Secondary mineralization pathways induced by dissimilatory iron reduction of ferrihydrite under advective flow

Iron (hydr)oxides not only serve as potent sorbents and repositories for nutrients and contaminants but also provide a terminal electron acceptor for microbial respiration. The microbial reduction of Fe (hydr)oxides and the subsequent secondary solid-phase transformations will, therefore, have a profound influence on the biogeochemical cycling of Fe as well as associated metals. Here we elucidate the pathways and mechanisms of secondary mineralization during dissimilatory iron reduction by a common iron-reducing bacterium, Shewanella putrefaciens (strain CN32), of 2-line ferrihydrite under advective flow conditions. Secondary mineralization of ferrihydrite occurs via a coupled, biotic-abiotic pathway primarily resulting in the production of magnetite and goethite with minor amounts of green rust. Operating mineralization pathways are driven by competing abiotic reactions of bacterially generated ferrous iron with the ferrihydrite surface. Subsequent to the initial sorption of ferrous iron on ferrihydrite, goethite (via dissolution/reprecipitation) and/or magnetite (via solid-state conversion) precipitation ensues resulting in the spatial coupling of both goethite and magnetite with the ferrihydrite surface. The distribution of goethite and magnetite within the column is dictated, in large part, by flow-induced ferrous Fe profiles. While goethite precipitation occurs over a large Fe(II) concentration range, magnetite accumulation is only observed at concentrations exceeding 0.3 mmol/L (equivalent to 0.5 mmol Fe[II]/g ferrihydrite) following 16 d of reaction. Consequently, transport-regulated ferrous Fe profiles result in a progression of magnetite levels downgradient within the column. Declining microbial reduction over time results in lower Fe(II) concentrations and a subsequent shift in magnetite precipitation mechanisms from nucleation to crystal growth. While the initial precipitation rate of goethite exceeds that of magnetite, continued growth is inhibited by magnetite formation, potentially a result of lower Fe(III) activity. Conversely, the presence of lower initial Fe(II) concentrations followed by higher concentrations promotes goethite accumulation and inhibits magnetite precipitation even when Fe(II) concentrations later increase, thus revealing the importance of both the rate of Fe(II) generation and flow-induced Fe(II) profiles. As such, the operating secondary mineralization pathways following reductive dissolution of ferrihydrite at a given pH are governed principally by flow-regulated Fe(II) concentration, which drives mineral precipitation kinetics and selection of competing mineral pathways.     

纳米铁还原高浓度硝基苯的实验

采用液相还原法制备纳米铁粒子,粒径约为50 nm,主要成分为ɑ-Fe。实验研究了纳米铁还原高浓度硝基苯的效果及反应动力学。结果表明:自制的纳米铁具有很高的活性,短时间内能将高浓度硝基苯彻底还原,效果远优于普通铁粉。当摩尔比满足n纳米铁∶n硝基苯=4∶1时,纳米铁还原硝基苯属于一级反应动力学,反应速率常数和半衰期分别为0.159 3 h-1和4.351 2 h。整个实验过程中,ρ(DO)在0.5 mg/L以下,pH值约为9.0,体系呈弱碱性的厌氧环境,有利于厌氧微生物的生长。纳米铁反应后的产物为Fe6(OH)12CO3和MgFe3O4,且以Fe6(OH)12CO3为主。纳米铁反应后较以前比粒径变化不大,分散性较好,但表面被严重腐蚀。     

Effects of humic substances on Fe(II) sorption onto aluminum oxide and clay

We studied the effects of humic substances (HS) on the sorption of Fe(II) onto Al-oxide and clay sorbents at pH 7.5 with a combination of batch kinetic experiments and synchrotron Fe K-edge EXAFS analyses. Fe(II) sorption was monitored over the course of 4 months in anoxic clay and Al-oxide suspensions amended with variable HS types (humic acid, HA; or fulvic acid, FA) and levels (0, 1, and 4 wt%), and with differing Fe(II) and HS addition sequences (co-sorption and pre-coated experiments, where Fe(II) sorbate was added alongside and after HS addition, respectively). In the Al-oxide suspensions, the presence of HS slowed down the kinetics of Fe(II) sorption, but had limited, if any, effect on the equilibrium aqueous Fe(II) concentrations. EXAFS analyses revealed precipitation of Fe(II)–Al(III)-layered double hydroxide (LDH) phases as the main mode of Fe(II) sorption in both the HA-containing and HA-free systems. These results demonstrate that HS slow down Fe(II) precipitation in the Al-oxide suspensions, but do not affect the composition or stability of the secondary Fe(II)–Al(III)-LDH phases formed. Interference of HS with the precipitation of Fe(II)–Al(III)-LDH was attributed to the formation organo-Al complexes HS limiting the availability of Al for incorporation into secondary layered Fe(II)-hydroxides. In the clay systems, the presence of HA caused a change in the main Fe(II) sorption product from Fe(II)–Al(III)-LDH to a Fe(II)-phyllosilicate containing little structural Al. This was attributed to complexation of Al by HA, in combination with the presence of dissolved Si in the clay suspension enabling phyllosilicate precipitation. The change in Fe(II) precipitation mechanism did not affect the rate of Fe(II) sorption at the lower HA level, suggesting that the inhibition of Fe(II)–Al(III)-LDH formation in this system was countered by enhanced Fe(II)-phyllosilicate precipitation. Reduced rates of Fe(II) sorption at the higher HA level were attributed to surface masking or poisoning by HA of secondary Fe(II) mineral growth at or near the clay surface. Our results suggest that HS play an important role in controlling the kinetics and products of Fe(II) precipitation in reducing soils, with effects modulated by soil mineralogy, HS content, and HS properties. Further work is needed to assess the importance of layered Fe(II) hydroxides in natural reducing environments.     

