRs even with Mössbauer spectroscopy. At least two complementary methods, such as X-ray diffraction (XRD) and Mössbauer spectroscopy, are required for proper identification and characterization of GRs. However, the precipitates collected in this work were poorly crystalline and did not allow for XRD analysis. Electron micrographs including energy dispersive X-ray spectroscopy, suspension colour and mineral stability calculations do not suffice as evidence. Hence, no convincing evidence of GR formation facilitated by denitrifying FeOB has been provided so far. The blue-green colours of metavivianite and green rust minerals originate from Fe(II)-Fe(III) charge transfer between adjacent Fe(II) and Fe(III) ions in edge-shared octahedra (Faye et al., 1968). The greenish suspension colour occurring during the intermediate phase has incited the idea of biogenic GRs in studies on nitrate-dependent Fe(II) biooxidation (Chaudhuri et al., 2001; Lack et al., 2002a&b). However, our results indicate that this reasoning is misleading.

2.3.2 Factors controlling the rate and extent of Fe(II) biooxidation Generally, it was found that maximally 20-64% of the initial Fe(II) amount was oxidized to Fe(III) (Figure 2.9). This indicates some limitations in the accessibility

Solid State Oxidation of Fe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 35 of Fe(II) in the system. Based solely on stoichiometry considerations, the microorganisms are expected to oxidize 5 mol Fe(II) for every 1 mol nitrate reduced to dinitrogen. However, as exemplified in Figure 2.8, this ratio was mostly <4, which can be explained by the consumption of nitrogen as a result of microbial – growth. At initial [Fe(II)]/[NO3 ] ratios <5, nitrate is in excess and should not limit the extent of the biooxidation. Thus, the lack of complete biooxidation could not be due to exhaustion of nitrate. Furthermore, all growth essential nutrients were more than sufficiently applied, hence, the incomplete Fe(II) biooxidation was not caused by lack of nutrients. The most reasonable explanation therefore seems to be that an increasingly limited access to the electron donor over time inhibits complete long- term Fe(II) biooxidation. At least four mechanisms may cause this inhibition: 1) the Fe(II) becomes isolated within the structure of the mixed Fe(II)-Fe(III) minerals forming during biooxidation or underneath a passive Fe(III)-bearing surface film on the initial Fe(II) precipitates, 2) the FeOB cell surface becomes covered with a passive Fe(III)-bearing surface film, 3) the Fe(II) biooxidation is controlled by the rate of dissolution of the initial Fe(II) minerals, or 4) the reaction proceeds primarily by biooxidation of dissolved Fe(II) whose concentration gradually decreases due to changes in solid phase composition. The actual mechanisms whereby the surface-associated Fe(III) can inhibit Fe(II) biooxidation are unknown, but they may involve both kinetic and thermodynamic constraints on the electron transfer. The Mössbauer results obtained in this work strongly suggest that the Fe(II) biooxidation occurred mainly in the solid state of the initial “vivianite” phase. However, we cannot rule out that some dissolved Fe(II) was oxidized as well.

36 Chapter 2

Figure 2.9. Concentration profiles of total Fe(II) and nitrate as a function of time during Fe(II) biooxidation.

No Fe(II) oxidation took place in cultures where nitrate had been omitted, confirming that the microbial Fe(II) oxidation is nitrate-dependent (data not shown). No Fe(II) oxidation was detected in the non-inoculated control experiments within the duration of the experiments and, thus, the chemical oxidation of dissolved Fe(II) by nitrate catalyzed by vitamins or trace elements (e.g. Cu(II)) can be neglected. Ammonium did not form in detectable amounts during Fe(II) biooxidation (data not shown) and therefore dinitrogen was assumed to be the end product as reported previously (Straub et al., 1996; Benz et al., 1998). The absence of ammonium formation does indirectly support the absence of biologically induced green rust formation as synthetic green rust is known to convert nitrate into ammonium in purely chemical reactions (Hansen et al., 1996). It was visually observed that the phosphate concentration in the media exerted a control on the microbial Fe(II)-oxidation. At phosphorus concentrations ≤ 2 mM, no Fe(II)-oxidation took place. However, the solubility product for vivianite (Ksp = 1.71·10-36 at 25°C (Al-Borno et al., 1994)) was still by far exceeded under these conditions. It is not known whether this phosphate limiting effect is due to growth constraints in the mixed bacterial community or whether specific Fe(II) phosphate

Solid State Oxidation ofFe(II) in Vivianite by Anaerobic Denitrifying Fe(II)-Oxidizing Bacteria 37 precipitates are prerequisites of the Fe(II) biooxidation to take place. Experiments are currently underway in our laboratory in order to elucidate the role of specific initial Fe(II) precipitates. It should be noted that the growth of the denitrifying FeOB could not be estimated as they were present in highly heterogeneous suspensions containing both solid iron phases as well as other bacteria (enrichment culture).

2.3.3 Morphology ofsolid iron phases The morphology of the various precipitates was studied by SEM. The initial whitish precipitate consisted of a web-like structure (Figure 2.1 Oa and background in Figure 2.1 Ob) whereas more distinct hexagonally shaped rosettes with particle size ~20 µm (Figure 2.lOb&c) formed during Fe(II) biooxidation. The energy dispersive spectroscopic element analyses showed that, other than iron, the initial whitish precipitate and the rosettes contained mainly phophorus.

Figure 2.10. Scanning electron micro graphs of precipitates fo1med at various stages of the experiment. a. Initial whitish precipitate. b&c. Rosettes observed in the intennediate greenish precipitate. d. Reddish precipitate sampled during the late biooxidation stage. 38 Chapter 2

These observations are interpreted as vivianite forming a web-like morphology in the initial whitish precipitate and partly transforming into hexagonal particles in the greenish colored stage. The interpretations are supported by similar vivianite morphologies reported including pseudo-hexagonal vivianite crystals of low symmetry resulting from microbial Fe(III) reduction of HFO and platy rosettes of vivianite crystals formed during bioreduction of Fe(III) in smectite (Fredrickson et al., 1998; Dong et al., 2003). It was not possible to associate the morphology observed in the reddish precipitate with the minerals identified in this phase (Figure 2.10d).

2.4 Conclusions This work demonstrated that anaerobic, autotrophic, denitrifying Fe(II)-oxidizing bacteria produce poorly crystalline goethite by solid state oxidation of “vivianite” via a “metavivianite” intermediate. The increasing amount of Fe(III) forming in the vivianite structure was accompanied by an increasing intensity in the green colour as the Fe(II) biooxidation progressed. Mössbauer spectroscopic analyses provided no significant evidence of green rust formation. The finding of microbially oxidized vivianite in this study raises the question of the oxidation state of vivianite specimens from natural sediments. Vivianite is generally believed to be an ideal Fe(II) hydroxy phosphate mineral and the presence of Fe(III) is explained by aerial oxidation upon sampling. The results presented here indicate that microbiological processes may be responsible for the oxidation of vivianite and metavivianite in natural subsurface environments.

Acknowledgments We would like to thank Dr. K. Straub for providing and advising us how to culture the nitrate- reducing FeOB. Furthermore, we thank Dr. C. B. Koch for performing the Mössbauer analyses and Dr. D. Mavrocordatos for performing the SEM analyses.

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Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 43

3 Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction

Abstract Many inorganic and organic pollutants may be degraded by microorganisms in the subsurface. However, a wide range of contaminants including chromate, nitrate, radionuclides, nitroaromatic compounds, chlorinated aliphatics and carbamate pesticides may also be chemically transformed by reduction reactions involving layered iron(II)-iron(III)-hydroxides (green rusts). Hence, green rusts (GRs) may play a potentially important role in the fate and transport of pollutants in iron-rich suboxic soils and sediments. Yet, only little is known about the formation of GRs in these environments. The biotic formation of GRs mediated by the anaerobic dissimilatory Fe(III)-reducing bacteria Shewanella spp. has been reported and proposed in several studies. However, the experimental conditions applied were mostly not natural and the evidence of GR formation provided may be questioned. This work investigated the Fe-containing products formed by the facultatively anaerobic Fe(III)-reducing microorganism, Shewanella algae BrY, in culture media containing 4-10 mM formate or lactate and 8-27 mM Fe(III). In order to simulate natural conditions, Fe(III) oxides were applied as coatings on silica (model system for sandy soils) or calcite particles (model system for calcareous soils) and synthetic electron shuttles as well as highly concentrated artificial pH buffers were excluded. S. algae BrY reduced 19-72% of the initial Fe(III) when grown in goethite/calcite, lepidocrocite/calcite or hydrous ferric oxide/sand mineral systems and green or blackish mineral phases were produced within 1-2 weeks after inoculation. Mössbauer spectroscopic analyses indicated that the green and blackish precipitates were dominated by vivianite (Fe3(PO4)2⋅8H2O) and green rust.

44 Chapter 3 3.1 Introduction The significance of bacteria in the biogeochemical cycling of iron has been broadly recognized over the past two decades. Chemical processes were previously considered to account for most of the Fe(III) reduction in subsurface environments. Dissimilatory Fe(III)-reducing bacteria (DIRB) that gain energy by coupling the oxidation of hydrogen or organic compounds to the reduction of Fe(III) oxides have been known for many years, but their biogeochemical importance was acknowledged only a decade ago (reviewed by Lovley, 1997). DIRB transfer electrons to extracellular Fe(III) without assimilating the iron. Fe(III) bioreduction accounts for a major fraction of the carbon oxidation in many different environments and in the presence of high amounts of reactive Fe(III), microbial Fe(III) reduction may even inhibit sulfate reduction and methanogenesis (King, 1990; Lovley & Phillips, 1986). In fact, most of the Fe(III) reduction in the Fe(III) reduction zone of aquatic sediments and aquifers is thought to be enzymatically catalyzed by microorganisms (Lovley et al., 1991). However, the relative importance of microbial and chemical processes involved in the Fe(III) reduction are still somewhat disputed among microbiologists and geochemists.

A wide diversity of Fe(III)-reducing bacteria, which fall in a number of different phylogenetic groups, is known today. Both organisms growing by respiration and by fermentation have been isolated and identified (Lovley, 1991; Nealson & Saffarini, 1994). Hydrogen, short- and long-chained fatty acids, amino acids, sugars and aromatic compounds may serve as electron donors for Fe(III) bioreduction. The enzymes responsible for dissimilatory Fe(III) reduction are outer membrane associated ferric reductases (Lower et al., 2001 and references therein).

Iron reducing bacteria may utilize alternative electron acceptors such as O2, nitrate, S0, sulfate, humic substances, contaminant metals and metalloids as well as chlorinated solvents. The first organism shown to couple respiratory growth to dissimilatory iron reduction was Pseudomonas ferrireductans now known as Shewanella oneidensis but previously classified as Alteromonas putrefaciens and

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 45 Shewanella putrefaciens (Venkateswaran et al., 1999). Various DIRB, including the obligate anaerobic Geobacter sp. and the facultatively anaerobic Shewanella sp., have been isolated from both marine and freshwater sediments, soil, and aquifers (Thamdrup, 2000 and references therein).

The redox potentials of oxidized and reduced iron couples, and thus the energy yield available from Fe(III) reduction, depend strongly on the specific iron phases involved. In soil and aquatic environments, Fe(III) oxides mainly occur in association with other sediment particles as aggregates or coatings. Amorphous and poorly crystalline Fe(III) oxides usually make up 20% or less of the iron content in a sediment (Thamdrup, 2000). They are the main products of abiotic and biotic Fe(II) oxidation in sediments and they constitute the most important phases for microbial Fe(III) reduction. Until recently, it was generally believed that DIRB reduced insoluble Fe(III) oxides only by direct contact with the Fe(III) oxide thereby allowing electron transfer from the cell to the Fe(III) oxide surface. However, over the past several years there has been a growing recognition that DIRB may use different strategies in order to access the solid Fe(III) oxides. These strategies include solubilization of Fe(III) by synthetic or natural Fe(III) chelators and Fe(III) reduction via electron shuttling with soluble humic substances or microbially produced electron shuttles (Nevin & Lovley, 2002 and references therein; Turick et al., 2003). The Fe(III) complexing agents may also stimulate Fe(III) oxide reduction indirectly, by chelation and, thus, removal of Fe(II) from the cell and the Fe(III) oxide surfaces. Both chelated Fe(III) and soluble electron shuttles are more accessible to Fe(III) reductases than solid Fe(III) oxides. In contrast to Geobacter metallireducens, S. algae produces and releases extracellular electron shuttling compounds (Nevin & Lovley, 2000). However, in the absence of soluble electron shuttles, reversible adhesion is required for reduction of solid Fe(III) oxides by S. algae BrY (Das & Caccavo, 2000). Shewanella algae BrY adheres readily and preferentially to a range of solid Fe(III) oxides, such as ferrihydrite, goethite and hematite (Das & Caccavo, 2001). The adhesion

46 Chapter 3 mechanisms are not completely understood, but recent results suggest that the adhesion is mediated by cell surface proteins and independent of cell motility (Caccavo & Das, 2002).

The microbial formation of GRs resulting from bioreduction of various Fe(III) oxides, including ferrihydrite, goethite and lepidocrocite, by strains of the anaerobic DIRB, Shewanella putrefaciens, has been reported repeatedly over the last years (Fredrickson et al., 1998; Kukkadapu et al., 2001; Liu et al., 2001; Parmar et al., 2001; Ona-Nguema et al., 2002a&b; Glasauer et al., 2003). However, no evidence of biogenic formation of GRs at natural geochemical conditions have been offered and it is still unknown whether this process may take place at natural conditions comprising low carbon and iron concentrations as well as the absence of synthetic electron shuttles and highly concentrated artificial pH buffers. GRs are layered iron(II)-iron(III)-hydroxides with anionic interlayers and II III x+ x- they hold the general formula [Fe (6-x)Fe x(OH)12] [(A)x/n·yH2O] , where x = 0.9 - 2- – 2- 4.2, A is an n-valent anion, e.g. CO3 , Cl or SO4 and y is the number of water molecules in the interlayer. In circumneutral solutions the oxidation of dissolved Fe(II) always passes through solid GR phases (Bernal et al., 1959). This agrees with the natural GR occurrences found in suboxic nonacid iron-rich environments such as hydromorphic soils and intertidal sediments (Al-Agha et al., 1995; Trolard et al., 1996; Genin et al., 1998). In addition, GRs have been found as corrosion products in numerous engineering systems, e.g. in a pipeline distribution system for drinking water, steel sheet piles in marine sediments, reinforced concrete (ferro-concrete) and permeable reactive barriers of zero-valent iron implemented for on-site remediation of organic and inorganic contaminants (Tuovinen et al., 1980; Nielsen, 1976; Genin et al., 1991; Roh et al., 2000). Through sequestration and reductive transformation, GRs may play an important role in the fate and transport of organic and inorganic pollutants in suboxic iron-rich soils and sediments (see Chapters 4 & 5 in this work and references therein).

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 47 The major goal of this work was to examine the iron minerals forming during the course of Fe(III) bioreduction of hydrous ferric oxide, goethite and lepidocrocite. Two model systems simulating sandy and calcareous soils in subsurface environments were designed in order to investigate the formation of iron minerals at conditions including low carbon levels, low Fe(III) concentrations applied as Fe(III) oxide coatings on sand or calcite, no electron shuttle and no synthetic pH buffers.

3.2 Materials and methods All handling and sampling of solutions and suspensions were carried out at strict anoxic conditions. Standard sterile techniques were used throughout (Hungate, 1969; Miller & Wolin, 1974). Only the iron oxide coatings were not autoclaved in order to avoid the iron oxides from transforming. Goethite (acicular particles with size 0.1 × 0.6 µm, specific surface area 16 m2/g) and lepidocrocite (acicular particles with size 0.05 × 0.3 µm, specific surface area 18 m2/g) were purchased as fine powders from Bayer (Bayferrox 910 and 943). Calcite (grain size 170-350 µm; Plüss-Staufer AG) and sea sand (dominantly quartz; grain size 0.1-0.3 mm; Riedel- de Haën) were used as Fe(III) oxide bearing minerals.

3.2.1 Preparation of iron oxide coatings Two grams of goethite (goe) or lepidocrocite (lep) and 100 g calcite were combined with 200 mL deionized water (DIW) in a 500 mL polyethylene flask.

Hydrous ferric oxide (HFO) was synthesized by dissolving 4 g Fe(NO3)3·9H2O in 70 mL DIW followed by slow neutralization under magnetic stirring till pH 7 with approximately 30 mL 1 M NaOH (method modified after Schwertmann & Cornell, 1991). The HFO coating was made by combining 100 mL freshly precipitated HFO with 900 mL deionized water and 50 g sea sand in a polyethylene bottle. The suspensions containing the iron oxide coatings were gently agitated on a reciprocating shaker for 24 h and left to stand for another 24 h. Excess Fe(III) oxides and salts were removed from the coated material by repeated decantation

48 Chapter 3 and washing with 0.03 M NaNO3 followed by washing with DIW until clear runoff. Finally, the coatings were collected on folding filters and air dried. The amount of HFO, goethite and lepidocrocite coated onto sand and calcite after washing and drying was quantified to 7-11 mg Fe(III)/g sand or calcite.

3.2.2 Mineral characterisation

The identity and purity of the HFO synthesized were examined by means of X-ray diffraction (XRD) measurements. The XRD analyses were performed on a Scintag XDS 2000 using Co Kα radiation (45 kV, 40 mA) using divergence, scatter and receiving slits of 1°, 0.5° and 0.2 mm, respectively. Samples were scanned between 6 and 80 °2θ with a scan speed of 1 °2θ/min. Mineral suspension samples for transmission Mössbauer spectroscopic analysis were collected on 0.2 µm filters in an anoxic glove box (Coy Laboratory Products Inc.), transferred to Perspex capsules and stored in liquid nitrogen until measurement. Mössbauer spectra were obtained between 250 and 5 K using a conventional constant acceleration spectrometer and a source of 57Co in Rh. The spectrometer was calibrated using a 12.5 µm foil of α-Fe at room temperature and isomer shifts are given relative to the centroid of this absorber. The spectra were fitted using simple Lorentzian line shape and it was assumed that all positions have identical f-factors.

3.2.3 Culture conditions and cell preparation Shewanella algae BrY is a motile gram-negative rod which was isolated first from anoxic estuary sediments (Caccavo et al., 1992). S. algae BrY was grown aerobically in tryptic soy broth (30 g/L CASO-bouillon, Merck) at 28°C on a rotary shaker at 150 rpm for 16-18 h. Cells were harvested by centrifugation (6000 rpm × g, 4ºC, 15 min) during the late exponential – early stationary growth phase at

OD660 ~ 0.6. Optimal Fe(III) reductase activity is expressed at this stage of growth (Roden & Zachara, 1996). The cells were washed twice in oxic 50 mM PIPES [piperazine-N,N´-bis(2-ethanesulfonic acid)] buffer (pH 7.0) and resuspended in culture medium containing no Fe(III) and no carbon source. Washed cell

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 49 suspensions were used as inocula for Fe(III) reduction experiments. Oxygen was expelled from the inoculum by extensive purging with 100% N2(g) (99.99999% purity). Working stock cultures of S. algae BrY were maintained aerobically on tryptic soy agar plates at ambient temperature.

3.2.4 Bioreduction experiments All anaerobic incubations were carried out in anoxic serum vials (25 mL) or test tubes (13 mL) sealed with thick (10-13 mm) butyl rubber stoppers and aluminum crimp caps or plastic screw caps. The basal culture medium (Table 3.1) was prepared according to Kostka & Nealson (1998) but with a phosphate concentration of 2 mM and the exclusion of Fe(II) and EDTA (ethylenediaminetetraacetic acid). The medium was amended with 4-10 mM lactate or formate and 8-27 mM Fe(III). The Fe(III) was applied as Fe(III) oxide coatings on sand or calcite. The suspensions were purged extensively with 100%

N2(g) (HFO/sand suspensions) or 99.5% N2/0.5% CO2(g) (goe/calcite and lep/calcite suspensions) prior to inoculation. The calcareous systems were buffered at pH ~ 7.6 through a natural buffer system (CaCO3(s) + 99.5% N2/0.5% CO2(g)) whereas the sandy systems contained no pH buffer (100% N2(g); pH 5.5-6.0). Inoculum size made up 5% of the total volume. Cultures were incubated dark at room temperature and gently agitated once every day. At different time intervals suspension samples for Fe(II) and Mössbauer analysis were withdrawn from the reaction mixture using 100% N2(g) or 99.5% N2/0.5% CO2(g)-flushed, sterile disposable syringes and hypodermic needles. Suspension samples for Fe(II) analysis were digested in 0.1 M HCl for 30 min.