Biomineralization of lepidocrocite and goethite by nitrate-reducing Fe(II)-oxidizing bacteria: Effect of pH, bicarbonate, phosphate, and humic acids

Fe(III) solid phases are the products of Fe(II) oxidation by Fe(II)-oxidizing bacteria, but the Fe(III) phases reported to form within growth experiments are, at times, poorly crystalline and therefore difficult to identify, possibly due to the presence of ligands (e.g., phosphate, carbonate) that complex iron and disrupt iron (hydr)oxide precipitation. The scope of this study was to investigate the influences of geochemical solution conditions (pH, carbonate, phosphate, humic acids) on the Fe(II) oxidation rate and Fe(III) mineralogy. Fe(III) mineral characterization was performed using ⁵⁷Fe-Mössbauer spectroscopy and μ-X-ray diffraction after oxidation of dissolved Fe(II) within Mops-buffered cell suspensions of Acidovorax sp. BoFeN1, a nitrate-reducing, Fe(II)-oxidizing bacterium. Lepidocrocite (γ-FeOOH) (90%), which also forms after chemical oxidation of Fe(II) by dissolved O₂, and goethite (α-FeOOH) (10%) were produced at pH 7.0 in the absence of any strongly complexing ligands. Higher solution pH, increasing concentrations of carbonate species, and increasing concentrations of humic acids promoted goethite formation and caused little or no changes in Fe(II) oxidation rates. Phosphate species resulted in Fe(III) solids unidentifiable to our methods and significantly slowed Fe(II) oxidation rates. Our results suggest that Fe(III) mineralogy formed by bacterial Fe(II) oxidation is strongly influenced by solution chemistry, and the geochemical conditions studied here suggest lepidocrocite and goethite may coexist in aquatic environments where nitrate-reducing, Fe(II)-oxidizing bacteria are active.     

Arsenic redox transformation by humic substances and Fe minerals

The toxicity and mobility of the redox-active metalloid As strongly depends on its oxidation state, with As(III) (arsenite) being more toxic and mobile than As(V) (arsenate). It is, therefore, necessary to know the biogeochemical processes potentially influencing As redox state to understand and predict its environmental behavior. The first part of this presentation will discuss the quantification of As redox changes by pH-neutral mineral suspensions of goethite [α-Fe^IIIOOH] amended with Fe(II) using wet-chemical and synchrotron X-ray absorption (XANES) analysis (Amstaetter et al., 2010). First, it was found that goethite itself did not oxidize As(III). Second, in contrast to thermodynamic predictions, Fe(II)–goethite systems did not reduce As(V). However, surprisingly, rapid oxidation of As(III) to As(V) was observed in Fe(II)–goethite systems. Iron speciation and mineral analysis by Mössbauer spectroscopy showed rapid formation of ⁵⁷Fe–goethite after ⁵⁷Fe(II) addition and the formation of a so far unidentified additional Fe(II) phase. No other Fe(III) phase could be detected by Mössbauer spectroscopy, EXAFS, scanning electron microscopy, X-ray diffraction or high-resolution transmission electron microscopy. This suggests that reactive Fe(III) species form as an intermediate Fe(III) phase upon Fe(II) addition and electron transfer into bulk goethite but before crystallization of the newly formed Fe(III) as goethite.The second part of the presentation will show that semiquinone radicals produced during microbial or chemical reduction of a humic substance model quinone (AQDS, 9,10-anthraquinone-2,6-disulfonic acid) can react with As and change its redox state (Jiang et al., 2009). The results of these experiments showed that these semiquinone radicals are strong oxidants and oxidize arsenite to arsenate, thus decreasing As toxicity and mobility. The oxidation of As(III) depended strongly on pH. More arsenite (up to 67.3%) was oxidized at pH 11 compared to pH 7 (12.6% oxidation) and pH 3 (0.5% oxidation). In addition to As(III) oxidation by semiquinone radicals, hydroquinones that were also produced during quinone reduction, reduced As(V) to As(III) at neutral and acidic pH values (less than 12%) but not at alkaline pH. In an attempt to understand the observed redox reactions between As and reduced/oxidized quinones present in humic substances, the radical content in reduced AQDS solutions was quantified and Eh-pH diagrams were constructed. Both the radical quantification and the Eh-pH diagram allowed explaining the observed redox reactions between the reduced AQDS solutions and the As.In summary these studies indicate that in the simultaneous presence of Fe(III) oxyhydroxides, Fe(II), and humic substances as commonly observed in environments inhabited by Fe-reducing microorganisms, As(III) oxidation can occur. This potentially explains the presence of As(V) in reduced groundwater aquifers.     