50 Chapter 3 Table 3.1. Composition of the mineral medium (modified from Kostka & Nealson (1998)).

Components Concentration (M)

(NH4)2SO4 0.0143 -4 KH2PO4 7.3·10 -3 K2HPO4 1.3·10 -3 MgSO4·7H2O 1.0·10 -4 CaCl2·2H2O 5.0·10 -5 H3BO3 5.6·10 -6 ZnSO4·7H2O 1.0·10 -6 Na2MoO4·2H2O 4.0·10 -7 CuSO4·5H2O 2.0·10 -6 MnSO4·H2O 1.0·10 -5 Na2SeO4 1.2·10 -6 CoCl2·6H2O 5.0·10 -6 NiCl2·6H2O 8.0·10 NaCl 1.0·10-5 L-arginine 1.1·10-4 L-serine 1.9·10-4 L-glutamic acid 1.4·10-4 Lactate or formate 4-10·10-3 Fe(III) 8-27·10-3

3.2.5 Analytical methods Fe(II) was determined using a modified phenanthroline method (Fadrus & Maly, 1975). The total amount of Fe(III) coated on calcite and sand was determined by atomic absorption spectroscopy following dissolution in 6 M HCl(aq) for 24 h.

3.3 Results and discussion

3.3.1 Fe(II) production and suspension colour changes Strongly chelating agents such as EDTA were omitted from the culture medium in order to prevent complexation of Fe(II) and Fe(III) which interferes with precipitation of Fe(II) and Fe(II)-Fe(III) mineral phases. Within 1-2 weeks after inoculation Shewanella algae BrY produced green mineral phases in media suspensions containing lepidocrocite and goethite as coatings on calcite and 4-10 mM formate or lactate (Figure 3.1). The formation of the green precipitates was generally slower for the lepidocrocite coating than for the goethite coating. The

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 51 blue-green colours of the phases produced most likely originate from Fe(II)-Fe(III) charge transfer between adjacent Fe(II) and Fe(III) ions in edge-shared octahedra (Faye et al., 1968). Dark brown and blackish products were formed when the bacteria were inoculated on HFO coated sand (Figure 3.2).

Figure 3.1. Culture tubes containing a) goethite and b) lepidocrocite coated calcite in culture medium. The left tubes of the pair were not inoculated whereas the right tubes were photographed 5 months after inoculation with S. algae BrY. Experimental conditions: [formate]0

= 4 mM, [Fe(III)]0 = 8 mM, 99.5% N2/0.5% CO2(g), pH 7.6.

52 Chapter 3

Figure 3.2. Culture tubes containing HFO coated sand in culture medium. Tubes 1 and 2 to the left were not inoculated whereas tubes 3-5 to the right were photographed a) 13 days and b) 21 days after inoculation with S. algae BrY. Experimental conditions: [lactate]0 = 10 mM, [Fe(III)]0

= 25 mM, 100% N2(g), pH 5.5-6.0.

The green and black colours did not change to other colours (observed for >1 year), indicating that the microbial Fe(III) reduction ceased at these mineral stages.

The concentrations of dissolved ferrous iron (Fe(II)sol) estimated during Fe(III) bioreduction were generally low (Figure 3.3). When comparing the final Fe(II)sol amounts produced and the slopes of the Fe(II)sol formation curves for HFO, goethite and lepidocrocite in Figure 3.3, it can be seen, that the final Fe(II)sol amount and the Fe(II)sol production rate both follow the order HFO > goethite > lepidocrocite at similar cell densities, regardless of the carbon source applied. This suggests that bioreduction by S. algae BrY is more facile for HFO than for goethite and lepidocrocite at the experimental conditions employed here. It should be noted that the final Fe(II)sol amounts and the Fe(II)sol production rates reported in this work have not been normalised with respect to the specific surface areas of the iron oxides and coating-bearing solids applied. The reactivity trend is consistent with previous findings demonstrating higher reducibility of natural and poorly

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 53 crystalline Fe(III) oxides as compared to synthetic crystalline Fe(III) oxides (Zachara et al., 1998). The authors ascribed these differences in reducibility to differences in particle size, surface area and crystal defects of the Fe(III) oxides. In some cases, the dissolved Fe(II) concentration decreased again with time (Figure 3.3 b-d). This indicates that the Fe(II) formed was incorporated into solid phases forming and/or adsorbed onto the calcite, sand or Fe(III) oxide surfaces. The solid Fe(II) concentrations were not estimated spectrophotometrically. The solid material was generally low in total iron and, therefore, saving it for Mössbauer spectroscopic analysis was given highest priority.

54 Chapter 3

Figure 3.3. Time course of dissolved Fe(II) production during bioreduction of HFO/sand, goethite/calcite and lepidocrocite/calcite by S. algae BrY. Experimental conditions: [Fe(III)]0 =

25 mM, [formate]0 = 10 mM (a-c) or [lactate]0 = 10 mM (d-f).

No color change and no Fe(II) production were observed in mineral suspensions lacking either a carbon source or S. algae BrY cells (data not shown).

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 55 3.3.2 Identification ofsolid iron phases The purity of the Fe(III) oxides used in the experiments were investigated by transmission Mossbauer spectroscopy (Figure 3.4). Single (or strongly dominating) sextets in the spectra with magnetic hyperfine fields of 48 .1, 50.5 and 45 .5 Tat 5 K demonstrated the purity of the HFO, goethite and lepidocrocite samples, respectively. A minor impurity of goethite in the lepidocrocite sample was resolved in the spectrum measured at 80 K (not shown for pure sample, but can be seen as a magnetically ordered sextet in Figure 3.5c). No Fe(II)-containing components were detected.

a) b)

· 12 -8 -4 4 8 12 ·12 -8 -4 4 8 12 Velocity (mmls} Velocity (mmls}

c)

•

·12 -8 -4 0 4 12 Velocity (mmls)

Figure 3.4. Transmission Mossbauer spectra measured at 5 K of a) HFO (magnetic hyperfine 1 field of 48.1 T, isomer shift of 0.48 1nrns- , negligible quadrupole shift, and line width of outer 1 1 lines 1.10 rmns- ) , b) goethite (magnetic hyperfine field of 50.5 T, isomer shift of 0.49 rmns- , 1 1 quadmpole shift of -0.13 1nrns- , and line width of outer lines 0.42 rnrns- ) and c) lepidocrocite 56 Chapter 3 (magnetic hyperfine field of 45.5 T, isomer shift of 0.50 mms-1, quadrupole shift of –0.01 mms-1, and line width of outer lines 0.60 mms-1) prior to inoculation. Simple Lorenztian fits are shown.

The oxidation state and coordination of Fe in the microbially reduced HFO, goethite and lepidocrocite samples were also examined by transmission Mössbauer spectroscopy (Figure 3.5) The bioreduced HFO, goethite and lepidocrocite samples cultured on formate contained Fe(II) holding similar coordination, as inferred from the similarity of the hyperfine parameters (see legend in Figure 3.5), but different relative intensities (72, 19 and 71%, respectively). The major part of the Fe(III) remaining in the bioreduced samples were coordinated similarly to the Fe(III) present in the initial Fe(III) oxide. The coordination of Fe(II) in the bioreduced lepidocrocite samples cultured on lactate was slightly different (a smaller quadrupole splitting of 2.88 mms-1 for the ferrous component dominates – data not shown). The exact mineralogy of the Fe(II) present in the green phases was not fully resolved, but its coordination is very akin to one of the Fe(II) sites in vivianite (see Chapter 2, this work) and synthetic green rusts (Koch, 1998). These findings agree with other reports on the bioformation of vivianite and green rusts by Shewanella putrefaciens CN32 although the evidence provided may be discussed (Fredrickson et al., 1998; Glasauer et al., 2003; Parmar et al., 2001). Our Mössbauer data on the green phases did not allow for a detailed account of the type of green rust produced. However, when considering solution composition (see Table 3.1) and the high affinity of GR interlayers for carbonate, it is reasonable to assume that carbonate GR was formed (Hansen & Taylor, 1991). Due to the high amounts of Fe(III) in the oxides present in the experiments, it was difficult to probe a possible content of Fe(III) in the vivianite with certainty. The differences in the number of Fe(II) positions in the Mössbauer spectra and particular the different temperatures at which magnetic ordering takes place can be employed in order to distinguish between green rust and vivianite. Preliminary Mössbauer data obtained for the blackish precipitates formed in the HFO/sand suspensions indicate that they hold no resemblance to magnetite eventhough the colour suggests so. On the contrary, the black precipitates seemed to be more similar to synthetic green rusts.

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 57 Mossbauer spectroscopic measurements are currently underway in order to resolve the Fe(II) coordinations in the greenish and blackish phases.

a) b)

c c ,g .Q e- 0 e.0 1l 1l .,"' "' ~ -~ iii • "' Qi • ~ a: •

-12 -8 4 0 4 12 4 -3 -2 -1 0 1 2 4 Velocity (mmts) Velocity ( rmi/s)

c)

-12 -8 4 8 12 v elocity (m mis)

Figure 3.5 Transmission Mossbauer spectra of the black and green phases fo1med within 1-2 weeks after inoculation of a) HFO (measured at 130 K), b) goethite/calcite (measured at 80 K) and c) lepidocrocite/calcite (measured at 80 K) with S. algae BrY. Experimental conditions:

[fo1mate]0 = 4 mM, [Fe(III)]o = 8 mM, 99.5% Ni/0.5% C02(g), pH 7.6. The quadrnpole splittings and isomer shifts for the Fe(II) components in the three systems are: a) 2.93 nnns-1 and 1 1 1 1 1 1.26 mms· , b) 3.08 rmns· and1.31 rmns· and c) 3.22 mms· and 1.32 mms· . Simple Lorenztian fits are shown.

The evidence provided in many of the studies proposing biogenic GRs is not all too convincing, but it strongly suggests the probability of microbially produced GR being present. The challenge encountered is that when present in low 58 Chapter 3 concentrations, especially in mixtures including other iron minerals, it is very difficult to identify GRs using conventional solid phase analysis methods even with Mössbauer spectroscopy. At least two complementary methods, such as XRD and Mössbauer spectroscopy, are required for proper identification and characterization of GRs. However, in this work the solid materials were generally too low in total iron to allow for XRD analysis. Moreover, the highly heterogeneous suspensions were dominated by the coating-bearing sand and calcite solids. Electron micrographs including energy dispersive X-ray spectroscopy, suspension colour and mineral stability calculations do not suffice as evidence. The most convincing evidence provided so far involves an atypical GR-CO3 with an

Fe(II)/Fe(III) ratio of 1 (Ona-Nguema et al., 2002a&b). This GR-CO3 was formed as a result of lepidocrocite reduction by Shewanella putrefaciens CIP 8040 at conditions comprising high nutrient levels (50-75 mM formate), high Fe(III) concentrations (80-300 mM) and a synthetic electron shuttle (100 µM anthraquinone-2,6-disulfonate (AQDS)) at initial pH 7.5. Hence, the results reported during recent years suggest that microbial formation of GR may be possible. The results presented here indicate that GRs may be produced microbially at conditions including low carbon and Fe(III) concentrations as well as the exclusion of synthetic electron shuttles and pH buffers.

3.3.3 Factors controlling the identity of the secondary iron minerals In general, one would expect that biogenic minerals have chemical compositions and crystal habits similar to those produced by nonenzymatic processes as they are governed by the same equilibrium principles. In fact, since the latter stages of mineralization are inorganically driven and the secondary Fe(II)-containing minerals are formed indirectly by electron transfer outside the bacterial cell and not directly inside the bacterial cell, the type of iron mineral formed is a function of the environmental conditions in which the bacteria live, i.e. the same microorganism form different minerals in different environments. The key factors controlling the identity of the secondary iron minerals include medium composition, electron

Formation of Layered Iron Hydroxides by Microbial Fe(III) Reduction 59 donor and electron acceptor concentrations, mineral aging as well as adsorbed ions (Zachara et al., 2002). The main factor controlling the nature of the secondary mineral products are the respiration-driven biogenic Fe(II) supply rate and magnitude and its surface reaction with the residual oxide and other sorbed ions (Zachara et al., 2002). Especially solution and medium composition have a strong impact on the nature of the Fe(II)-containing biomineralization products forming.

Accordingly, siderite (FeCO3) and magnetite (Fe3O4) were the secondary solid phases resulting from the bioreduction of ferrihydrite by Shewanella putrefaciens CN32 in bicarbonate buffered medium (pH 7.1) containing no phosphate, whereas siderite and vivianite were the secondary iron minerals dominating in bicarbonate buffered medium (pH 7.4) containing 4 mM phosphate (Zachara et al., 2002). This is explained by the inhibiting effect of phosphate on crystallization of magnetite (Couling & Mann, 1985; Fredrickson et al., 1998).

3.3.4 Factors controlling the rate and extent of Fe(III) bioreduction In this study, the extent of Fe(III) bioreduction was estimated to 19-72% by transmission Mössbauer measurements. In fact, complete microbial reduction of crystalline Fe(III) minerals has never been observed in laboratory batch culture studies (Roden & Urrutia, 2002). It has been found that Fe(II) does not inhibit Fe(III) reductase activity through an enzyme inhibition mechanism (Roden & Urrutia, 2002). Hence, other chemical and/or physiologic factors control the bioavailability of solid Fe(III) phases and, thus, the extent of their microbial Fe(III) reduction. The initial rate and long-term extent of microbial reduction of amorphous and crystalline Fe(III) oxides, including HFO, goethite and hematite, were linearly correlated with oxide surface area (Roden & Zachara, 1996). Association of biogenic Fe(II) with Fe(III) oxide and DIRB cell surfaces reduced the long-term extent of crystalline Fe(III) oxide bioreduction (Roden & Urrutia, 2002). These results were explained by Fe(II) surface complexes and/or precipitates creating a passive Fe(II)-bearing surface film providing direct physical interference with the electron transfer from the DIRB cells to Fe(III). However, the

60 Chapter 3 real mechanisms whereby the surface-associated Fe(II) inhibits Fe(III) oxide bioreduction are unclear but they most likely involve both kinetic and thermodynamic constraints on the electron transfer. Culture medium composition, in particular the presence and the concentration of phosphate as well as Fe(II) chelating ligands, also exert an influence on the extent of the microbial reduction of Fe(III) oxides. The extent of Fe(III) bioreduction was inhibited by high phosphate concentrations which favoured surface/bulk precipitation processes (Urrutia et al., 1998). The carbon sources most frequently applied in Fe(III) bioreduction studies include malate, citrate and other di- and tricarboxylic acids which are not only easily metabolizable carbon sources but also eminent Fe(II) and Fe(III) chelators. In this study, we employed formate and lactate as carbon sources since they are the weakest complexing agents of Fe(II) and Fe(III) among the C1-

C3 monocarboxylic acids (Martell, 1964). Thus, we expect less dissolution of prevailing precipitates by complexation as compared to other studies.

3.4 Conclusions This work demonstrated that Shewanella algae BrY reduced 19-72% of initial Fe(III) when grown in culture media containing 4-10 mM formate or lactate and 8- 27 mM Fe(III) applied as goethite or lepidocrocite coatings on calcite (pH 7.6) or HFO coatings on sand (pH 5.5-6.0). Within 1-2 weeks after inoculation, green mineral phases were produced in the goethite/calcite and lepidocrocite/calcite mineral systems whereas black precipitates formed in the HFO/sand suspensions. Mössbauer spectroscopic analyses indicated that the greenish and blackish phases most likely were mineral mixtures dominated by vivianite and green rust. Thus, the results indicate that GRs may be produced microbially at conditions including low carbon and Fe(III) concentrations as well as the exclusion of synthetic electron shuttles and pH buffers.

Acknowledgments We would like to thank Dr. R. Gerlach for providing us the Shewanella algae BrY culture and Dr. C. B. Koch for performing the Mössbauer analyses.

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Zachara, J.M.; Fredrickson, J.K.; Li, S.; Kennedy, D.W.; Smith, S.C.; Gassman, P.L. (1998) Bacterial reduction of crystalline Fe3+ oxides in single phase suspensions and subsurface materials. American Mineralogist, 83, 1426-1443.

Zachara, J.M.; Kukkadapu, R.K.; Fredrickson, J.K.; Gorby, Y.A.; Smith, S.C. (2002) Biomineralization of poorly crystalline Fe(III) oxides by dissimilatory metal reducing bacteria (DMRB). Geomicrobiology Journal, 19, 179-207.

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 65

4 Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust: The Effect of Probe Compound Charge

Abstract Layered iron(II)-iron(III)-hydroxides (green rusts) may play an important role in controlling the fate and transport of many organic and inorganic contaminants in iron-rich suboxic soils and sediments. Unlike most other iron oxides, green rusts (GRs) contain not only external Fe(II) reactive sites at the basal planes and at the edges but also internal sites in the space between consecutive Fe(II)-Fe(III) hydroxide layers. The GR interlayer thickness is a function of both the size and the charge of the interlayer anion. Whether a given oxidant has access to the internal sites in GRs is dependent on its charge. We investigated the reductive transformation of nitroaromatic compounds (NACs) by GR-SO4 and studied the effect of NAC charge on the reactivity towards GR-SO4. A series of structurally closely related compounds with different charge properties including nitrobenzene, 4-nitrotoluene, 4-chloronitrobenzene and 4-nitrophenylacetic acid were used as probe compounds. The NACs were completely reduced to their corresponding anilines by GR-SO4. The reactions followed pseudo 1. order kinetics with respect to NAC and the surface area-normalised pseudo 1. order rate constants obtained -4 -1 -2 were 0.16–4.65·10 s ·m ·L at [Fe(II)GR]0 = 1.03-12.60 mM, [NAC]0 = 20-102 µM and pH 8.4-8.6. Neither mass transfer control nor surface saturation kinetics could account for the rather unexpected similarity of the surface area-normalised pseudo 1. order rate constants obtained for the reduction of the neutral and anionic

NACs by GR-SO4. These observations suggest that the anionic NACs did not have an enhanced access to the inner or outer Fe(II)-GR reactive sites as compared to the neutral NACs. Hence, the reaction between NAC and GR-SO4 primarily took place at the edges of GR-SO4.

66 Chapter 4 4.1 Introduction

Layered iron(II)-iron(III)-hydroxides (green rusts) are intermediate phases formed by partial oxidation of Fe(II) or partial reduction of Fe(III). In neutral and weakly alkaline solutions, the oxidation of dissolved Fe(II) always passes through solid green rust (GR) phases (Bernal et al., 1959). This agrees with the natural GR occurrences found in suboxic non-acid iron-rich environments such as hydromorphic soils and intertidal sediments (Al-Agha et al., 1995; Trolard et al., 1996; Genin et al., 1998). In addition, GRs have been found as corrosion products in numerous engineered systems, i.e. a pipeline distribution system for drinking water, steel sheet piles in marine sediments, reinforced concrete (ferro-concrete) and permeable reactive barriers of zero-valent iron implemented for on-site remediation of organic and inorganic contaminants (Tuovinen et al., 1980; Nielsen, 1976; Genin et al., 1991; Roh et al., 2000). Furthermore, the microbial formation of GRs resulting from bioreduction of Fe(III) oxides by strains of the anaerobic dissimilatory Fe(III) reducing bacteria Shewanella putrefaciens has been reported increasingly over the last 5 years (Fredrickson et al., 1998; Kukkadapu et al., 2001; Parmar et al., 2001; Ona-Nguema et al., 2002; Glasauer et al., 2003). Moreover, the biotic formation of GRs by anaerobic denitrifying Fe(II) oxidizing bacteria has been proposed but proper identification of the GR phases still lacks (Chaudhuri et al., 2001). All these indications of microbial GR formation infer the importance of GRs as a link between geochemical and biological processes in natural systems.

II GRs form platy crystals with the general formula [Fe (6- III x+ x- x)Fe x(OH)12] [(A)x/n·yH2O] , where x = 0.9 - 4.2, A is an n-valent anion, e.g. 2- – 2- CO3 , Cl or SO4 , and y is the number of water molecules in the interlayer. The crystal structure consists of positively charged hydroxide sheets with Fe(II) and Fe(III) cations having octahedral hydroxyl coordination. The Fe(III) in the hydroxide layers creates a net positive charge which is balanced by hydrated anions in the interlayers (Figure 4.1). The interlayers have a higher affinity for

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 67 divalent anions than for monovalent anions (Miyata, 1983). Among the 3 most 2- 2- – common GR forms the affinity follows the order: CO3 > SO4 > Cl . The extreme preference shown for carbonate hinders further access and exchange except under certain conditions (Hansen & Taylor, 1991). Non-carbonate forms are readily exchanged with other anions when dispersed in a solution containing the exchanging anion (Mendiboure & Schöllhorn, 1986).

c

a b

Figure 4.1. Green rust layer structure. The hydroxide layers and the interlayers are connected by hydrogen bonds (not shown). The GR-SO4 crystal structure is characterised by the hexagonal unit cell having a = b = 0.55 nm and c = 1.10 nm (Simon et al., 2003). The unit cell consists of one double layer (a double layer is a hydroxide layer and an interlayer), i.e. the hydroxide layer constitutes 0.49 nm and the interlayer 0.61 nm in GR-SO4.

The GR interlayer thickness (extending in the c axis direction; Figure 4.1) is a function of both the size and the charge of the interlayer anion. Tetrahedrally coordinated anions like sulphate lead to larger interlayer distances than smaller monoatomic anions like chloride or planar ions like carbonate (Mendiboure & Schöllhorn, 1986). Not only size but also charge density plays a role for the interlayer spacing. That is, for anions having the same number of valence electrons, anions with smaller ionic radii (higher electron density) are bound more strongly and therefore result in smaller interlayer spacings.

Due to their layered structures, anionic interlayers and high specific surface areas, GRs represent reactive ion exchangers and sorbents of anions, e.g. arsenate, selenate and phosphate (Myneni et al., 1997; Hansen & Poulsen, 1999; Randall et al., 2001). In addition, GR may incorporate heavy metal cations by isomorphic

68 Chapter 4 substitution into the GR hydroxide layers (Tamaura, 1985; Tamaura, 1986). Furthermore, GRs have been shown to reduce a range of inorganic contaminants such as nitrite, nitrate, selenate, chromate, uranyl, pertechnetate and the transition metals AgI, AuIII, CuII and HgII as well as organic pollutants including halogenated ethanes, ethenes and methanes (Hansen et al., 1994; Hansen et al., 1996; Myneni et al., 1997; Erbs et al., 1999; Loyaux-Lawniczak et al., 1999; Cui & Spahiu, 2002; Lee & Batchelor, 2002b; Heasman et al., 2003; O’Loughlin et al., 2003a & 2003b; Pepper et al., 2003; Elsner et al., 2004; O’Loughlin & Burris, 2004). Thus, through sequestration and reductive transformation, GRs may play an important role in controlling the fate and transport of contaminants in suboxic soils and sediments.

In a previous study, the effects of interlayer anion and Fe(II):Fe(III) ratio in GRs on the reduction rate of nitrate were investigated (Hansen et al., 2001). It was found that the rate of nitrate reduction to ammonium increased with increasing Fe(II):Fe(III) ratio and decreased when exchanging a monovalent interlayer anion (chloride) with a divalent anion (sulphate). The results suggest that for anionic oxidants like nitrate, Fe(II) within the hydroxide layer is available from the outside basal planes and from the edges as well as through the interlayer under certain conditions (Figure 4.2). However, oxidants with different charge properties (cations, neutral molecules) may exhibit different affinities for the various reactive Fe(II) sites present in GR.

As the reactive sites are located in/at the Fe(II)-Fe(III) hydroxide layers, the rate of reaction depends on the hydroxide layer area which can be accessed by the oxidant. If the oxidant can exchange with the interlayer anion, reaction can take place both at outer and inner surfaces of the GR particles and, in total, more reactive sites are available for the reaction. However, it was found that nitrate cannot penetrate the interlayer when carbonate or sulphate constitutes the interlayer anions (Hansen & Koch, 1998). This agrees with the fact that the interlayers have a lower affinity for monovalent anions than for divalent anions

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 69 (Miyata, 1983). However, when nitrate was forced into the interlayer by extracting the interlayer sulphate through precipitation of barium sulphate outside the GR particles, the observed 40 fold increase in rate of nitrate reduction almost equalled the increase in exposed surface area of the Fe(II)-Fe(III) hydroxide layers (Hansen & Koch, 1998). From these observations it is expected that the rate of reaction depends on the particular GR form, the crystallite size and the ease with which an oxidant can exchange with An- in the GR interlayer (Figure 4.2). Due to electrostatic interactions, we expect anions to be attracted to the positively charged outer and inner surfaces to a higher degree than cations and neutral compounds. If this theory holds, we may expect oxidants with similar intrinsic reactivity (similar one-electron reduction potentials) to react in the following order: anionic > non- charged > cationic (Figure 4.3), granting that we do not normalise the rate constants with respect to the amount of oxidant sorbed.

Figure 4.2. Reaction of a probe compound at basal planes, at edges and in the interlayer of GR.

The hypothesis only holds in cases where the oxidants possess the same intrinsic reactivities. If the relative reactivities of the probe compounds differ greatly from what would be expected when considering only their reduction potentials,

70 Chapter 4 compound specific effects such as charge properties might explain this and the relative reactivities may follow a pattern like the one depicted in Figure 4.3.

Figure 4.3. Hypothetical plot of observed reaction rate constants for the reactions between cationic, neutral and anionic probe compounds and GR-SO4, assuming that the oxidant charge controls its reactivity towards GR.

In this work, we investigated the reductive transformation of NACs by GR-SO4. Furthermore, the effect of NAC charge on the rate of reaction and the possible access to the internal reactive sites in GR-SO4 were assessed. When quantifying Fe(II) in GRs by means of acid digestion, it is not possible to distinguish between the reactive sites accessible from the outside (at the basal planes or at the edges) or through the interlayer. However, we designed an indirect method to gain insight into the relative importance of the various reactive sites by using a series of structurally closely related compounds with different charge properties as “reactive probes”. Neutral and anionic probes were needed in order to access all Fe(II) reactive sites. According to our hypothesis, cationic and non-charged oxidants should provide information about the reactivity of the outer Fe(II) reactive sites in GR whereas the anionic oxidants should provide information about the reactivity

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 71 of both outer and inner Fe(II) reactive sites. We chose five nitro aromatic compounds (NACs) - representing an important group of reducible organic pollutants - as probe compounds (Figure 4.3). This class of compounds is not only of great environmental concern but also comprises suitable model compounds for studying redox reactions potentially relevant in the environment. Moreover, they react readily with Fe(II) surface species associated with iron oxides or clay minerals transforming them into well-defined, easily detected products allowing mass and electron balances to be established (Hofstetter et al., 2003; Klausen et al., 1995; Schultz & Grundl, 2000). Our main goals were to establish the rate law and estimate the surface area-normalised reaction rates for the reaction of the probe compounds with GR-SO4 in order to assess the importance of the Fe(II) reactive sites accessible through the interlayer relative to the Fe(II) reactive sites accessible at the outer surface in GR-SO4.

4. 2 Materials and methods

All handling and sampling of solutions and suspensions were carried out under strict anoxic conditions. All chemicals were p.a. quality or better. Methanolic stock solutions (5 mM) of nitrobenzene (NB), 4-nitrotoluene (4-NT), 4- chloronitrobenzene (4-CNB) and 4-nitrophenylacetic acid (4-NPA) were prepared in deoxygenated methanol. Several attempts to synthesize the cationic probe compound, 4-(N,N,N-trimethylammonium)-nitrobenzene, failed and therefore the study had to be carried out with only neutral and anionic oxidants. The sulphate GR form was chosen as it is the most stable form and, thus, the easiest to work with in the lab.

4.2.1 Synthesis of GR-SO4

GR-SO4 was synthesized by controlled air oxidation of an FeSO4 solution at a constant pH of 7.00 according to the procedure given by Koch & Hansen (1997).

The GR-SO4 suspension was washed with deoxygenated deionised water and

72 Chapter 4 separated on a folding filter, redispersed in deoxygenated 25 mM Na2SO4(aq) in order to stabilize the GR-SO4 and prevent it from transforming into magnetite spontaneously. Washing, separation and redispersion of the GR-SO4 suspension were conducted in an anoxic glove box (Coy Laboratory Products Inc.). All suspensions and solutions were deoxygenated by Ar-purging (99.9998% Ar; Carbagas).

4.2.2 Mineral characterisation

The identity and purity of the GR-SO4 suspensions were examined by means of X- ray diffraction measurements. The XRD analyses were performed on a Scintag XDS 2000 using Cu Kα radiation (45 kV, 40 mA). Glycerol smears made according to Hansen (1989) were scanned between 6 and 80 °2θ with a scan speed of 1 °2θ/min.

4.2.3 Lyophilization and determination of specific surface area Simple air-drying of the GR mineral in the glove box resulted in big flakes with very low surface areas, hence, a more suitable lyophilization method was adopted from Elsner et al. (2004). The GR-SO4 suspensions were lyophilised using Schlenk-type glassware. The set-up consisted of a 1 L round bottom flask and a 200 mL glass finger connected by a crescent-shaped equipped with an evacuation outlet and a stopcock. All ground joints and fittings were attached using high-vacuum grease. The washed and resuspended GR-SO4 suspensions were filled into the glass finger and the freeze-drying apparatus was assembled and closed before taking it out of the glove box. The suspension was frozen by carefully submerging the lower part of the glass finger into liquid nitrogen for a few hours. Subsequently, the evacuation outlet was connected to a vacuum pump by a metal hose. Following a short evacuation of the metal hose, the lyophilization apparatus was evacuated for several minutes by gently opening the stopcock. The evacuation was terminated by closing the stopcock and disconnecting the vacuum pump. The apparatus position was now reversed by removing the glass finger from

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 73 and immersing the round bottom flask into liquid nitrogen. As any other lyophilization method, this method depends on sublimation of the ice from the frozen sample and its recondensation on a cool surface, in this case the round bottom flask. Generally, it took 1-2 d for the mineral to dry. The apparatus was disassemled in the glove box and the fine powder stored under anoxic conditions.

The specific surface area (SSA) of GR-SO4 was determined by the BET multi- point method using N2 adsorption (Brunauer et al., 1938). Powder samples were filled into sample burettes in the glove box and the generously greased stopcocks closed. Samples and burettes were evacuated prior to connecting them to the BET- instrument (Sorptomatic 1990, Fisons).

4.2.4 Estimation of the one-electron reduction potential for 4-NPA Kinetic experiments in 100 mL Viton stoppered and alu-crimp capped serum vials were carried out under the exclusion of oxygen as described by Hofstetter et al.

(1999). The homogeneous aqueous solutions contained 50 mM KH2PO4 buffer

(pH = 6.60), 5 mM Na2S redox buffer and 20 µM juglone (8-hydroxy-1,4- naphthoquinone) added as deoxygenated 20 mM methanolic stock solution. The solutions were equilibrated at least one day prior to 4-NPA addition. To start the reaction, 50 µM 4-NPA was added as deoxygenated 20 mM methanolic stock solution The vials were agitated on a roller apparatus in the dark at 21ºC. Control experiments were prepared similarly except for the addition of juglone. At different time intervals aqueous samples were withdrawn with a syringe and collected in 1.8 mL HPLC vials containing 100 µL 1 M HCl. The sample vials were sealed with Teflon-coated silicone septa and plastic screw caps and vortexed for 10 s. The samples were stored at -20°C and analysed without further treatment. For comparison, experiments with 4-NT were also conducted. See Supporting Information 7.1 for more information on the one-electron reduction potentials.

74 Chapter 4 4.2.5. Kinetic experiments

All reactions took place at pH 8.4-8.6 where GR-SO4 tends to stabilize and buffer itself. Samples for Fe(II) and XRD analysis were withdrawn prior to reaction. Due to the fast reactions, the experiments were conducted in 10 mL single-use polyethylene syringes (BD Plastipak) in the glove box. To start reaction, 40-200 µL 5 mM methanolic stock solutions of NAC were quickly added to 10 mL GR-

SO4 suspension (1-12 mM Fe(II)GR) washed and resuspended in 25 mM

Na2SO4(aq). A Teflon filter (25 mm x 0.2 µm; BGB Analytik) was quickly mounted on the tip of the syringe and the syringe was vigorously shaken between sampling. At different time intervals, filtered suspension samples were collected in 1.8 mL HPLC vials. The HPLC vials were sealed with Teflon-coated silicone septa and plastic screw caps. The samples were stored at -20°C and analysed without further treatment. Absorption of NAC in the syringe and in the Teflon filter evaluated in blank experiments with NAC added to 25 mM Na2SO4(aq) was found to be negligible.

4.2.6 Analytical methods Initial total and aqueous Fe(II) were determined using a modified phenanthroline method (Fadrus and Maly, 1975). In order to determine [Fe(II)aq] and [Fe(II)total], 1 mL filtered (0.22 µm) and 1 mL unfiltered GR-SO4 suspension samples were withdrawn and each treated with 18 mL 0.1 M HCl for at least 30 min. From these acid digests, 0.1 mL was added to 0.5 mL Fe(II)-reagent and 1.9 mL deionised water (DIW) added up. The Fe(II) content in GR-SO4 was estimated as the difference: [Fe(II)GR] = [Fe(II)total] - [Fe(II)aq]. The NACs and their corresponding intermediates and products formed during reduction by GR-SO4 were quantified by reversed-phase HPLC. Separation was performed on a LiChrospher 100 RP-18 (5 µm, 125 × 4 mm ID) reversed-phase column coupled with a LiChroCART 100 RP- 18 (4 × 4 mm ID) precolumn. Analytical conditions were isocratic and the eluent consisted of 10 mM hydroxylammonium chloride in various DIW/CH3OH mixtures (v/v 35%/65% and pH 7.0 for 4-NT and 4-CNB; 95%/5% and pH 6.0 for

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 75 4-NPA). The injection volume was 20 µLand the flow-rate 1.0 mL/min. HPLC analyses were performed using a Gynkotek High Precision Pump M480, Gynkotek Gina 50 autosampler and a diode array UV detector (340s, Gynkotek). UV-VIS detection was carried out at the wavelengths of maximum absorption for the various nitro aromatic and aniline analytes.

4.3 Results and discussion

4.3.1 Productformation and reaction kinetics The reduction of the aromatic nitro group occurs via nitroso- and hydroxylamino- intermediates where 2 electrons are transferred in each reaction step (Figure 4.4).

0 '\. -::;::-0 -::;::-0 H'\. OH H'\. H N N N/ N"/

2e-, 2H+ 2e-, 2H+ H20 + 2e-, ~ H20 ~ .. ~ ...

R R R Nitrobenzene Nitrosobenzene Hydroxylamine Aniline

Figure 4.4. Reductive transfonnation pathway of NACs .

Thus, in order to reduce 1 Ar-N02 completely to Ar-NH2, 6 electrons corresponding to 6 mol Fe(II) are needed. As magnetite was the major iron phase formed during reaction (XRD results not shown), we assume the following reaction stoichiometry:

The aniline product was not formed at the same rate as the nitro compound degraded which is consistent with the detection of early eluting hydroxylamine intermediates during the course of the reaction (Figure 4.5a & 4.5c). No traces of 76 Chapter 4 nitrosobenzene intermediates or side products such as azoxy-, azo- or hydrazobenzene were found. In Figure 4.5, pseudo 1. order kinetic plots and ln

[Ar-NO2]t/[Ar-NO2]0) versus time plots for the neutral probe compounds, 4-CNB and 4-NT, are shown as examples. The plots for NB and 4-NPA look similar.

Figure 4.5. a. Concentration versus time plots for reaction of GR-SO4 with 4-CNB ([Fe(II)GR]0 =

12.6 mM, [4-CNB]0 = 30 µM). b. ln [Ar-NO2]t/[Ar-NO2]0) versus time plots for reaction of GR-

SO4 with 4-CNB ([Fe(II)GR]0 = 12.6 mM + [4-CNB]0 = 30 µM; [Fe(II)GR]0 = 6.3 mM + [4-

CNB]0 = 50 µM). c. Concentration versus time plots for reaction of GR-SO4 with 4-NT

([Fe(II)GR]0 = 1.31 mM, [4-NT]0 = 20 µM). d. ln [Ar-NO2]t/[Ar-NO2]0) versus time plots for reaction of GR-SO4 with 4-NT ([Fe(II)GR]0 = 1.31 mM, [4-NT]0 = 20 µM; [Fe(II)GR]0 = 1.31 mM, [4-NT]0 = 50 µM). The hydroxylamino intermediate shown in µM equals the deficit in the mass balance and in abs equals the detector response (peak area). Solid lines represent 1. order kinetic fits (a & c) and ln [Ar-NO2]t/[Ar-NO2]0) versus time fits (d) whereas symbols and dotted lines represent actual data.

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 77

At intial Fe(II)GR concentrations in large excess of initial Ar-NO2 concentration, we found a pseudo 1. order rate law for the degradation of Ar-NO2 by GR-SO4:

d[]ArNO a b − 2 k ⋅= [][Fe(II) ⋅ ArNO ] dt GR 2 where a = 1, b = 1 and the observed pseudo 1. order rate constant kobs = k ·

[Fe(II)GR]. At high [Fe(II)GR]0/[Ar-NO2]0 ratios, the nitro compound was transformed completely into the aniline product within reaction duration and the degradation curves of the nitro compound were shaped according to pseudo 1. order kinetics (data points follow solid line in Figure 4.5a). In some instances, i.e. at low [Fe(II)GR]0/[Ar-NO2]0 ratios, the reactions did not follow pseudo 1. order kinetics for the whole duration of reaction (data points deviate from solid line in Figure 4.5c). Hence, in order to allow comparison, all the pseudo 1. order rate constants were calculated as initial rates (i.e. max. first two half-lives) from linear fits of (time, ln [Ar-NO2]t/[Ar-NO2]0)-plots (Figure 4.5b & 4.5d). Surface area- normalised pseudo 1. order rate constants are shown in Table 4.1.

Table 4.1. Surface area-normalised pseudo 1. order rate constants for the reductive transformation of 4-nitrotoluene (4-NT), 4- chloronitrobenzene (4-CNB) and 4-nitrophenylacetic acid (4-NPA) by GR-SO4.

Age GR [Fe(II)GR]0/ ∆[ArNO2] b -1 c -1 -2 d Experiment [Fe(II)GR]0 (mM) [NAC]0 (µM) a f kobs (s ) kobs (s ·m ·L) (d) [NAC]0 (µM)

-4 -5 GR-SO4 + 4-NT 3 1 1.03 20 51.5 10.9 54.5% 7.65·10 6.95·10

-4 -5 GR-SO4 + 4-NT 3 1 1.03 50 20.6 17.4 34.8% 7.41·10 6.74·10

-4 -5 GR-SO4 + 4-NT 3 1 1.03 100 10.3 21.4 21.4% 2.63·10 2.39·10

-4 -5 GR-SO4 + 4-CNB 1 1 1.03 55 18.7 17.7 32.2% 4.21·10 3.83·10

-4 -5 GR-SO4 + 4-CNB 1 1 1.03 102 10.1 16.5 16.2% 2.37·10 2.15·10

-4 -5 GR-SO4 + 4-NPA 1 1 1.03 25 41.2 9.7 38.8% 4.82·10 4.38·10

-4 -5 GR-SO4 + 4-NPA 1 1 1.03 46 22.4 9.9 21.5% 6.37·10 5.79·10

-4 -5 GR-SO4 + 4-NPA 1 1 1.03 100 10.3 13.7 13.7% 1.96·10 1.78·10

-4 -5 GR-SO4 + 4-NT 2 15 1.31 20 65.5 5.4 27.0% 6.74·10 4.82·10

-4 -5 GR-SO4 + 4-NT 2 15 1.31 50 26.2 9.7 19.4% 5.89·10 4.21·10

-2 -5 GR-SO4 + 4-NT 4 2 12.60 50 252 49.1 98.2% 1.10·10 8.17·10

-3 -5 GR-SO4 + 4-NT 4 2 6.30 50 126 42.6 85.2% 1.86·10 2.76·10

-3 -5 GR-SO4 + 4-CNB 2 2 12.60 30 420 29.0 96.7% 9.25·10 6.87·10

-3 -5 GR-SO4 + 4-CNB 2 2 6.30 50 126 38.0 76.0% 1.36·10 2.02·10

-3 -5 GR-SO4 + 4-NPA 2 2 12.60 40 315 37.1 92.8% 5.96·10 4.43·10

-3 -5 GR-SO4 + 4-NPA 2 2 6.30 45 140 27.3 60.7% 1.09·10 1.62·10

a. Amount of NAC reduced by GR-SO4 at reaction termination. b. Fraction of initially added NAC transformed by GR-SO4 at reaction termination. c. Pseudo 1. order rate constants calculated as initial rates, i.e. max. first two half-lives. d. Surface area-normalised pseudo 1. order rate constants. The area of GR-SO4 per L -1 2 -1 suspension was calculated as ¼·[Fe(II)GR]0·600 g·mol ·71.2 m ·g .

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 79 4.3.2 Comparison of rate constants for the different NACs Even for NACs holding very different one-electron transfer reduction potentials

1 ( Eh ' ), their reactivities differed only little in Fe(II)-Fe(III) systems, such as the Fe(II)/goethite system (slope a = 0.6 for linear free energy relationship (LFER)

1 between kobs and Eh ' ; Hofstetter et al., 1999) and the Fe(II)/magnetite system

1 (LFER slope a = 0.34; Klausen et al., 1995). When considering only the Eh ' for the reductive transformation reactions of the NACs applied in this study (Table 4.2), we expect the surface area-normalised pseudo 1. order rate constants for the reduction of the NACs to follow the order 4-CNB > NB > 4-NT > 4-NPA. Based

1 on log kobs versus Eh ' correlations obtained in Fe(II)/goethite systems, we expect 4-

CNB to react 6 times faster than 4-NPA (Hofstetter et al., 1999).

Table 4.2. One-electron reduction potentials and relative reactivities in Fe(II)-magnetite and GR-

SO4 systems for the nitro aromatic probe compounds.

h ´ c,d c,e Compound pKa E 1 (mV) krel (Fe3O4) krel (GR-SO4) 4-Chlornitrobenzene - -450 a 1.22 1.48 Nitrobenzene - -486 a 1 1 4-Nitrotoluene - -500 a 0.57 1.76 4-Nitrophenylacetic acid 3.85 -543 b - 1.23 a. Values from references cited in Hofstetter et al., 1999. b. Estimated at pH 6.60 using a LFER (Hofstetter et al., 1999; see Supporting Information 7.1). c. Reactivity relative to NB. d. Values from Klausen et al., 1995. e. Values from this work.

A comparison of the relative rate constants of the NACs obtained for their transformation by GR-SO4 (this work) and by magnetite (Klausen et al., 1995) shows that they do not differ significantly from each other in any of the mineral systems (Table 4.2). When considering charge effects, we expect the anionic probe compounds to react faster with GR-SO4 than the neutral probe compounds provided that they sorb preferentially within the GR-SO4 interlayers and that Fe(II) in the interlayers are equally or more reactive than external Fe(II) sites. Still, the surface area-normalised kobs values obtained for NB, 4-NT, 4-CNB and 4-NPA under various experimental conditions did not differ significantly from each other

80 Chapter 4 (Figure 4.6; Table 4.1). The anionic probe compound 4-NPA did not react significantly faster with GR-SO4 than the neutral probe compounds NB, 4-NT and 4-CNB. This may indicate that 4-NPA does not significantly interact with reactive Fe(II) sites in the interlayer. Alternatively, the negative charge carried by 4-NPA may be compensating for the lower intrinsic reactivity as compared to the neutral probe compounds, thus, explaining the similarity in rate constants for 4-NPA and the neutral probe compounds. Finally, other factors than intrinsic reactivity or charge of the probe compounds such as regeneration of reactive sites or formation of the magnetite phases may control the overall reactivity of the system.

Figure 4.6. Actual plot of surface area-normalised pseudo 1. order rate constants for the reactions between neutral and anionic probe compounds and GR-SO4.

In heterogeneous reactions, mass transfer in bulk solution becomes the rate- limiting step when the surface reaction is much faster than the diffusion of the reacting species to the reactive surface. In cases where mass transfer controls the overall rate of reaction, the observed pseudo 1. order rate constant, kobs ≥ kL·a, -1 where kL is the calculated mass transfer coefficient (m·s ) and a is the ratio of the external (geometric) specific surface area to volume of solution (m-1) (see

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 81

Supporting Information 7.2). Mass transfer controlled reactions between GR-SO4 particles and the NACs in bulk solution would explain the similar pseudo 1. order rate constants obtained for the NACs in this work. However, when comparing our estimates of kL·a with kobs (see Supporting Information 7.2), we found that the rates of mass transfer for all 4 NACs exceed the observed rate constants by at least 3 orders of magnitude at every initial Fe(II)GR concentration. Thus, the reactions of the given NACs with GR-SO4 are not likely to be mass transfer limited under the experimental conditions applied here.

Since mass transfer in bulk solution does not control the reaction between GR-SO4 and NACs, the overall reaction rate may be surface saturation controlled. During the reductive transformation of NACs, not only the parent compound but also various intermediates forming may compete for the restricted number of reactive sites present in GR-SO4. This competition may constitute the rate limiting step in the overall reactivity and may even be enhanced if the number of reactive sites is depleted during reaction. However, surface saturation kinetics would not explain the unexpected similarity of the pseudo 1. order rate constants obtained for the

NACs but it could explain the bent curves observed at low initial Fe(II)GR concentrations (Figure 4.5d). The kinetically deviating cases at low [Fe(II)GR]0 were evaluated according to Langmuir-Hinshelwood kinetics (see Supporting Information 7.2). Our experimental data did not agree with the Langmuir- Hinshelwood rate law for any of the NACs (regression results not shown). Simplifying the rate law, by assuming that the aniline product or the hydroxylamino intermediate or both did not compete for the reactive sites, did not improve the regression. Thus, the Langmuir-Hinshelwood model cannot explain the deviations from pseudo 1. order kinetics observed at [Fe(II)GR]0 in our GR-SO4 system and it does not suffice as the correct reaction mechanism nor as the rate- limiting step.

82 Chapter 4 If the adsorption follows a saturation-type sorption isoterm (e.g. Langmuir), the sorbate (oxidant) concentration at the surface will vary non-linearily with the total amount of oxidant added. This dependence will have to be taken into account when establishing rate laws for the heterogeneous reactions and when testing the hypothesis that the reaction rates depend on the sorbed concentration of the oxidants. However, at the high reaction rates observed here we could not quantify sorption. Since the measured initial NAC concentrations corresponded to the nominal amount of NAC added, we assume that transformation and not sorption was responsible for the consumption of NAC.

4.3.3 Factors influencing the reaction rate In general, numerous compound- and system-specific factors influence redox reactions. One very important factor is pH which influences the speciation of dissociable compounds as well as the stability of GR and the formation of other iron minerals in the system. pH has a strong impact on the sorption and therefore the availability of ionisable oxidants such as carboxylic acids. At pH ~ 8.4, where our experiments were conducted, 4-NPA (pKa = 3.85) is completely dissociated.

Our experiments conducted with GR-SO4 and NB showed that pH was constant during reaction. In addition, solution pH has an effect on the surface speciation. From other Fe(II)-Fe(III) systems, such as Fe(II) surface species associated with iron oxides or clay minerals, it is well-known that other reactive hydroxylated Fe(II)-Fe(III)-hydroxo surface complexes can form at higher pH (Charlet et al., 1998; Liger et al., 1999). Williams & Scherer (2001) reported a small decrease (5 fold) in the reduction rate of chromate with GR-CO3 when increasing pH from 5.0 to 9.0. This decrease may be due to the alternating speciation of the Fe(II) surface – sites on GR-CO3 and of chromate in solution (pKa (H2CrO4) = 0.8; pKa (HCrO4 ) = 6.5) when raising pH (Williams & Scherer, 2001). In contrast, other studies have reported small increases (4 fold) in the reduction rates of nitrate and trichloroethene with GR-SO4 when increasing pH from 7.1 to 8.4 and from 6.8 to 10.1, respectively (Koch & Hansen, 1997; Lee & Batchelor, 2002b).

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 83

In this work, all experiments were carried out in the presence of 25 mM

Na2SO4(aq) in order to minimize GR-SO4 dissolution and spontaneous transformation into magnetite. Preliminary results from experiments conducted with NB show that the bulk concentration of Na2SO4 has only a very small impact on the rate, i.e. increasing the concentration of Na2SO4(aq) in the GR-SO4 suspension from 5 to 25 mM, reduced the observed rate constant by a factor of 2.

At Na2SO4 concentrations above 25 mM, the effect leveled off and therefore

[Na2SO4] = 25 mM was chosen for this work. Portions of the same GR-SO4 suspension were used for kinetic experiments over a period of two weeks. No significant aging effects, e.g. rate constants decreasing as a function of GR age, were observed within this time frame.

4.3.4 Comparison with rate constants obtained for other Fe(II) containing mineral systems A recent study compared the reactivity of various Fe(II) containing iron mineral systems towards organic probe compounds representing different classes of pollutants (Elsner et al., 2004). The reductive transformation of 4-CNB was investigated for the Fe(III) minerals, goethite (α-FeOOH), lepidocrocite (γ-

FeOOH) and hematite (α-Fe2O3), as well as for the Fe(II)-Fe(III) oxide, magnetite

(Fe3O4). All experiments were conducted in the presence of 1 mM dissolved Fe(II) at pH 7.2. As seen from the surface area-normalised pseudo 1. order rate constants in Figure 4.7, the reduction of 4-CNB by the Fe(II)-amended goethite, lepidocrocite and magnetite systems was up to 40 times faster than its reduction by

GR-SO4. The reduction rate obtained for the Fe(II)/hematite system was only slightly higher than the rate for the blank containing no iron mineral but aqueous Fe(II) solely.

84 Chapter 4

Figure 4.7. Surface area-normalised pseudo 1. order rate constants for the degradation of 4-CNB by GR-SO4 (open square; this work) and various Fe(II) containing mineral systems (solid circles; Elsner et al., 2004). Experimental conditions applied by Elsner et al.: 1 mM aqueous Fe(II), pH 2 7.2, 25 m mineral surface area/L. GR-SO4 = green rust sulphate; α-FeOOH = goethite; Fe3O4 = magnetite; γ-FeOOH = lepidocrocite; α-Fe2O3 = hematite.

The experiments with GR-SO4 in our study were carried out at pH 8.4 whereas the experiments comprising the other systems in Figure 4.7 were conducted at pH 7.2. As the reactivity of GR is expected to increase with pH (Lee & Batchelor, 2002b;

Koch & Hansen, 1997), the lower of GR-SO4 compared to other Fe(II) systems cannot be explained by differences in pH values. GR-SO4 might just contain fewer or less reactive surface sites than Fe(II)-amended goethite, lepidocrocite and magnetite suspensions. These findings contrast those of other studies which found higher surface area-normalised pseudo 1. order rate constants for dechlorination reactions for GR-SO4 than for magnetite (Lee & Batchelor, 2002a & 2002b; Elsner et al., 2004). The different reactivity orders of the Fe(II)-bearing minerals found for chlorinated aliphatics and nitro aromatics suggest that effects other than pH and

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 85 the intrinsic reduction potentials of the reacting species play a role for the reactivity of these Fe(II)-bearing minerals.

4.3.5 Depletion of reactive sites

Assuming that the platy hexagonal GR-SO4 crystals hold an average width of 1 µm and an average particle thickness of 35 nm (Hansen & Koch, 1998), the ratio of outer surface area to total surface area Aouter/Atotal ~ 1/31 (see Supporting

Information 7.3). This means that only 3% of the total surface area in GR-SO4 is available at the external surface. Thus, as the interlayer sulphate in GR-SO4 is not readily exchanged with the anionic NAC applied, we conclude that 4-NPA and other NACs only react with the Fe(II) sites at the external GR-SO4 surface. In

Table 4.3 the actual amounts of NAC reduced by GR-SO4 during reaction is compared with the amount of NAC which theoretically can be reduced by the initial amount of external reactive Fe(II) sites in GR-SO4 at Aouter/Atot ~ 1/31.

Table 4.3. The actual amounts of NAC reduced by GR-SO4 during reaction, ∆[NAC]act, compared with the amount of NAC which stoichiometrically should be reduced by the initial amount of external reactive Fe(II) sites in GR-SO4, ∆[Ar-NO2]theory (calculated as

[Fe(II)GR]0/(31·6) assuming an even distribution of Fe(II) throughout the GR-SO4 structure).

Calculated for [NAC]0 ~ 50 µM.

[Fe(II)GR]0 ∆[Ar-NO2]theory ∆[4-NT]act ∆[4-CNB]act ∆[4-NPA]act (mM) (µM) (µM) (µM) (µM)

1.03 5.5 17.4 17.7 9.9 6.3 33.9 42.6 38.0 27.3

As seen in Table 4.3, the actual amounts of NAC reduced by GR-SO4 during reaction are in most cases higher than the amount of NAC which should be reduced at the given [Fe(II)GR]0 according to reaction stoichiometry. This indicates that new external reactive sites were regenerated, e.g. the Fe(III) phases produced peel off the GR surface exposing new Fe(II) sites or that outermost internal

86 Chapter 4 reactive sites in close vicinity to the edges are available for reaction as well. Lee and Batchelor (2000b) also found the experimentally observed reduction capacity of GR-SO4 for chlorinated ethylenes to be 2-3 orders of magnitudes lower than the estimated reduction capacity including all Fe(II) in GR-SO4.

At low initial Fe(II)GR concentrations, only a fraction of NAC was reduced within the reaction time observed (Figure 4.5c) even though there was stoichiometric excess of Fe(II)-GR present. The fraction of initial Ar-NO2 reduced by GR-SO4 at reaction termination decreased as [Fe(II)GR]0 decreased (Table 4.1) and was accompanied by a change in apparent rate laws with time (compare Figures 4.5b & d). In order to explain these observations, we propose that the NACs react only at external reactive Fe(II) sites and that the regeneration of new external reactive sites is much slower than the reduction of NAC by GR-SO4. Thus, the fast reduction of NAC taking place at the external reactive sites represents the pseudo 1. order behaviour whereas depletion of external reactive sites and their slow regeneration are represented by the second bent part of the (time, ln [Ar-NO2]t/[Ar-NO2]0)- curves deviating from pseudo 1. order kinetics. Hence, at low [Fe(II)GR]0, the regeneration of reactive sites will eventually control the overall reaction rate. Depletion of available Fe(II) was also observed during the fast reduction of chromate by GR-CO3 when the initial chromate concentration was increased or when the GR-CO3 suspension was respiked with chromate repeatedly (Williams & Scherer, 2001).

4.3.6 The role of external and internal reactive sites It is reasonable to assume that GRs hold adsorption properties similar to other layered double hydroxides such as hydrotalcites. The sorption of 2,4,6- trinitrophenol (TNP) and 2,4,5-trichlorophenol (TCP) on chloride and carbonate intercalated hydrotalcites (HT-Cl = Mg3Al(OH)8Cl·yH2O; HT-CO3 =

Mg6Al2(OH)16CO3·yH2O) has been investigated (Hermosin et al., 1993; Ulibarri et al., 1995; Ulibarri et al., 2001). The authors found that the adsorption of TCP on

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 87

HT-CO3 was very low and that TCP adsorbs only on the external surface sites of

HT-CO3 (Hermosin et al., 1993). Furthermore, is was reported that the adsorption of TNP on HT was dramatically affected by the nature of the interlayer anion, i.e. the adsorption of TNP was considerably higher on HT-Cl than on HT-CO3 (Ulibarri et al., 2001). For HT-Cl, interlayer anion exchange of chloride with TNP was detected by XRD analysis and an expansion of the characteristic basal d003 spacing from 7.9 Ǻ to 13.2 Ǻ confirmed the presence of TNP in the HT interlayer (Ulibarri et al., 1995). Collating the results reported for HTs with GRs, it is not 2- likely that the divalent SO4 in GR-SO4 is exchanged with the monovalent 4-NPA. Châtelet et al. (1996) investigated the adsorption of mono- and divalent anions on/in the outer and inner adsorption sites in HT by varying the zetapotential with 2- pH in the presence of various . The authors found that SO4 adsorbs on the external HT surfaces by formation of outer-sphere complexes whereas chloride hardly adsorbed on HT. Moreover, it was reported that the adsorption of sulphate onto HT was not strongly affected by the presence of chloride while sulphate on the contrary inhibited the adsorption of chloride on HT. Studies applying spectroscopic analyses have investigated the sorption of oxyanions at external and internal GR-SO4 surfaces (Myneni et al., 1997; Randall et al., 2001). Selenate was adsorbed only on the outer GR-SO4 surface when added after GR formation whereas it was primarily coprecipitated into the interlayer when present during GR formation. Thus, for selenate, its presence during GR formation is a prerequisite of its incorporation in the GR interlayer. Selenate is readily reduced by GR-SO4 and the rates of reduction of coprecipitated selenate were very similar to the reduction rates of selenate adsorbed at the outer GR surface (Myneni et al., 1997). This finding suggests that the outer and inner reactive Fe(II) sites in GR-SO4 hold similar reactivities.

Results based on electron microscopy reported that the reduction of uranyl took place primarily at the edges of hexagonal GR-SO4 particles (O’Loughlin et al.,

2003a). In another recent study, XRD characterization of the GR-SO4 crystals

88 Chapter 4 during reaction with trichloroacetate (TCA) indicated that TCA did not enter the

GR-SO4 interlayer during reaction (Chapter 5, this work). The average GR-SO4 particle thickness perpendicular to the basal plane was constant during reaction, implying that TCA reacts only at the edges and not at the basal planes. Assuming that the platy hexagonal GR-SO4 crystals hold an average width of 1 µm and an average particle thickness of 35 nm (Hansen & Koch, 1998), the ratio of edge surface area to outer surface area is Aedge/Aouter ~ 1/30 (see Supporting Information

7.3). Hence, only 3% of the outer surface area in GR-SO4 is available at the edges. Once more the regeneration of new external reactive sites is strongly inferred as the actual amounts of NAC reduced by GR-SO4 during reaction are much higher than the amount of NAC which may be reduced by the reactive edge sites present initially. Assuming that the NACs react at the edges only and if employing the

Aedge/Aouter in the estimation of the rate constants, the surface area-normalised pseudo 1. order rate constants for GR-SO4 would be 30 times higher than the rate constants depicted in Figure 4.7. Thus, the reactivity of GR-SO4 normalised to its reactive surface area is higher than the reactivity normalised to its outer surface area determined by the BET method (N2 adsorption).

The reduction of chromate has been examined in the presence of all the common GR forms (Bond & Fendorf, 2003; Loyaux-Lawniczak et al., 1999; Loyaux- Lawniczak et al., 2000; Williams & Scherer, 2001). The results reported by Bond & Fendorf (2003) confirm that not only the surface area of GR but also the interlayer spacing (interlayer anion size) and interlayer anion charge play an important role for the reaction rate. Hence, it follows that coordination (size) and charge of the oxidant determine its access to the internal sites in GRs.

The results obtained for all 4 NACs support what has been reported for nitrate and - TCA. At [Fe(II)GR]0 = 2-10 mM and [NO3 ]0 = 14.3 mM, pseudo 1. order rate -7 -1 -2 constants for the reduction of nitrate by GR-SO4 were 1.58·10 s ·m ·L (Hansen et al., 2001). This reaction rate increased 40 times by adding barium nitrate instead

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 89 of nitrate, thereby precipitating the interlayer sulphate as barium sulphate and enhancing access to the interlayer. Though barium addition changes the GR-

SO4 system dramatically, it indicates the importance of interlayer anion exchange (Hansen & Koch, 1998). The rate constant reported for nitrate (no barium added) is 100-1000 smaller than the rate constants obtained for the NACs in this work. Moreover, the reaction kinetics for nitrate did not deviate from pseudo 1. order kinetics. At [Fe(II)GR]0 = 0.25-10.4 mM and [TCA]0 = 50 µM-1 mM, pseudo 1. -7 order rate constants for the reduction of TCA by GR-CO3 or GR-SO4 were 6.5·10 s-1·m-2·L (Chapter 5, this work). The rate constant for TCA is 10-1000 smaller than the rate constants for the NACs and the reaction kinetics for TCA did not deviate from pseudo 1. order kinetics. This suggests that the overall reductive transformation of slowly reacting oxidants such as nitrate and TCA is not controlled by the rate of regeneration of external Fe(II) reactive sites. Altogether, the results reported for selenate, chromate and nitrate clearly demonstrate that these anionic oxidants react primarily with external reactive sites in GR-SO4. Only under certain conditions, i.e. adding the oxidant prior to GR-SO4 formation or extracting the interlayer sulphate through precipitation with barium outside the

GR-SO4 particles, do the oxidants have access to the interlayer. Our findings suggest that both the neutral and anionic nitro aromatic probe compounds applied here also react exclusively with the external reactive sites in GR-SO4. Supposedly, the neutral and monovalent charge states of the NACs hinder their access to the

GR-SO4 interlayer. A divalent anionic nitro aromatic probe compound might exchange with the interlayer sulphate more readily and gain access to the inner

Fe(II) reactive sites in GR-SO4, only divalent anionic NACs are not commercially available.

4.4 Conclusions This work demonstrates that NACs are completely reduced to their corresponding anilines by GR-SO4. The surface area-normalised pseudo 1. order rate constants obtained for the reduction of the neutral and anionic NACs by GR-SO4 under

90 Chapter 4 various experimental conditions did not differ significantly from each other despite their different charges. Neither mass transfer control nor surface saturation kinetics could account for the similarity of the pseudo 1. order rate constants obtained for the NACs. These observations suggest that the anionic NACs do not have an enhanced access to inner or outer Fe(II)-GR reactive sites as compared to the neutral NACs. Based on our estimations of the molecular sizes of the NACs, we propose that the charge and not the size of the NACs controls their access to the internal reactive sites in GRs. Hence, the reaction between NAC and GR-SO4 takes place primarily at the external reactive Fe(II) sites. This work further demonstrated that the reduction of the NACs by GR-SO4 only followed pseudo 1. order kinetics throughout the whole reaction at high initial Fe(II)GR concentrations. At low initial

Fe(II)GR concentrations, the NACs were not reduced completely within the reaction time observed though, according to reaction stoichiometry, the total Fe(II)-GR present should be sufficient to reduce the whole amount of NAC. This means that at some point during the reaction the external reactive Fe(II) sites were depleted and the regeneration of new external reactive sites was much slower than the reduction of the NACs by GR-SO4. The reduction of 4-CNB by GR-SO4 reported here was 10-100 times slower than its reduction by other Fe(II)-Fe(III) systems such as goethite, lepidocrocite and magnetite suspensions amended with Fe(II) (Elsner et al., 2004).

The results obtained in this work infer that under natural geochemical conditions where GR-SO4 presumably forms in low concentrations, the rate of regeneration of external Fe(II) reactive sites may control the overall reductive transformation of fast reacting pollutants by GR-SO4. Thus, not only abiotic processes such as interchanging redox conditions created by water level alterations but also the direct microbial formation through Fe(III)-reducing bacteria may govern the formation of GRs and the renewal of external Fe(II) reactive sites in GRs. This holds both for natural systems like iron-rich suboxic soils and sediments as well as engineered

Reduction of Nitroaromatic Probe Compounds by Sulphate Green Rust 91 systems like permeable reactive barriers of zero-valent iron implemented for on- site remediation of organic and inorganic contaminants.

Acknowledgments We would like to thank Henrik T. Andersen for performing the NB kinetic experiments, and Hanne Nancke-Krogh for technical assistance in the laboratory.

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Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 97

5 Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems

Abstract Trichloroacetate (TCA) is a widespread environmental contaminant with proven phytotoxicity and suspected human carcinogenicity. In order to assess the global cycling of TCA and to predict its fate in subsurface environments, information regarding the reactivity and product distribution of TCA degradation is needed. Due to the high oxidation state of TCA, conditions for oxidative transformation pathways in soils and groundwater are unfavorable. However, in suboxic soils and sediments, Fe(II)-bearing minerals are potential reactants for reductive dehalogenation reactions of TCA as has been demonstrated for other halogenated contaminants. We examined the reactivity of various Fe(II)-Fe(III) mineral systems towards TCA and dichloroacetate (DCA), its expected transformation product, in laboratory batch experiments imitating natural conditions, i.e. low initial Fe(II), Fe(III) and TCA/DCA concentrations and no artificial buffer. The

Fe(II)-Fe(III)-systems investigated included sulfate green rust (GR-SO4), carbonate green rust (GR-CO3), magnetite, Fe(II)/goethite and Fe(II)/lepidocrocite. Trichloroacetate was readily reduced to DCA by all Fe(II)-bearing minerals. The reactions generally followed pseudo 1. order kinetics with respect to TCA. The surface area-normalised pseudo 1. order rate constants obtained (0.35–7.6·10-5 min- 1 -2 ·m ·L at [Fe(II)]0 = 0.20–12.2 mM, [TCA]0 = 15–1000 µM and pH 7.0–8.7) showed no striking differences regarding product distribution and surface area- normalised reaction rate constants between the Fe(II)-Fe(III)-systems. The stoichiometrically formed DCA was not further reduced to monochloroacetate (MCA) or acetate in any of the systems within the time frame in our experiments. To our knowledge, this is the first published report on abiotic transformation of TCA by Fe(II)-bearing minerals. Our results imply that processes involving reactive Fe(II)-bearing minerals may play a significant role in controlling the fate

98 Chapter 5 of TCA in natural subsurface environments and that DCA found in the subsurface may be formed by such processes.

5.1 Introduction Trichloroacetic acid (TCA) has been applied as a herbicide for many years until its use was banned in the late 1980´s (Berg et al., 2000). Today, TCA is mainly used as an etching agent in the metal industry, as a swelling solvent in the plastic production and as a bleaching agent in the paper and pulp manufacture (Müller et al., 1996). Other anthropogenic sources include formation of TCA as a result of the chlorine based disinfecting process used in drinking water treatment and the atmospheric photooxidation of chlorinated solvents including tetrachloroethene and 1,1,1-trichloroethane (McCulloch, 2002). Only very little information is available on the TCA production volumes and even less is known about the amount of TCA released into the environment as a result of its industrial applications. Due to its low volatility and high aqueous solubility, TCA is easily washed out of the atmosphere into the aquatic and terrestrial biospheres. As TCA is found in almost every ecosystem around the globe, including non-urban and non-industrial sites, the relative contributions from anthropogenic and natural sources are currently being debated (McCulloch, 2002; Ahlers et al., 2003). Trichloroacetic acid is omnipresent in soils and the concentrations reported are very variable ranging from <0.05 µg/kg to 380 µg/kg (Euro Chlor, 2001; McCulloch, 2002; Ahlers et al., 2003). Both abiotic and enzymatically catalyzed formation of TCA from humic acids have been demonstrated in laboratory studies (Haiber et al., 1996; Hoekstra et al., 1999b; Fahimi et al., 2003). Furthermore, the in situ natural formation of TCA from anthropogenic or natural tetrachloroethene or 1,1,1-trichloroethane in biota has been suggested (Hoekstra et al., 1999a; McCulloch, 2002). Such natural sources may explain part of the TCA concentrations found in soils but their environmental significance is still unknown.

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 99 On account of its phytotoxicity, suspected human carcinogenicity and widespread occurrence, TCA is of considerable environmental concern especially in the terrestrial compartment. The TCA concentrations found in soil, air and water in pre-industrial times were far below the present ones (Jordan & Frank, 1999; Ahlers et al., 2003). Based on the current TCA concentrations detected in soils, the European Commission proposed risk reduction measures concerning tetrachloroethene - a precursor of TCA - to be taken immediately (Ahlers et al., 2003 and references therein). Occurrences of monochloroacetic acid (MCA) and dichloroacetic acid (DCA) reported include surface waters, marine waters, precipitation, ice (glaciers) and air (Reimann et al., 1996; Berg et al., 2000; Scott et al., 2000; Scott et al., 2002). Based on the concentrations reported for the aquatic environments, it is reasonable to assume that MCA and DCA are omnipresent in soils as well. Sources of MCA and DCA include production in the chemical industry, photooxidation of chlorinated aliphatics in the atmosphere and reductive transformation of TCA (Reimann et al., 1996; Ahlers et al., 2003 and references therein). MCA and DCA are also toxins and suspected human carcinogens (Kühn & Pattard, 1990), hence, not only TCA but also its daughter compounds are pollutants of environmental concern.

In subsurface environments TCA may be removed by sorption, seepage, chemical transformation, microbial degradation, and plant uptake followed by metabolic degradation and/or physical removal at harvest (Foy, 1975). There is little or no evidence of abiotic transformations of TCA in the literature. Only one recent study demonstrated the reductive dechlorination of TCA to MCA by Fe(0) (Hozalski et al., 2001). It has been reported that the degradation of TCA in soil is slow and mainly mediated by microorganisms but only little is known about the bacteria and processes involved (Lignell et al., 1984). Biodegradation of TCA has been found at both oxic and anoxic conditions. An aerobic microorganism capable of growing on TCA as the sole carbon and energy source has been characterised (Yu & Welander, 1995). Moreover, anaerobic bacteria coupling co-metabolic growth to reductive

100 Chapter 5 dechlorination of TCA have been isolated (Weightman et al., 1992; De Wever et al., 2000). However, more information regarding the abiotic and biotic transformation of TCA is needed in order to assess the fate and transport of TCA in natural subsurface environments.

It is well-known that Fe(II) present in minerals or associated with mineral surfaces is a much stronger reductant than Fe(II) in solution. The enhanced reactivity of a structural or surface-bound Fe(II) center can be rationalized by the increased electron density donated by hydroxyl ligands and a stabilization of the Fe(III) oxidation state by the hydroxyl ligands (Luther, 1990). Fe(II)-bearing minerals including layered Fe(II)-Fe(III) hydroxides (green rusts), magnetite (Fe3O4), siderite (FeCO3), Fe(II) sulfides as well as Fe(II)-carrying Fe(III) oxides and clay minerals have also been shown to reduce a range of organic and inorganic contaminants such as nitro aromatic compounds, chlorinated aliphatics, chromate, uranyl, pertechnetate, nitrate, monochloramine and carbamate pesticides (Chapter 4, this work; Klausen et al., 1995; Cui & Eriksen, 1996; Erbs et al., 1999; Liger et al., 1999; Loyaux-Lawniczak et al., 1999; Amonette et al., 2000; Hansen et al., 2001; Pecher et al., 2002; Vikesland & Valentine, 2002; Hofstetter et al., 2003; O’Loughlin and Burris, 2003; O’Loughlin et al., 2003a & 2003b; Strathmann & Stone, 2003; Elsner et al., 2004a). Laboratory and field studies showed that even in geochemically highly heterogeneous anoxic aquifer sediments, Fe(II) adsorbed to Fe(III) (hydr)oxide surfaces was the dominant reductant of nitroaromatic and halogenated contaminants (Rügge et al., 1998; Hofstetter et al., 1999; Kenneke & Weber, 2003). Only little is known about the nature of the Fe(II) species associated with Fe(III) oxide surfaces but reactive hydroxylated Fe(II)-Fe(III)-hydroxo surface complexes associated with hematite and magnetite above pH 6.5 have been proposed (Charlet et al., 1998a&b; Liger et al., 1999). Due to the presence of structural Fe(II) within the mineral lattice, the reactivity of Fe(II) associated with mixed valent Fe(II)-Fe(III) minerals such as green rusts, magnetite and reduced ferruginous clay minerals may hold another reactivity than Fe(II) associated with

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 101 pure Fe(III) oxides. However, Fe(II) adsorbed on Fe(III) oxides such as goethite, hematite and lepidocrocite may also hold different reactivities as the Fe(III) oxides contain different crystal and surface structures.

Since chlorinated ethanes and ethenes, such as hexachloroethane, 1,1,1- trichloroethane, tetrachloroethene and trichloroethene, are susceptible to chemical reduction by a range of Fe(II)-bearing minerals including magnetite, GR-SO4, Fe(II) sulfides and Fe(II)-carrying Fe(III) oxides (Butler & Hayes, 1998 & 1999; Hwang & Batchelor, 2000; Gander et al., 2002; Lee & Batchelor, 2002a&b; Elsner et al., 2004a), we hypothesized that TCA may be transformed by Fe(II)-bearing minerals as well. The main goals of this work were to study such reactions and establish product distribution and surface area-normalised reaction rates for the reductive dechlorination of TCA by Fe(II)-Fe(III) mineral systems common in nature.

5. 2 Materials and methods No synthetic buffers were applied and iron concentrations were kept low. The calcareous systems were pH-controlled at 7.6 through a natural buffer system

(CaCO3(s) + 99.5% N2/0.5% CO2(g)). All handling and sampling of solutions and suspensions were carried out under strict anoxic conditions. Goethite (acicular particles with size 0.1 × 0.6 µm, specific surface area 16 m2/g) and lepidocrocite (acicular particles with size 0.05 × 0.3 µm, specific surface area 18 m2/g) were purchased as fine powders from Bayer (Bayferrox 910 and 943). Calcite (grain size 170-350 µm; Plüss-Staufer AG) was used as a buffer or as a Fe(III)-oxide-bearing mineral. In order to simulate natural conditions, the iron minerals were applied as coatings on calcite particles (model system for calcareous soils) in some experiments. Trichloroacetic acid, dichloroacetic acid and monochloroacetic acid were p.a. quality (Fluka).

102 Chapter 5 5.2.1 Synthesis of GRs and magnetite

GR-CO3 was synthesized by controlled air oxidation of an FeCl2 solution at a constant pH of 7.00 (titrated with 1 M Na2CO3) according to the procedure given by Hansen & Koch (1997). 0.5 M aqueous stock solutions of FeCl2 were prepared in 100 mL glass flasks by reacting 65 mmol of iron powder (particle size 10 µm; Merck) with 100 mL deoxygenated 1.0 M HCl. The solutions were magnetically stirred and heated (~80°C) during reaction until the H2(g) production had ceased (≥

2 hours). The FeCl2 solutions were stored in the dark under a small Ar overpressure at 5°C. The GR-CO3 suspensions were washed with deoxygenated deionised water (DIW), separated on a folding filter (medium filtration rate; cotton linter/high alpha pulp; Schleicher & Schuell) and redispersed in deoxygenated

DIW. Washing, separation and redispersion of the GR-CO3 suspension were conducted in an anoxic glove box (92% N2/8% H2; Coy Laboratory Products Inc.). All suspensions and solutions were deoxygenated by Ar-purging (99.9998% Ar;

Carbagas). Magnetite was synthesized by further aerial oxidation of GR-CO3 at pH

7.00 until consumption of 1 M Na2CO3 ceased. GR-SO4 was synthesized by controlled air oxidation of an FeSO4 solution at a constant pH of 7.0 according to the procedure given by Koch & Hansen (1997). The GR-SO4 suspension was washed with deoxygenated DIW, separated on a glass filter funnel (pore size 4; Duran) and redispersed in deoxygenated DIW. Washing, separation and redispersion of the GR-SO4 suspension were conducted in an anoxic glove bag (99.9995% Ar; Aldrich).

5.2.2 Preparation of iron oxide coatings Two grams of goethite (goe) or lepidocrocite (lep) and 100 g calcite were combined with 200 mL DIW in a 500 mL polyethylene flask. The suspension was gently agitated on a reciprocating shaker for 24 h and left to stand for another 24 h. Excess Fe(III) oxides and salts were removed from the coated material by repeated decantation and washing with DIW in polyethylene flasks until clear runoff. Finally, the coatings were collected on folding filters and air dried. The amount of

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 103 goethite and lepidocrocite coated onto calcite after washing and drying was quantified to 10-11 mg Fe(III)/g calcite.

5.2.3 Mineral characterisation

The identity and purity of the GR-CO3, GR-SO4 and magnetite suspensions were examined by means of X-ray diffraction (XRD). The XRD analyses were performed on a Scintag XDS 2000 using Co Kα radiation (45 kV, 40 mA) or a Siemens D5000 XRD applying Co Kα radiation (40 kV, 40 mA). Glycerol smears made according to Hansen (1989) were scanned between 6 and 80 °2θ with a scan speed of 1 °2θ/min. The specific surface area (SSA) of calcite was determined by the BET multi-point method using N2 adsorption (Brunauer et al., 1938). Powder samples were filled into sample burettes in the glove box and the generously lubricated stopcocks closed. Samples and burettes were evacuated prior to connecting them to the BET-instrument (Sorptomatic 1990, Fisons).

5.2.4 Kinetic experiments All reactions were carried out in 25-100 mL serum vials sealed with stoppers (Viton or Teflon coated rubber) and aluminum crimp caps. Kinetic experiments were conducted with GR-SO4, GR-CO3, magnetite, Fe(II)/goethite and Fe(II)/lepidocrocite at room temperature. In most cases pH was controlled through the carbonate-bicarbonate buffer system by adding calcite to suspensions containing the iron minerals solely or by adding the iron minerals as coatings on calcite. Furthermore, the calcite containing suspensions were deoxygenated with

0.5% CO2/99.5% N2(g) thereby attaining an initial pH of 7.6-7.7. The GR-CO3 and magnetite suspensions were deoxygenated with 100% N2(g) and no additional pH buffer was added. The goethite and lepidocrocite suspensions were amended with

300-1000 µM FeCl2(aq) and equilibrated > 20 h prior to TCA/DCA addition. See Table 5.1 for more details on the experimental conditions. To start the reaction, 50 µM - 1 mM TCA or DCA was added to the mineral suspensions from aqueous anoxic stock solutions. The reaction vials were agitated gently on a roller apparatus

104 Chapter 5 or a shaking table (35 rpm) in order to minimize abrasion of the iron oxide mineral coatings. At appropriate time intervals, suspension samples were withdrawn using

Ar(g)-, 100% N2(g)- or 99.5% N2/0.5% CO2(g)-flushed sterile disposable syringes and hypodermic needles. The suspension samples were filtered (0.2 µm; Teflon) and collected for quantification of chloride and the chlorinated acetic acids. The samples were stored at -20°C and analysed without further treatment.

5.2.5 Analytical methods Total and aqueous Fe(II) were determined using a modified phenanthroline method

(Fadrus and Maly, 1975). For determining [Fe(II)aq] and [Fe(II)total], 1 mL filtered (0.2 µm; Teflon) and 1 mL unfiltered mineral suspension were added to 18 mL 0.1 M HCl, respectively, and allowed to dissolve for 30 min. From these acid digests, 0.1 mL was added to 0.5 mL Fe(II)-phenanthroline-buffer-reagent and 1.9 mL DIW added up. Estimates of the structural or adsorbed Fe(II) in the Fe(II)-Fe(III) mineral systems were estimated as the difference [Fe(II)solid] = [Fe(II)total] -

[Fe(II)aq]. The total amount of Fe(III) coated on calcite was determined by atomic absorption spectroscopy following dissolution in 6 M HCl(aq) for 24 h. At low initial TCA concentrations (≤ 50 µM), the chlorinated acetic acids were quantified by means of a modified ion interaction (or paired-ion) chromatographic method (Sarzanini et al., 1999). Separation was performed on a LiChrospher 100 RP-18 (5 µm, 125 × 4 mm ID) reversed-phase column coupled with a LiChroCART 100 RP- 18 (4 × 4 mm ID) precolumn. Analytical conditions were isocratic and the eluent consisted of 50% aqueous solution of 3.5 mM cetyltrimethylammonium chloride

(pH 5.0) and 50% CH3CN. The injection volume was 20 µL and the flow-rate 1.0 mL/min. HPLC analyses of the chloroacetates were performed using a Gynkotek Pump M480, Gynkotek Gina 50 auto sampler and a diode array UV detector (340s, Gynkotek). UV-VIS detection was carried out at 200 nm. At higher initial TCA concentrations, the chlorinated acetic acids were quantified by a modified HPLC method (Husain et al., 1992). Separation was performed on a ChromSphere C-18 (10 µm, 250 × 4.6 mm ID) reversed-phase column. Analytical conditions were

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 105 isocratic and the eluent consisted of 0.15 M (NH4)2SO4(aq), pH 5.5. The injection volume was 20 µL and the flow-rate 1.0 mL/min. HPLC analyses were performed using a Series 10 Liquid Chromatographic Pump (Perkin-Elmer) and a SPD-10 A VP UV-VIS detector (Shimadzu). UV-VIS detection was carried out at 210 nm.

Chloride was determined in the GR-SO4 kinetic experiments using a flow injection system with spectrophotometric detection (Cheregi & Danet, 1997).

5.3 Results and discussion

5.3.1 Product formation and reaction kinetics Trichloroacetate was readily reduced to DCA by all the Fe(II)-bearing minerals examined. Only DCA was detected within the reaction time in all the Fe(II)-Fe(III) mineral systems. Experiments conducted with the various Fe(II)-Fe(III) mineral systems and DCA confirmed that no significant reduction of DCA took place (data not shown). Hence, it is reasonable to assume that the further hydrogenolysis of DCA to MCA is too slow to be detected within the experimental time frame here. The mass balance of TCA and DCA was almost complete in all suspensions ruling out any alternative reaction pathways to reductive dechlorination. Decarboxylation of TCA producing chloroform and requires high temperatures and is therefore assumed not to take place at the experimental conditions applied here (Atkins et al., 1984). Based on these results, we propose that the reductive dechlorination of TCA by Fe(II)-bearing minerals proceeds via hydrogenolysis (replacement of halogen by hydrogen) as reported for the transformation of TCA by zero-valent iron (Hozalski et al., 2001). Thus, in order to reduce TCA to DCA, 2 electrons corresponding to 2 Fe(II) are needed (Figure 5.1).

106 Chapter 5

O O 2e-, H+ Cl- C C - Cl3C O Cl2HC O-

TCA DCA

Figure 5.1. Proposed reductive transformation pathway of TCA.

In the Fe(II)/goe and Fe(II)/lep systems, we detected no TCA transformation in the absence of either aqueous Fe(II) or pure or calcite-associated goethite and lepidocrocite. These results strongly indicate that reactive Fe(II) species associated with the goethite and lepidocrocite surfaces are the reductants for TCA in these systems. The Fe(III) phases forming in the mineral suspensions were not characterised and therefore the reaction stoichiometry cannot be assessed.

At initial Fe(II) concentrations in large excess of initial TCA concentration, we found a pseudo 1. order rate law for the degradation of TCA by Fe(II):

d[]TCA a b k [][Fe(II) ⋅⋅=− TCA] dt where a = 1, b = 1 and the observed pseudo 1. order rate constant kobs = k · [Fe(II)].

At all [Fe(II)]0/[TCA]0 ratios studied (6-738), TCA was transformed almost quantitatively into DCA and the reaction kinetics followed pseudo 1. order kinetics with respect to TCA (Figure 5.2). The observed pseudo 1. order rate constants for the transformation of TCA by the various Fe(II)-Fe(III) mineral systems were calculated as initial rates (i.e. max. first two half-lives) from linear fits of (time, ln

[TCA]t/[TCA]0)-plots (Table 5.1). The amount of chloride produced during reaction with GR-SO4 was always equivalent to the amount of TCA transformed into DCA (Figure 5.2c). This also indicates that no significant further reduction of

DCA took place in GR-SO4 suspensions.

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 107

Figure 5.2. Time course of TCA consumption and DCA and chloride production for a)

Fe(II)/Goe ([Fe(II)tot]0 = 0.95 mM), b) Fe(II)/Lep ([Fe(II)tot]0 = 0.91 mM), c) GR-SO4

([Fe(II)GR]0 = 9.62 mM), d) GR-CO3 ([Fe(II)tot]0 = 6.33 mM) and e) Magnetite ([Fe(II)tot]0 = 3.50 mM). Solid lines represent 1. order kinetic fits whereas symbols and dotted lines represent actual data. = TCA, = DCA, = Cl–.

Table 5.1. Experimental conditions and pseudo 1. order rate constants for the reductive transformation of TCA by various Fe(II)-Fe(III) containing mineral systems.

Suspension age [Fe(II)] a [Fe(II)] b Surface area k f System solid aq [TCA] (µM) pH c pH d k e (min-1) obs (d) (mM) (mM) 0 init end obs (m2/L) (min-1·m-2·L)

g Fe(II)aq 1 0 0.30 43.4 n.d. 7.6 n.d. n.d. n.d.

Fe(II)/Goe 1 0.02 0.24 42.9 n.d. 7.8 g 1.02·10-4 7.1 i 1.43·10-5

Fe(II)/Goe 1 0.13 0.94 54.3 7.7 7.0 2.25·10-4 7.1 i 3.16·10-5

Fe(II)/Goe 1 0.234 0.066 48.4 7.65 8.0 g 6.40·10-4 54.0 j 1.19·10-5 coating Fe(II)/Goe 1 0.15 0.80 48.6 n.d. 7.6 g 12.43·10-4 54.0 j 2.30·10-5 coating

Fe(II)/Lep 1 0.02 0.23 15.7 n.d. 7.8 g 0.75·10-4 8.0 i 0.94·10-5

Fe(II)/Lep 1 0.163 0.137 47.0 7.65 8.0 g 2.82·10-4 54.0 j 0.52·10-5 coating Fe(II)/Lep 1 0.10 0.81 41.7 n.d. 7.7 g 8.31·10-4 54.0 j 1.54·10-5 coating

g -4 k -4 Fe3O4 1 3.38 0.12 51.3 8.10 7.8 8.30·10 1.6 5.31·10

-4 k -5 Fe3O4 77 11.2 5.6 55.6 7.0 7.0 1.53·10 5.2 2.95·10

g -4 l -5 GR-CO3 1 5.94 0.39 50.3 7.65 8.4 7.61·10 41.9 1.82·10

-4 l -5 GR-CO3 2 7.6 0.1 47.8 8.56 8.29 40.8·10 53.6 7.61·10

-4 l -5 GR-CO3 32 7.3 0.03 56.3 8.5 8.0 4.90·10 51.5 0.95·10

g -4 l -5 GR-CO3 142 3.53 0.005 62.9 n.d. 8.7 5.13·10 24.9 2.06·10

-4 m -5 GR-SO4 1 5.17-12 17 0.86-1.39 105 n.d. n.d. 3.60·10 92.6 0.39·10

-4 m -5 GR-SO4 1 6.22-10 27 0.93-1.45 270 n.d. n.d. 3.76·10 88.1 0.43·10

-4 m -5 GR-SO4 1 7.05-10 14 0.77-1.79 500 n.d. n.d. 3.74·10 91.8 0.41·10

-4 m -5 GR-SO4 1 5.17-10 51 0.60-1.65 1000 n.d. n.d. 2.89·10 83.7 0.35·10

n.d. = not detected. a. Initial structural or adsorbed Fe(II) estimated as [Fe(II)total] – [Fe(II)aq]. b. Initial dissolved Fe(II) measured. c. Suspension pH prior to TCA addition. d. Suspension pH at reaction termination. e. Pseudo 1. order rate constants for the consumption of TCA calculated from initial rates (max. first two half-lives). f. Surface area-normalised pseudo 1. order rate constants. g. pH control through pure CaCO3 and 0.5% CO2(g). h. pH control through Fe(III) oxide-coated calcite and 0.5% CO2(g). i. Estimated using the SSA of the Fe(III) oxide applied. j. Estimated using the SSA of calcite 2 2 -1 2 -1 2 ~1 m /g. k. Estimated assuming SSA = 4 m /g (Schwertmann & Cornell, 1991):½·[Fe(II)solid]0·232 g·mol ·4 m ·g . l. Estimated assuming SSA = 47 m /g (Williams & Scherer, 2001): -1 2 -1 2 -1 ¼·[Fe(II)GR]0·600 g·mol ·47 m ·g . m. Estimated as in l but using SSA = 71.2 m ·g (Chapter 4, this work).

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 109 5.3.2 Comparing rate constants obtained for the various Fe(II)-Fe(III) mineral systems Data for the systems containing iron oxide coated calcite were very similar to the data obtained for the pure iron oxides (not shown in Figure 5.3). Since no SSA was determined for magnetite in this study, a SSA of 4 m2/g was assumed (Schwertmann & Cornell, 1991). However, it should be noted that the magnetite synthesized by Schwertmann and Cornell (1991) was prepared differently (i.e. oxidation of Fe(II) by nitrate in a heated alkaline solution) from the magnetite applied in this study. The surface area-normalised pseudo 1. order kobs values obtained for GR-CO3, GR-SO4, Fe(II)/goethite and Fe(II)/lepidocrocite were all within the same order of magnitude (Figure 5.3a).

Figure 5.3. Average surface area-normalised pseudo 1. order rate constants for the degradation of a) TCA (this work), b) hexachloroethane (Elsner et al., 2004a) and c) carbon tetrachloride (Amonette et al., 2000; Pecher et al., 2002; O’Loughlin et al., 2003c; Elsner et al., 2004b) by

GR-SO4, GR-CO3 (suspension age 1 d), Fe3O4, Fe(II)/α-FeOOH and Fe(II)/γ-FeOOH.

Experimental conditions applied in this work: [Fe(II)tot]0 = 0.25-1.07 mM in the goethite and lepidocrocite suspensions, [Fe(II)tot]0 = 0.25-11.6 mM in the GR-SO4 and GR-CO3 suspensions, pH 7.0-8.6, 7.1-92.6 m2 mineral surface area/L. Experimental conditions applied by Elsner et al.: 1 mM aqueous Fe(II), 25 m2 mineral surface area/L. Experimental conditions applied in 2 references employed in c): [Fe(II)tot]0 = 1-8.3 mM, 25-275 m mineral surface area/L. GR-SO4 =

110 Chapter 5 sulfate green rust; GR-CO3 = carbonate green rust; Fe3O4 = magnetite; α-FeOOH = goethite; γ- FeOOH = lepidocrocite.

When comparing the rate constants for the Fe(II)-Fe(III) mineral systems found for reduction of TCA in this study (Figure 2a), mixed valent Fe(II)-Fe(III) minerals such as green rusts and magnetite containing structural Fe(II) within the mineral lattice do not seem to be significantly more reactive than Fe(II)-Fe(III) mineral systems containing Fe(II) associated with pure Fe(III) oxides. Unlike most other iron oxides, GRs contain not only external Fe(II) reactive sites at the surface but also internal sites in the space between consecutive Fe(II)-Fe(III) hydroxide layers. The GR interlayer thickness is a function of both the size and the charge of the interlayer anion. For solutes, the Fe(II) within the GR hydroxide layer is accessible at the outside basal planes and at the edges as well as through the interlayer under certain conditions (see Figure 4.2, Chapter 4, this work). Due to electrostatic forces, oxidants holding different charge properties (anions, cations, neutral molecules) may exhibit different affinities for the various reactive Fe(II) sites present in GR. As the reactive sites are located in/at the Fe(II)-Fe(III) hydroxide layers, the rate of reaction depends on the hydroxide layer area which can be accessed by the oxidant. If the oxidant is able to exchange with the interlayer anion, reaction can take place both at outer and inner surfaces of the GR particles and, in total, more reactive sites are available for the reaction. Thus, oxidant size and charge primarily control its access to the internal sites in GRs. XRD characterization of the GR-SO4 crystals during reaction with TCA demonstrated that the GR-SO4 interlayer spacing did not vary during reaction (Table 5.2). This may indicate that TCA did not enter the GR-SO4 interlayers.

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 111

Table 5.2. Diffraction angle, d-spacing and width at half peak height (W½) for the 001 GR-SO4 diffraction peak as a function of time during reaction with TCA ([Fe(II)GR]0 = 4 mM, [TCA]0 = 1 mM).

Angle d -spacing W½ Time (min) 001 (°2θ) (nm) (°θ)

0 9.483 1.0821 0.273 10 9.494 1.0809 0.287 215 9.522 1.0777 0.263 330 9.550 1.0745 0.273 510 9.524 1.0775 0.277 855 9.509 1.0791 0.253 1160 9.467 1.0839 0.268

We roughly estimated the molecular size of TCA by summing the covalent radii of the individual atoms (see Supporting Information 7.4). When comparing the molecular size of TCA with the GR-SO4 interlayer spacing of 0.61 nm, it can be concluded that only when the C-C bond is oriented perpendicular to the interlayer plane does the size of TCA exceed the GR-SO4 interlayer spacing. In contrast, the size of TCA exceeds the GR-CO3 interlayer spacing (0.26 nm) regardless of its orientation. Hence, if TCA was intercalated in the GR-CO3 interlayer we would expect the interlayer spacing to expand. The same holds for intercalation of a vertically oriented TCA in the GR-SO4 interlayer. Supposedly, both the low charge and the size of TCA impeded its access to the GR-SO4 and GR-CO3 interlayers, i.e. the divalent sulphate and carbonate in the GR interlayers did not readily exchange with the monovalent TCA since GR interlayers generally have a higher affinity for divalent anions than for monovalent anions (Miyata, 1983). Thus, TCA did neither access nor react with internal Fe(II) reactive sites in GR-SO4 which means that the reaction between TCA and GR-SO4 took place at the external reactive Fe(II) sites solely. It is reasonable to assume that the same holds for the reaction between TCA and GR-CO3. No significant aging effects, e.g. rate constants varying as a function of GR age, were observed within 142 days (see Table 5.1). However, the SSAs of

112 Chapter 5 the GR suspensions holding ages up to 142 days were not measured but estimated assuming that the GR SSA did not decrease within the time frame.

According to the Scherrer formula, the width at half peak height (W½) of a diffraction peak is inversely proportional to the average crystal dimension perpendicular to the given crystal plane (Klug & Alexander, 1974). The average

GR-SO4 particle thickness perpendicular to the basal plane (W½; Table 5.2) was constant during reaction, implying that TCA reacts only at the edges and not at the basal planes. Assuming that the platy hexagonal GR-SO4 and GR-CO3 crystals hold an average width of 1 µm and an average particle thickness of 35 nm (Hansen and Koch, 1998), the ratio of edge surface area to outer surface area is Aedge/Aouter ~

1/30 for GR-SO4 and 1/21 for GR-CO3 (see Supporting Information 7.3). This means that only 3% of the outer surface area in GR-SO4 and 5% of the outer surface area in GR-CO3 are available at the edges. Assuming that TCA reacts at the edges only and if employing the Aedge/Aouter in the estimation of the rate constants, the surface area-normalised pseudo 1. order rate constants for GR-SO4 and GR-

CO3 would be 20-30 times higher than the rate constants depicted in Figure 5.3a. Thus, the reactivity of GRs normalised to their reactive surface area is much higher than the reactivity normalised to their total surface area.

5.3.3 Comparing with rate constants obtained for other chlorinated aliphatic compounds Though care must be taken when comparing kinetic parameters obtained at different experimental conditions (e.g. pH, [Fe(II)]0/[TCA]0 ratios, surface area to volume ratios etc.), it is interesting to compare our results to those reported for hexachloroethane (Figure 5.3b; data from Elsner et al., 2004a). The reductive transformation of hexachloroethane was investigated for various Fe(II)-bearing minerals including Fe(II)/goethite, Fe(II)/lepidocrocite and GR-SO4 in the presence of 1 mM dissolved Fe(II) and 25 m2 mineral surface area/L at pH 7.2 except for the

GR-SO4 suspensions in which the dissolved Fe(II) concentrations were slightly

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 113 higher and pH = 8. The pseudo 1. order rate constants reported for hexachloroethane are in the range 1.8·10-4 – 7.5·10-3 h-1·m-2·L (Elsner et al., 2004a). When comparing Figure 5.3a with Figure 5.3b, it can be seen that the differences in intrinsic reactivity of the Fe(II)-bearing mineral systems are more pronounced for hexachloroethane than for TCA.

Caution should also be advised to the different reaction mechanisms by which hexachloroethane and TCA react. The transfer of a single electron and the formation of an alkyl radical upon removal of a chlorine atom constitute the first and, in most cases, the rate-limiting step in the reduction of chlorinated aliphatic compounds (Vogel et al., 1987). Depending on the chemical structure of the chlorinated aliphatic compound, the resulting free alkyl radical may undergo hydrogenolysis, chloroelimination or dimerization/coupling. In the case of TCA, the free dichloroacetate radical most likely undergoes hydrogenolysis. The almost quantitative transformation of TCA to DCA confirms that hydrogenolysis is the prevalent reaction mechanism in our mineral systems. The pentachloroethyl radical formed from hexachloroethane may undergo hydrogenolysis (producing pentachloroethane) or dichloroelimination (producing tetrachloroethene). Elsner et al. (2004a) found that hexachloroethane was transformed quantitatively into tetrachloroethylene for all minerals which strongly indicates that dichloroelimination was the dominating reaction mechanism. Another polychlorinated aliphatic compound transformed mainly by hydrogenolysis under reducing conditions is carbon tetrachloride. Several studies have investigated the reductive dechlorination of carbon tetrachloride by various Fe(II)-bearing minerals, including Fe(II)/goethite and GR-SO4, and reported pseudo 1. order rate constants in the order 1.52·10-4 – 6.40·10-4 h-1·m-2·L for Fe(II)/goethite and 8.64·10-4 h-1·m-2·L for GR-SO4 (Amonette et al., 2000; Pecher et al., 2002; O'Loughlin et al., 2003c; Elsner et al., 2004b). When comparing Figure 5.3a with Figure 5.3c, it can be seen that the range of magnitude of the rate constants and the differences in intrinsic

114 Chapter 5 reactivity of the Fe(II)-bearing mineral systems are similar for carbon tetrachloride and TCA.

5.3.4 Factors controlling the reactivity of surface-bound Fe(II) The reactivity of an oxidant towards Fe(II) surface species cannot be predicted from the reduction potentials of the redox couple alone. In heterogeneous systems, processes such as mass transfer and adsorption/desorption may have a rate-limiting effect on the overall reaction rate. If the adsorption follows a saturation-type sorption isotherm (e.g. Langmuir), the sorbate (oxidant) concentration at the surface will vary non-linearly with the total amount of oxidant added. This dependence will have to be taken into account when establishing rate laws for the heterogeneous reactions and when testing the hypothesis that the reaction rates depend on the sorbed concentration of the oxidants. pH has a strong impact on the sorption and thereby on the availability of ionizable oxidants. At the pH values applied here, the chloroacetates are fully dissociated (pKa (TCA) = 0.66, pKa

(DCA) = 1.35, pKa (MCA) = 2.87). However, we found the sorption of TCA to be negligible in suspensions of pure calcite, goethite/calcite and lepidocrocite/calcite at pH 7.6-7.7. Moreover, the mass balance of TCA and DCA was almost complete in all suspensions and therefore loss of TCA or DCA due to adsorption at mineral surfaces or incorporation in the GR anion interlayers can be ruled out. Calcite has a much lower adsorption capacity than most iron oxides, hence, we anticipate that goethite and lepidocrocite control the adsorption of TCA and DCA in both the pure FeOOH and the FeOOH/calcite suspensions. This was supported by our experimental results demonstrating that the presence of a calcite surface - either pure or as a support for goethite and lepidocrocite coatings - did not exert any noticeable effect on the reaction rates (see Table 5.1). In addition, the surface area- normalised rate constants for mineral systems containing goethite or lepidocrocite in pure form and mineral systems containing goethite or lepidocrocite as coatings on calcite were very similar. In heterogeneous reactions, mass transfer in bulk solution becomes the rate-limiting step when the surface reaction is much faster

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 115 than the diffusion of the reacting species to the reactive surface. However, at the low rate constants obtained here, the reaction of TCA with the Fe(II)-bearing minerals is not likely to be mass transfer limited (see Supporting Information 7.2).

One very important factor affecting heterogeneous redox reactions is pH which influences the speciation of the complexes in solution and at mineral surfaces as well as the stability of the more soluble Fe(II)-containing minerals such as GRs. In contrast to aqueous Fe(II) complexes, it is not possible to predict the reactivity of Fe(II) surface species as their reduction potentials are unknown. In the absence of specifically adsorbing solutes other than H+, the surface charge of the Fe(III) oxides goethite and lepidocrocite is determined by the surface densities of the + – charged surface species ≡FeOH2 and ≡FeO whereas the surface charge of calcite – + – is determined by the density of the surface species ≡CO3 , ≡CaOH2 and ≡CaO

(Stumm, 1992; Van Cappellen et al., 1993). The point of zero charge (pHpzc) of pure calcite is in the pH range 7-11 and depends on the partial pressure of carbon dioxide, pCO2. The higher the pCO2, the lower the pHpzc. At the experimental conditions applied here (0.5% = 0.005 atm CO2(g)), the pHpzc = 8.2 for calcite (Table 5.3). As only 10-11 mg Fe(III) of goethite and lepidocrocite was coated onto calcite, we assumed a pHpzc of 8.2 for the goethite and lepidocrocite coated calcite particles as well. The pHpzc values for green rusts are unknown.

116 Chapter 5 Table 5.3. Specific surface areas and point of zero charge of the various iron minerals in pure form as well as goethite and lepidocrocite coated onto calcite.

2 Mineral Structural formula SSA (m /g) pHpzc

II III a GR-SO4 Fe 4Fe 2(OH)12SO4·3H2O 71 -

II III b GR-CO3 Fe 4Fe 2(OH)12CO3·3H2O 47 -

e Magnetite Fe3O4 - 6.9 Goethite α-FeOOH 16 c 8.5 f Lepidocrocite γ-FeOOH 18 c 7.3 e

d g Calcite CaCO3 ≤ 1 8.2 Goe coating - ≤ 1 d 8.2 h Lep coating - ≤ 1 d 8.2 h

a. Chapter 4, this work. b. Williams & Scherer, 2001. c. Product information by Bayer. d. The SSA of calcite was quantified to ≤ 1 m2/g. The detection limit of our BET method was 1 m2/g. e. Charlet et al., 1998a. f. Liger et al., 1999. g. Van Cappellen et al., 1993. h. Same as for calcite.

The surface hydroxyl groups on iron oxides may be both singly (≡Fe-OH), doubly

(≡Fe2-OH), triply (≡Fe3-OH) and geminally (≡Fe-(OH)2) coordinated (Cornell & Schwertmann, 1996; Stumm, 1992). The differently coordinated surface hydroxyl groups are not equally reactive. Adsorption reactions involve only singly coordinated surface groups and therefore only this kind of hydroxyl groups on iron oxides will be considered here (Cornell & Schwertmann, 1996). Hence, the predominant surface sites available for adsorption in pure suspensions of Fe(III) 0 + – oxides are ≡FeOH , ≡FeOH2 and ≡FeO . In the presence of dissolved Fe(II), ≡FeIIIOFeIIOH0, ≡FeIIIOFeIIO– and ≡FeIIIOFeII+ constitute the main reactive sites at the Fe(III) oxide surfaces (Liger et a., 1999). Assuming that Fe2+ and other cationic Fe(II) species are the dominating adsorbates on the mineral surfaces in our experiments, we expect the actual pHpzc to be higher than the pHpzc of the pure oxides listed in Table 5.3. Hence, at pH<8.2 where most of our experiments were conducted, all the mineral surfaces presumably carry net positive charges.

At pH 7.0 where Fe2+ is still the predominant Fe(II) species in solution (~50%), we expect that ≡FeIIIOFeIIOH, ≡FeIIIOFeIIO– and ≡FeIIIOFeII+ constitute the main

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 117 reactive sites at the Fe(III) oxide surfaces as suggested by Liger et a., 1999. As pH increases from 7.0 to 8.7, the Fe(II) carbonate complexes become increasingly 2+ + 0 + important in solution at the expense of the Fe , FeCl , FeSO4 and FeOH species (King, 1998). Fe(II) carbonate complexes do not bind at the oxide surface as readily as the aquo or hydroxo complexes of Fe(II) but carbonate itself sorbs readily to Fe(III) oxide surfaces through which the Fe(III) oxide surface is coated by inner-sphere monodentate ≡FeIIIOCOOH0 surface complexes (Villalobos & Leckie, 2000 & 2001). The presence of carbonate shifted the sorption edge for the Fe(II) adsorption on goethite from pH 5.8 to 7.8 and the authors hypothesized this to be a result of the formation of aqueous and surface Fe(II)-carbonate complexes and to competition between carbonate and Fe(II) for Fe(III) oxide surface sites (Vikesland & Valentine, 2002). Similarly, monodentate surface complexes like III 0 III – ≡Fe Cl and ≡Fe OSO3 as well as ternary monodentate surface complexes like III II 0 III II – ≡Fe OFe Cl and ≡Fe OFe OSO3 and ternary bidentate surface complexes such III II as (≡Fe O)2Fe OSO3 may form at Fe(III) oxide surfaces when Fe(II), chloride and sulfate are present in solution (Ostergren et al., 2000; Kim et al., 2004). However, the effects of anionic ligands such as chloride and sulfate on Fe(II) adsorption at Fe(III) oxide surfaces and the reactivity of Fe(II) carbonate, chloride and sulfate surface sites are still unknown and need to be evaluated (see Supporting Information 7.5). Thus, we do not know whether chloride and sulfate decrease or increase the Fe(II) sorption in our mineral systems. We can only report that we did not detect any significant differences in the rate of TCA transformation between the mineral suspensions containing carbonate, chloride and sulfate, respectively. Hence, we anticipate that ≡FeIIIOFeIIOH, ≡FeIIIOFeIIO– and ≡FeIIIOFeII+ constitute the main reactive sites at the Fe(III) oxide surfaces within the whole pH range 7.0- 8.7. This might also explain why we did not detect any obvious systematic pH effect in the Fe(II)-Fe(III)-systems (see Table 5.1). In the case of hexachloroethane, the reactivity order GR-SO4>goethite>magnetite>lepidocrocite may be rationalized by the variations in surface site densities and total amount of

118 Chapter 5 Fe(II) sorbed on the iron minerals (see Supporting Information 7.5) as well as the different speciations and reactivities of the Fe(II) surface sites on the iron minerals.

5.3.5 Comparison with biotic and other abiotic systems Only one report on abiotic transformation of TCA is found in the literature and the study demonstrates the reductive dechlorination of TCA to MCA by Fe(0) (Hozalski et al., 2001). The authors reported a pseudo 1. order rate constant of 6.0·10-4 min-1·m-2·L for the transformation of TCA to DCA and a pseudo 1. order rate constant of 2.25·10-4 min-1·m-2·L for the transformation of DCA to MCA at

[Fe(0)]0 = 0.25 M, [TCA]0 = 100-200 µM and pH 3.6-6.2. The rate constant for TCA reduction by Fe(0) is 10-300 times faster than the rate constants reported for the Fe(II)-bearing mineral systems here.

There is abundant evidence that soil microorganisms and fungi can dechlorinate TCA but only little is known about the bacteria and processes involved in the biodegradation of TCA. Biotransformation of TCA has been found at both oxic and anoxic conditions. Most of the microorganisms isolated grow feebly on TCA as a sole source of carbon (Foy, 1975; Weightman et al., 1992; De Wever et al., 2000). Only one bacterium capable of growing on TCA as the sole carbon and energy source has been characterized (Yu & Welander, 1995). In addition, anaerobic bacteria coupling co-metabolic growth to reductive dechlorination of TCA have been isolated (Weightman et al., 1992; De Wever et al., 2000). The inability to grow on the less chlorinated acids DCA and MCA is a notable feature of both the aerobic and anaerobic bacteria. Complete transformation of TCA to methane and carbon dioxide has only been found when abiotic and biotic processes were combined (Egli et al., 1989). The abiotic transformation of TCA to DCA occurred spontaneously in the presence of sterile activated charcoal whereas the DCA formed was further degraded to methane and carbon dioxide by a mixed culture of methanogenic bacteria. However, the abiotic reductant(s) responsible for the transformation of TCA to DCA was not reported (Egli et al., 1989).

Reductive Transformation of Trichloroacetate in Abiotic Fe(II)-Fe(III) Mineral Systems 119

The rate constants obtained in this work suggest that the Fe(II)-bearing mineral systems may be important reductants of TCA in natural suboxic environments. In natural iron-rich soils holding specific surface areas of 22 m2/g (Kenneke & Weber, 2003), average bulk densities of 2.65 g/cm3 and porosities of 25% and containing 2% iron oxides, a rough estimation of the half-life of TCA amounts to 47 minutes when applying the average surface area-normalised rate constant obtained for all the Fe(II)-Fe(III) mineral systems in this work (1·10-3 h-1·m-2·L). This estimation is based on the assumption that enough reactive Fe(II) is available in these soils. The natural iron-reducing sediment investigated by Kenneke and Weber (2003) contained 80 µM Fe(II) in the soil solution and 315 µmole Fe(II) per g sediment At such low Fe(II) concentrations, the overall rate of abiotic transformation of TCA in natural soils and sediments is most likely limited by the regeneration of reactive Fe(II). Hence, the continuous regeneration of reactive Fe(II) surface sites by adsorption of abiotically or microbially produced Fe(II) may further the long-term abiotic transformation of TCA in such environments.

5.4 Conclusions This work demonstrates that various Fe(II)-Fe(III) minerals systems including GR-

SO4, GR-CO3, magnetite, Fe(II)/goethite and Fe(II)/lepidocrocite readily transform TCA to DCA. Dichloroacetate was not further reduced to MCA or acetate by any of the Fe(II)-bearing minerals. The surface area-normalised pseudo 1. order rate constants obtained for the reductive transformation of TCA by the various Fe(II)- bearing minerals did not differ significantly from each other. The results obtained in this work infer that under natural geochemical conditions, Fe(II)-bearing mineral systems may play an important role in the overall transformation of TCA. Thus, not only microbial degradation but also abiotic reductive transformation of TCA by Fe(II)-bearing minerals may govern the fate of TCA in natural subsurface environments. This holds both for natural systems like iron-rich suboxic soils and sediments as well as engineered systems like permeable reactive barriers of zero-

120 Chapter 5 valent iron implemented for on-site remediation where both Fe(0) and solid or surface-bound Fe(II) corrosion intermediates may transform TCA.

Acknowledgments We would like to thank Susanne Guldberg for performing the experimental work comprising

GR-SO4.

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Conclusions and Outlook 125

6 Conclusions and Outlook

The work presented in this dissertation adds to the understanding of how Fe(II)- bearing minerals like green rusts (GRs), vivianite (Fe2(PO4)2⋅8H2O), magnetite

(Fe3O4) and Fe(II) associated with goethite and lepidocrocite may form and react in nature. In order to elucidate the role of bacteria in the formation of GRs in natural soils and sediments, we studied the iron mineral phases forming as a result of the activity of iron-respiring bacteria. In chapter 2, the Fe-containing products formed by anaerobic, autotrophic denitrifying Fe(II)-oxidizing bacteria (FeOB) were examined. The culture medium applied contained high levels of bicarbonate and phosphate and is typically used in this kind of studies as it provides excellent conditions for the nitrate-reducing FeOB. Fe(II) was present initially as a whitish solid Fe(II) hydroxy phosphate (vivianite) and as soluble Fe(II). The results obtained demonstrate that the denitrifying FeOB produce poorly crystalline goethite via a greenish Fe(III)-enriched vivianite intermediate. Mössbauer spectroscopic analyses provided no evidence of green rust formation. At low phosphate concentrations where vivianite does not control the Fe(II) activity, it is reasonable to assume that siderite (FeCO3) precipitates initially and that carbonate GR phases may form during biooxidation. At low bicarbonate concentrations, we would expect Fe(II) sulfate or chloride species to dominate initially (depending on the Fe(II) source applied) and sulfate GR or chloride GR to form during biooxidation. In chapter 3, we investigated the Fe-containing products formed during reduction of common Fe(III) oxides by the anaerobic dissimilatory Fe(III)- reducing microorganism, Shewanella algae BrY. S. algae BrY reduced substantial amounts of the initial Fe(III) and green and blackish mineral phases were produced within 1-2 weeks after inoculation. Mössbauer spectroscopic analyses showed that the green and black precipitates consisted of green rust and vivianite.

We studied the reactivity of synthetic GRs towards reducible organic pollutants in order to asses the potential significance of GR phases for the fate of such

126 Chapter 6 compounds. To this end, we used nitroaromatic compounds (NACs) and chlorinated acetates as suitable model compounds for studying redox reactions potentially relevant in the environment. In chapter 4, we investigated the relative reactivity of outer and inner Fe(II) reactive sites in synthetic sulfate green rust

(GR-SO4) by using a series of structurally closely related compounds with different charge properties as “reactive probes”. The probe compounds included nitrobenzene, 2-nitrophenol, 4-nitrotoluene, 4-chloronitrobenzene and 4- nitrophenylacetic acid. Our results demonstrated that NACs are completely reduced to their corresponding anilines by GR-SO4. The reactions followed pseudo 1. order kinetics with respect to NAC and the surface area-normalised pseudo 1. -4 -1 -2 order rate constants obtained were 0.16–4.65·10 s ·m ·L at [Fe(II)GR]0 = 1.03-

12.60 mM, [NAC]0 = 20-102 µM and pH 8.4-8.6. Neither mass transfer control nor surface saturation kinetics could account for the similarity of the surface- normalised pseudo 1. order rate constants obtained for the reduction of the neutral and anionic NACs by GR-SO4. These observations suggest that the reaction between NAC and GR-SO4 takes place at the external reactive Fe(II) sites. At low initial Fe(II)GR concentrations, the external reactive Fe(II) sites were depleted and the regeneration of new external reactive sites eventually controlled the reduction of the NACs by GR-SO4. In chapter 5, we examined the reactivity of various Fe(II)-Fe(III) mineral systems towards trichloroacetic acid (TCA) and dichloroacetate (DCA) in laboratory batch experiments imitating natural conditions. The Fe(II)-Fe(III)-systems investigated included GR-SO4, carbonate green rust, magnetite, Fe(II)/goethite and Fe(II)/lepidocrocite. TCA was readily reduced to DCA by all Fe(II)-containing minerals. The reactions followed pseudo 1. order kinetics with respect to TCA and the surface area-normalised pseudo 1. -5 -1 -2 order rate constants obtained were 0.33–7.6·10 min ·m ·L at [Fe(II)]0 = 0.25–

11.6 mM, [TCA]0 = 15–1000 µM and pH 7.0–8.7. Our results showed no significant differences regarding product distribution and surface area-normalised reaction rate constants between the Fe(II)-Fe(III)-systems. DCA was not further

Conclusions and Outlook 127 reduced to monochloroacetate (MCA) or acetate in any of the systems within the time frame in our experiments.

As suggested in chapters 2 and 3, sufficient evidence must be provided and caution should be exercised when proclaiming new biogenic minerals. The study of microbially produced GRs is still in its infancy and more research is needed in order to elucidate the role of bacteria in the formation of GRs in natural soils and sediments. The results presented in chapter 2 indicate that microbiological processes may be responsible for the oxidation of vivianite and metavivianite II III ((Fe 3-x,Fe x)(PO4)2(OH)x·(8-x)H2O, x > 1.2) in natural subsurface environments. In chapter 3, we demonstrated that GRs may be produced microbially at conditions including low carbon and Fe(III) concentrations as well as the exclusion of synthetic electron shuttles and pH buffers. The role of microbial processes in the redox cycling of iron in the subsurface and the ways in which these processes can be coupled to contaminant remediation are currently active areas of research. Zero- valent iron has been the most extensively studied reductant for the treatment of many inorganic and organic contaminants and is currently the most commonly used material for the construction of permeable reactive barriers (PRB) but a detailed understanding of the processes involved in the reduction of these pollutants by Fe(0) is lacking (Scherer et al., 2000). Potentially reactive Fe(II)- bearing corrosion products identified in iron metal columns and barriers include magnetite, siderite, Fe(II) sulfides, green rusts as well as Fe(II) sorbed to mineral surfaces (Gu et al., 1999; Roh et al., 2000). The formation of reactive Fe(II)- bearing minerals like GRs may explain the effective long-term operation of zero- valent iron PRBs despite the formation of thick oxide films. Thus, natural in situ PRBs might be created by stimulating the activity of anaerobic dissimilatory Fe(III)-reducing bacteria and the subsequent formation of Fe(II) species such as GRs. Furthermore, suspensions of synthetic GRs, which are easily prepared from relatively inexpensive commodity chemicals, may also be injected and dispersed into the subsurface.

128 Chapter 6

The reductive transformation of NACs and TCA by GRs is relevant to understanding the processes responsible for their degradation in the subsurface and the development of innovative technologies for their remediation. The results obtained in chapters 4 and 5 indicate that GRs may play a significant role in the reductive transformation of NACs and TCA in natural subsurface environments. Furthermore, our results suggest that mainly the outer Fe(II) sites in GRs are utilized in the reaction with neutral and monovalent anionic compounds and that these sites may be replenished, e.g. by reduction of the oxidized surface sites or adsorption of Fe(II) from solution. The continuous restoration of Fe(II) surface sites in GRs may promote their long-term reactivity towards reducible contaminants.

References

Gu, B.; Phelps, T.J.; Liang, L.; Dickey, M.J.; Roh, Y.; Kinsall, B.L.; Palumbo, A.V.; Jacobs, G.K. (1999) Biochemical dynamics in zero-valent iron columns: Implications for permeable reactive barriers. Environmental Science and Technology, 33, 2170-2177.

Roh, Y.; Lee, S.Y.; Elless, M.P. (2000) Characterization of corrosion products in the permeable reactive barriers. Environmental Geology, 40, 184-194.

Scherer, M.M.; Richter, S.; Valentine, R.L.; Alvarez, P.J.J. (2000) Chemistry and microbiology of permeable reactive barriers for In Situ groundwater clean up. Critical Reviews in Environmental Science and Technology, 30, 363-411.

Supporting Information I

7 Supporting Information

7.1 Estimation of the one-electron reduction potential for 4-NPA

1 The one-electron reduction potential Eh ' of the half-reaction for a given NAC

- ArNO2 + e ArNO2 can be used for comparing reduction rates of different NACs in a given system. The formation of the nitroaryl radical is the rate-determining step in the overall rate of the reduction of a NAC to the corresponding aniline. The difference between the

1 Eh ' of a NAC and a given reductant is proportional to the change in standard free energy for the transfer of the first electron ∆G1°’. If a linear relationship between

1 the free energy of activation and ∆G1°’ is assumed, the Eh ' values of various NACs can be a measure of their relative reactivity with a given reductant.

As neither the one-electron reduction potential for 4-nitrophenylacetic acid (4- NPA) nor the Hammett constant for the acetic acid substituent could be found in the literature, the one-electron reduction potential for 4-NPA was estimated by application of a linear free energy relationship (LFER) to experimental data. Kinetic experiments were conducted in order to obtain the pseudo 1. order rate constant, kHJUG− , for the reduction of 4-NPA by a model hydroquinone (reduced juglone (8-hydroxy-1,4-naphthoquinone) in the presence of HS–). The reduction of a NAC by juglone follows the rate law

[]ArNOd 2 − − []kArNOk − [HJUG ] []− − [][]JUGfkArNO ⋅⋅⋅=⋅⋅=⋅= ArNO dt obs 2 HJUG 2 HJUG HJUG tot 2

1 and the Eh ' was deducted from a LFER

II Chapter 7

1 Eh ' log − ak +⋅= b HJUG 05916.0 for which a and b values have been established for a range of NACs with known

1 1 Eh ' values (Hofstetter et al., 1999). An excellent correlation of Eh ' and log kHJUG− has been found to exist over a range of 250 mV corresponding to more than 5 order of magnitude for kHJUG− . This is due to the fact that the actual transfer of the first electron is the rate-determining step under the experimental conditions chosen.

For comparison, experiments with 4-nitrotoluene (4-NT) were also conducted. The pseudo 1. order rate constants for the reduction of 4-NPA with juglone were corrected for the reduction of 4-NPA with only HS– (control experiments containing no juglone):

∗ − − − HJUG kk HS k − = HJUG []HJUG −

-1 -1 where kHJUG− (M ·s ) is the rate constant for a compound in the presence of only

∗ -1 juglone, kHJUG− (s ) is the pseudo 1. order rate constant for a compound in the – -1 presence of both juglone and HS , kHS − (s ) is the pseudo 1. order rate constant for the control reaction in the presence of only HS– and [HJUG–] (M) is the concentration of the reactive dissociated HJUG- form (nondissociated hydroquinone species are very nonreactive as compared to the monophenolate species).

Supporting Information III

OH 0 OH OH

+ e- + H+ = + e- + H+ =

0 0

pl(,.(ox) = 8 00 PK.,1 (red) = 6 60

JUG HJUG

OH OH

OH

pKa,2(red) = 10 60

Figure 7 .1 Oxidized and reduced juglone fo1m s.

Kinetic experiments in homogeneous anoxic aqueous solutions contained 5 mM

HS-, 20 µM total juglone, 50 mM KH2P04 buffer and were conducted at pH = 6.60 corresponding to a concentration of the reactive dissociated juglone form [HJUG"] = 10 µM.

By using the LFER

E1' (Hofstetter et al., 1999) logkHJUG- = 1.25 · 0.05~ 16 + 9.23

the fo llowing values were obtained

-1 -1) NAC kHJUG- (M ·s log kHJUG- E~' (mV)

4-NT 3.11 ·10·7 8.47· 10-8 2.26·10·2 -1.65 -515

4-NPA 1.64·10·7 1.16· 10·7 4.89·10·3 -2.31 -546 IV Chapter 7

∗ kHJUG− -values are averages of triplicates whereas kHS − -values are averages of

1 duplicates. The Eh ' determined for 4-NT in this work (-515 mV) differs 3% from

1 the Eh ' -value of -500 mV reported in the literature (Meisel & Neta, 1975;

1 Wardman, 1989). Hence, it is assumed that the Eh ' -value determined for 4-NPA also differs by 3%.

1 Note that even for NACs holding very different Eh ' values, the difference in their reactivities are much less pronounced in Fe(II)-Fe(III) systems, such as the Fe(II)/goethite system (LFER slope a = 0.6; Hofstetter et al., 1999) and the Fe(II)/magnetite system (LFER slope a = 0.34; Klausen et al., 1995) as compared to the juglone/H2S system (a = 1.25). Furthermore, it should be noted that all LFERs mentioned here were established for neutral NACs and in this work we have simply assumed that the LFERs are also valid for anionic NACs.

7.2 The rate-limiting step The overall rate of a reaction is equal to the rate of the slowest step in the mechanism. In heterogeneous reactions, e.g. a compound reacting at the surface of suspended particles in bulk solution, the overall process by which the heterogeneous reactions proceed may be broken down into a sequence of individual diffusion steps and reaction steps: 1) Mass transfer (diffusion) of the reactant from the bulk fluid to the external surface of the solid phase. 2) Adsorption of reactant onto the solid surface. 3) Reaction on the solid surface. 4) Desorption of the products from the solid surface. 5) Mass transfer of the products from the external solid surface to the bulk fluid. Hence, the rate of reaction of a compound reacting at the surface of suspended particles in bulk solution may be either mass transfer, adsorption/desorption or surface reaction limited. When the diffusion steps are much faster than the reaction steps, the mass transfer or diffusion steps do not affect the overall reaction rate. However, if the reaction steps

Supporting Information V are very fast compared with the diffusion steps, mass transport affects the reaction rate. Here only the external mass transfer is considered, i.e. the diffusion of reactants or products between the bulk fluid and the external surface of the solid phase. The additional internal mass transfer resistance for particles containing substantial internal surface area is not addressed.

7.2.1 Mass transfer (diffusion) limited kinetics The overall rate constant can be represented by a system of resistances in series (Fogler, 1999; Arnold et al., 1999):

⎛ 1111 ⎞ ⎜ += ⎟ ⎜ ⎟ obs ⎝ kkak −geomSAL ⎠

-1 where kobs is the observed rate constant, kL is the mass transfer coefficient (m·s ), a is the ratio of the external (geometric) specific surface area to volume of solution -1 (m ) and kSA-geom is the intrinsic rate constant of the reaction normalized to the external specific surface area rather than the BET specific surface area. By comparing kL·a with kobs one can estimate the role of mass transfer on the rate of reaction. Thus, if kL·a >> kobs, mass transfer is so fast that it has no impact on the reaction rate whereas if kL·a ≤ kobs, mass transfer is the rate limiting step.

In fluid dynamics, the Reynolds number, Re, is used for determining whether a flow is laminar or turbulent: ⋅ ud Re = tp ν

where dp is the particle diameter (m), ut is the terminal particle settling velocity (m·s-1) and ν is the kinematic fluid viscosity (m2·s-1): ν = η / ρ where η is the (absolute) dynamic fluid viscosity in centipoise (1 centipoise = 1 mPa·s = 10-3 kg·m-1·s-1) and ρ is the fluid density (kg·m-3).

VI Chapter 7

At Re < 1, we can apply Stoke’s particle settling velocity. Stoke’s law is an equation relating the terminal settling velocity of a smooth, rigid sphere in a viscous fluid of known density and viscosity to the diameter of the sphere when subjected to a known force field:

dg 2 ( −⋅⋅ ρρ ) u = pp (m·s-1) t 18 ⋅η

-2 -3 where g = 9.81 m·s is the gravitational constant, ρp is the particle density (kg·m ).

The Sherwood number is the main parameter for prediction of the mass transfer process in fluid dynamics: ∗ ⋅ dk Sh∗ = pL ⋅⋅+= ScRe6.02 3/12/1 Diw

2 -1 ∗ where Diw is the diffusion coefficient of the compound i in water (m ·s ), kL is the minimum (uncorrected) value of the mass transfer coefficient and Sc is the Schmidt number. This relation is often referred to as the Frössling correlation. The particle diameter is a key parameter in the Frössling correlation and the external mass transfer coefficient varies with square of the particle size for smaller particles.

The Schmidt number is the ratio of the kinematic fluid viscosity and the diffusion coefficient of the compound i in water: ν Sc = Diw

Supporting Information VII

According to Harriott (1962), the actual mass transfer coefficient kL is 1.5 times

∗ greater than the minimum value of the mass transfer coefficient kL . The uncertainty in kL·a associated with particle sphericity and roughness issues are believed not to exceed a factor of 2.

The diffusion coefficient of a compound i in water can be estimated as (Hayduk & Laudie, 1974):

−9 ⋅1026.13 2 -1 Diw = (m ·s ) 141 5890 η ⋅V i

3 -1 where Vi is the molar volume of the compound i (cm ·mol ) estimated according to Fuller et al., 1966.

Assuming spherical particles, the external (geometric) specific surface area and the particle diameter are calculated from the measured BET specific surface area, Atot, assuming that our GR-SO4 has a Atot/Aouter ~ 30 similar to the one reported by Hansen & Koch (1998):

π ⋅ 2 Atot SAp d p 6 2 -1 SAgeom == = = (m ·g ) 30 V ⋅ 1000ρ 1 3 d ⋅ 1000ρ p ()p π d ⋅⋅ ()1000ρ p (p ) 6 p p

-3 In our aqueous GR-SO4 system, the density ρ = 1000 kg·m , the absolute dynamic viscosity η = 10-3 Pa·s and the kinematic viscosity ν = 10-6 m2·s-1 for water. The

GR-SO4 particle specific parameters used is found below.

VIII Chapter 7

GR-SO4

2 Atot (m /g) 71.2

2 SAgeom (m /g) 2.37

3 Particle density ρp (kg/m ) 1500

-6 Particle diameter dp (m) 1.69·10

-7 Settling velocity ut (m/s) 7.78·10 Reynolds number 1.32·10-6

The molar volumes, the diffusion coefficients in water and the Schmidt numbers for the NACs including 4-chloronitrobenzene (4-CNB) and nitrobenzene (NB) were:

3 -1 2 -1 Compound Vi (cm ·mol ) Diw (m ·s ) Sc 4-NT 126.0 7.68·10-10 1302 4-CNB 123.0 7.79·10-10 1284 4-NPA 153.5 6.84·10-10 1462 NB 105.5 8.53·10-10 1173

The ratio of the external (geometric) specific surface area to volume of solution were calculated for GR-SO4 at the various concentrations applied:

-1 [Fe(II)GR]0 (mM) a (m ) 1.03 3.66·102 6.3 2.24·103 12.6 4.48·103 6.0 2.13·103

The uncorrected mass transfer coefficients were estimated for the NACs using the Frössling correlation:

Supporting Information IX

∗ -1 [Fe(II)GR]0 (mM) kL (m·s ) 4-NT 9.12·10-4 4-CNB 9.25·10-4 4-NPA 8.12·10-4 NB 1.01·10-3

Finally, kL·a was calculated and compared with the experimental 1. order rate constants, kobs, obtained for the NACs:

-1 -1 a Compound [Fe(II)GR]0 (mM) kL·a (s ) kobs (s ) 4-NT 1.03 0.50 4.20·10-4 6.3 3.07 1.40·10-3 12.6 6.13 5.90·10-3 4-CNB 1.03 0.51 7.40·10-4 6.3 3.11 1.70·10-3 12.6 6.22 4.60·10-3 4-NPA 1.03 0.45 6.40·10-4 6.3 2.73 1.09·10-3 12.6 5.46 4.73·10-3 NB 6.0 3.24 1.37·10-3 b

a. Experimental pseudo 1. order rate constant at 50 µM [Ar-NO2]0. b. Experimental pseudo 1. order rate constant at 10 µM [Ar-NO2]0.

When comparing kL·a with kobs, it can be seen that the rates of mass transfer for all 3 NACs exceed the observed rate constants by at least 3 or 4 orders of magnitude at every initial Fe(II)GR concentration Thus, the reaction of the given NACs with

GR-SO4 is not subject to mass transfer limitations under the experimental conditions applied here.

7.2.2 Surface saturation limited kinetics More than 75% of all heterogeneous reactions that are not diffusion-limited are surface-reaction-limited rather than adsorption- or desorption-limited. We now

X Chapter 7 look at the reaction A = B = C, where an intermediate B is formed. In our system,

A = Ar-NO2, B = Ar-NHOH and C = Ar-NH2. In this case the surface reaction is assumed to be a single-site mechanism where only the site S on which A or B is adsorbed is involved in the reaction forming B or C:

KA Adsorption 1 A + S = A·S

kS1 Surface reaction 1 A·S = B·S

-1 KB

Desorption 1 B·S = B + S

KB Adsorption 2 B + S = B·S

kS2 Surface reaction 2 B·S = C·S

-1 KC

Desorption 2 C·S = C + S

The rate law for this surface-reaction limited single-site mechanism involving an intermediate follows Langmuir-Hinshelwood kinetics (adopted from Fogler, 1999):

dC ⋅ ⋅ ⋅ CKCk A =− S1 AAsites dt 1 ⋅+⋅+⋅+ CKCKCK CCBBAA

Supporting Information XI where kS1 is the intrinsic rate constant of the surface reaction transforming A into the intermediate B, Csites is the concentration of reactive sites S on the solid surface, KA, KB and KC are the adsorption constants for A, B and C at the reactive surface sites and CA, CB and CC are the concentrations of A, B and C in the bulk fluid. Two major assumptions of the Langmuir isotherm imply that there is a fixed number of localised surface sites present on the surface and that the activity of the surface towards adsorption, desorption or surface reaction is independent of surface coverage.

Hence, fitting -∆CA/∆t to CA, CB and CC using a nonlinear curve fitting software such as SigmaPlot may provide one with the intrinsic rate constant and the adsorptions constants. If KB and KC >> KA, the intermediate and the product are strongly competing with the reactant for vacant reactive surface sites.

Our data was not fitted successfully by the Langmuir-Hinshelwood rate law (regression results not shown). Simplifying the rate law by excluding either the term KC·CC or KB·CB or both (assuming that the aniline product or the hydroxylaniline intermediate or both did not compete for the reactive sites) did not improve the regression. The Langmuir-Hinshelwood rate law for a dual-site mechanism did not fit our data either. Thus, Langmuir-Hinshelwood kinetics cannot explain the reaction mechanism of the given NACs in our GR-SO4 system.

7.3 External surface area of GR-SO4 and GR-CO3

The GR-SO4 unit cell consists of one double layer (d001 = 1.1 nm), i.e. one hydroxide layer (0.49 nm ) and one interlayer (0.61 nm). Hexagonal GR-SO4 particles holding an average width of 1 µm (Figure 7.2), an average particle thickness of 35 nm (Hansen & Koch, 1998) and a hydroxide layer thickness of 0.49 nm have a surface area of the basal plane

2 Abasal = 1 µm · 1 µm – 2 · 0.5 µm · 0.25 µm = 0.75 µm

XII Chapter 7 and a surface area of the edges

2 Aedge = (2 · 0.5 µm + 4 · 0.56 µm) · 0.00049 µm = 0.0016 µm

Figure 7.2. The hexagonal platy morphology of GR particles holding an average width of 1 µm.

The particle thickness is the mean crystal thickness perpendicular to the 003 plane as determined from the 003 reflections in an X-ray diffractogram. A GR-SO4 particle holding a thickness of 35 nm contains 35 nm/1.1 nm = 31.8 double layers.

The GR-CO3 unit cell consists of one double layer (d001 = 0.75 nm), i.e. one hydroxide layer (0.49 nm ) and one interlayer (0.26 nm). Hence, a GR-CO3 particle holding a thickness of 35 nm contains 35 nm/0.75 nm = 46.7 double layers.

The outer surface area of a GR-SO4 particle, including outer basal planes and edges, is

2 2 2 Aouter = 0.752 31.8µm ⋅+⋅ 0.0016 = 1.55µm µm

and the total surface area of a GR-SO4 particle, including both inner and outer basal planes as well as edges, is

Supporting Information XIII

2 2 2 Atot = 0.752(31.8 µm +⋅⋅ 0.0016 = 47.8)µm µm

Hence, the ratio of outer surface area to total surface area is

2 Aouter 1.55 µm = 2 ≈ 1/31 A tot 47.8 µm

Furthermore, the ratio of edge surface area to outer surface area is

2 Aedge 31.8 ⋅ 0.0016 µm = 2 ≈ 1/30 Aouter 1.55 µm

For GR-CO3, the outer surface area including outer basal planes and edges, is

2 2 2 Aouter = 0.752 ⋅+⋅ 0.00167.46µm = 1.57µm µm

and the total surface area of a GR-CO3 particle, including both inner and outer basal planes as well as edges, is

2 2 2 Atot = 0.752(46.7 µm +⋅⋅ 0.0016 = 70.1)µm µm

Hence, the ratio of outer surface area to total surface area is

A 1.57 µm 2 outer = ≈ 2 1/45 A tot 70.1 µm

Furthermore, the ratio of edge surface area to outer surface area is

2 Aedge ⋅0.001646.7 µm = ≈ 2 1/21 Aouter 1.57 µm

XIV Chapter 7

7.4 Van der Waals radii The size of polyatomic molecules can be estimated by summing the van der Waals radii of the individual atoms. Van der Waals radii or nonbonded radii can be pictured as the radii of hard spherical atoms (Figure 7.3).

Figure 7.3. Schematic of neighboring nonbonded atoms with van der Waals radii rA and rB.

Assuming that the spheres of neighboring nonbonded atoms just touch (Figure

7.3), the highest possible ion or molecule size, Ms, can be estimated as the sum of the van der Waals radii:

Ms = 2·rA + 2·rB + (1)

Taking Pauling’s rule for nonmetals into account, we can estimate the real size of polyatomic ions bound by covalent bonds (Pauling, 1960). The van der Waals radius is larger than the covalent radius because it involves the interposition of two electron pairs between the atoms rather than one. The rule states that the van der Waals radius of an atom exceeds its covalent radius by ~0.08 nm (overlap in Figure 7.4).

Figure 7.4. Schematic of atoms undergoing covalent bonding.

Supporting Information XV

Thus, the size of polyatomic ions bound by covalent bonds, Ms, can now be estimated as the sum of the van der Waals radii subtracted by 0.08 nm:

Ms = 2·(rA - 0.08 nm) + 2·(rB - 0.08 nm) + (2)

We estimated the molecular size of the NACs (Table 7.2) by means of equation (2) and the van der Waals radii of the atoms in Table 7.1.

Table 7.1. Van der Waals radii of various atoms. Values from Pauling, 1960.

Atom vdW radii (nm) H 0.120 O 0.140 N 0.150 C 0.170 Cl 0.181 S 0.185

In order to make the calculations, it was assumed that all atoms were spherical and that all bond angles were 90° or 180° (linear structures). In addition, no distinctions were made between single and double bonds. The molecular sizes of the NACs were estimated with the benzene ring representing the xy plane.

Table 7.2. Molecular sizes of the NACs. a. Thickness z of the xy plane.

a Compound Ms (x) (nm) Ms (y) (nm) Ms (z) (nm) NB 0.54 0.80 0.36 4-NT 0.54 1.06 0.36 4-CNB 0.54 1.00 0.36 4-NPA 0.54 1.36 0.36

XVI Chapter 7 Note that the molecular sizes in Table 7.2 are only rough estimations.

For comparison with the GR-SO4 interlayer spacing (0.61 nm), we consider three possible orientations of the NACs in the GR-SO4 interlayer: 1) The NAC xyz coordination is equivalent to the crystal abc coordination (z = c = 0.36 nm), 2) the NAC xy plane is parallel to the crystal bc plane (z = a = 0.54 nm) and 3) the NAC xy plane is parallel to the crystal ac plane (z = b = 0.80-1.36 nm). Hence, the sizes of the NACs do not hinder their access to the GR-SO4 interlayer. Only when oriented vertically do the sizes of the NACs (z = b = 0.80-1.36 nm) exceed the GR-

SO4 interlayer spacing.

The molecular size of trichloroacetate (TCA) was also estimated by means of equation (2) and the atomic van der Waals radii in Table 7.1. When the TCA aliphatic chain is assumed to represent the x direction (Ms (x) = 0.66 nm), the molecular size in the y and z directions ranges from 0.45-0.53 nm depending on the free rotation of the C-C bond. Thus, only if the C-C bond is oriented perpendicular to the crystal ab plane does the size of TCA exceed the GR-SO4 interlayer spacing

(0.61 nm). In contrast, the size of TCA exceeds the GR-CO3 interlayer spacing (0.26 nm) regardless of its orientation.

7.5 Adsorption of Fe(II) onto Fe(III) oxides As seen from the Fe(II) sorption isotherms, Fe(II) sorption varies widely between the Fe(III) oxides as a function of solution pH (Figure 7.5). Average surface densities of approximately 2 singly coordinated sites/nm2 iron oxide have been suggested for goethite and lepidocrocite (Cornell & Schwertmann, 1996). The similar surface site densities of goethite and lepidocrocite might explain their similar Fe(II) adsorption isotherms (Figure 7.5).

Supporting Information XVII

Figure 7.5. Fe(II) adsorption edges for ferrihydrite, goethite, hematite, lepidocrocite and magnetite in the absence of other specifically adsorbing cations and anions (from Vikesland & Valentine, 2002 and references therein). The total number of surface sites was in excess of the total Fe(II) concentrations in all experiments.

Dissolved cations or anions may specifically adsorb at the calcite and Fe(III) oxide + – 0 0 0 surfaces by exchanging for H or OH at the ≡CO3H , ≡CaOH , ≡FeOH and ≡FeIIIOFeIIOH0 surface sites. At the experimental conditions applied here within a 2+ – pH range 7.0-8.7, the dominant species of interest in solution are Fe , HCO3 , 2– – 2– CO3 , Cl , SO4 (only in the GR-SO4 systems) and the anionic TCA and DCA. In addition, Fe2+ readily forms aqueous complexes with hydroxide, carbonate, + + – chloride and sulfate whereby the species FeOH , FeHCO3 , Fe(OH)(CO3) , 0 2– + 0 FeCO3 , Fe(CO3)2 , FeCl and FeSO4 may occur (Millero & Hawke, 1992). At 2+ 0 – + pH 7.0-8.7, we expect the Fe(II) species Fe , FeCO3 , Fe(OH)(CO3) , FeOH and 2– Fe(CO3)2 to dominate in the GR-CO3 and CaCO3(s)/CO2(g) buffered magnetite suspensions. In the goe/calcite and lep/calcite suspensions, we expect the FeCl+ 2+ 0 + species to dominate as well whereas the Fe , FeSO4 and FeOH species most

XVIII Chapter 7 likely dominate in the GR-SO4 suspensions. Anionic inorganic ligands like carbonate, chloride and sulfate can lower or enhance the adsorption of Fe(II) due to a) formation of stable nonadsorbing Fe(II) ligand aqueous complexes, b) formation of Fe(II) ligand Fe(III) oxide surface complexes, which can lead to surface precipitation at high Fe(II) and ligand concentrations, c) competitive ligand sorption to the Fe(III) oxide surface blocking reactive sorption sites at the surface, and d) diminution of the positive charge at the Fe(III) oxide surface (at pH levels below the point of zero charge (pHpzc) of the Fe(III) oxide) thereby decreasing the electrostatic repulsion of cations by the Fe(III) oxide surface. Specifically adsorbed cations increase the pHpzc whereas specifically adsorbed anions decrease the pHpzc.

References

Arnold, W.A.; Ball, W.P.; Roberts, A.L. (1999) Polychlorinated ethane reaction with zero-valent : Pathways and rate control. Journal of Contaminant Hydrology, 40, 183-200.

Cornell, R.M.; Schwertmann, U. (1996) The iron oxides. Structure, properties, reactions, occurrence and uses. VCH Verlagsgesellschaft mbH, Weinheim.

Fogler, H.S. (1999) Elements of engineering, 3rd ed., Prentice Hall.

Fuller, E.N.; Schettler, P.D.; Giddings, J.C. (1966) A new method for prediction of binary gas- phase diffusion coefficients. Industrial and Engineering Chemistry, 58, 19-27.

Hansen, H.C.B.; Koch, C.B. (1998) Reduction of nitrate to ammonium by sulphate green rust: Activation energy and interlayer reaction mechanism. Clay Minerals, 33, 87-101.

Harriott, P. (1962) Mass transfer to particles: Part I. Suspended in agitated tanks. AIChE Journal, 8, 93-102.

Hayduk, W; Laudie, H. (1974) Prediction of diffusion coefficients for nonelectrolytes in dilute aqueous solutions. AIChE Journal, 20, 611-615.

Hofstetter, T.B.; Heijmann, C.G.; Haderlein, S.B.; Holliger, C.; Schwarzenbach, R.P. (1999) Complete reduction of TNT and other (poly)nitroaromatic compounds under iron-reducing subsurface conditions. Environmental Science and Technology, 33, 1479-1487.

Klausen, J.; Tröber, S.P.; Haderlein, S.B.; Schwarzenbach, R.P. (1995) Reduction of substituted nitrobenzenes by Fe(II) in aqueous mineral suspensions. Environmental Science and Technology, 29, 2396-2404.

Meisel, D.; Neta, P. (1975) One-electron redox potentials of nitro compounds and radiosensitizers. Correlation with spin densities of their radical anions. Journal of the American Chemical Society, 97, 5198-5203.

Supporting Information XIX

Millero, F.J.; Hawke, D.J. (1992) Ionic interactions of divalent metals in natural waters. Marine Chemistry, 40, 19-48.

Pauling, L. (1960) The nature of the chemical bond. 3rd ed. Cornell University Press, Ithaca.

Vikesland, P.J.; Valentine, R.L. (2002) Iron oxide surface-catalyzed oxidation of ferrous iron by monochloramine: implications of oxide type and carbonate on reactivity. Environmental Science and Technology, 36, 512-519.

Wardman, P. (1989) Reduction potentials of one-eletron couples involving free radicals in aqueous solution. The Journal of Physical Chemistry Reference Data, 18, 1637-1755.

Curriculum Vitae

13.01.1973 Born in Haderslev, Denmark

1988-1991 Mathematical high school, Haderslev Katedralskole, Denmark

1991-1992 Sabbatical year

1992-1995 B. Sc. in environmental chemistry, University of Copenhagen (KU), Denmark. B. Sc. thesis, 1995: “Methane oxidizing bacteria in soil”.

1995-1998 M. Sc. in environmental chemistry, University of Copenhagen, Denmark 1997-1998, diploma thesis: “Reductive dechlorination of carbon tetrachloride and chloroform in presence of iron(II)iron(III)-hydroxides (green rust)”.

1998-1999 Research and teaching assistant at the Chemistry Department, The Royal Veterinary & Agricultural University (KVL), Denmark.

1999-2004 Ph.D. in environmental sciences, Swiss Federal Institute of Technology Zürich (ETHZ) and Swiss Federal Institute for Environmental Science and Technology (EAWAG), Switzerland. Docoral thesis: “Formation and redox reactions of green rusts under geochemical conditions found in natural soils and sediments”.

2000-2002 Teaching assistent at the Swiss Federal Institute of Technology Zürich and supervision of diploma students.

2002 Microbial Diversity summer course (7 weeks) at the Marine Biological Laboratory, Woods Hole, Massachusetts, USA.