Chemical and physical properties of iron hydroxide precipitates associated with passively treated coal mine drainage in the Bituminous Region of Pennsylvania and Maryland

Changes in precipitate mineralogy, morphology, and major and trace element concentrations and associations throughout 5 coal mine drainage (CMD) remediation systems treating discharges of varying chemistries were investigated in order to determine the factors that influence the characteristics of precipitates formed in passive systems. The 5 passive treatment systems sampled in this study are located in the bituminous coal fields of western Pennsylvania and northern Maryland, and treat discharges from Pennsylvanian age coals. The precipitates are dominantly (>70%) goethite. Crystallinity varies throughout an individual system, and lower crystallinity is associated with enhanced sorption of trace metals. Degree of crystallinity (and subsequently morphology and trace metal associations) is a function of the treatment system and how rapidly Fe(II) is oxidized, forms precipitates, aggregates and settles. Precipitates formed earlier in the passive treatment systems tend to have the highest crystallinity and the lowest concentrations of trace metal cations. High surface area and cation vacancies within the goethite structure enable sorption and incorporation of metals from coal mine drainage-polluted waters. Sorption affinities follow the order of Zn > Co ≈ Ni > Mn. Cobalt and Ni are preferentially sorbed to Mn oxide phases when these phases are present. As pH increases in the individual CMD treatment systems toward the pH_pzc of goethite, As sorption decreases and transition metal (Co, Mn, Ni and Zn) sorption increases. Sulfate, Na and Fe(II) concentrations may all influence the sorption of trace metals to the Fe hydroxide surface. Results of this study have implications not only for solids disposal and resource recovery but also for the optimization of passive CMD treatment systems.     

Iron isotope fractionation and atom exchange during sorption of ferrous iron to mineral surfaces

The application of stable Fe isotopes as a tracer of the biogeochemical Fe cycle necessitates a mechanistic knowledge of natural fractionation processes. We studied the equilibrium Fe isotope fractionation upon sorption of Fe(II) to aluminum oxide (γ-Al₂O₃), goethite (α-FeOOH), quartz (α-SiO₂), and goethite-loaded quartz in batch experiments, and performed continuous-flow column experiments to study the extent of equilibrium and kinetic Fe isotope fractionation during reactive transport of Fe(II) through pure and goethite-loaded quartz sand. In addition, batch and column experiments were used to quantify the coupled electron transfer-atom exchange between dissolved Fe(II) (Fe(II)_aq) and structural Fe(III) of goethite. All experiments were conducted under strictly anoxic conditions at pH 7.2 in 20 mM MOPS (3-(N-morpholino)-propanesulfonic acid) buffer and 23 °C. Iron isotope ratios were measured by high-resolution MC-ICP-MS. Isotope data were analyzed with isotope fractionation models. In batch systems, we observed significant Fe isotope fractionation upon equilibrium sorption of Fe(II) to all sorbents tested, except for aluminum oxide. The equilibrium enrichment factor, , of the Fe(II)_sorb-Fe(II)_aq couple was 0.85 ± 0.10‰ (±2σ) for quartz and 0.85 ± 0.08‰ (±2σ) for goethite-loaded quartz. In the goethite system, the sorption-induced isotope fractionation was superimposed by atom exchange, leading to a δ^56/54Fe shift in solution towards the isotopic composition of the goethite. Without consideration of atom exchange, the equilibrium enrichment factor was 2.01 ± 0.08‰ (±2σ), but decreased to 0.73 ± 0.24‰ (±2σ) when atom exchange was taken into account. The amount of structural Fe in goethite that equilibrated isotopically with Fe(II)_aq via atom exchange was equivalent to one atomic Fe layer of the mineral surface (∼3% of goethite-Fe). Column experiments showed significant Fe isotope fractionation with δ^56/54Fe(II)_aq spanning a range of 1.00‰ and 1.65‰ for pure and goethite-loaded quartz, respectively. Reactive transport of Fe(II) under non-steady state conditions led to complex, non-monotonous Fe isotope trends that could be explained by a combination of kinetic and equilibrium isotope enrichment factors. Our results demonstrate that in abiotic anoxic systems with near-neutral pH, sorption of Fe(II) to mineral surfaces, even to supposedly non-reactive minerals such as quartz, induces significant Fe isotope fractionation. Therefore we expect Fe isotope signatures in natural systems with changing concentration gradients of Fe(II)_aq to be affected by sorption.     

Kinetics of zinc and arsenate co-sorption at the goethite-water interface

Little or no information is available in the literature about reaction processes of co-sorbing metals and arsenate [As(V)] on variable-charged surfaces or factors influencing these reactions. Arsenic and metal contamination are, however, a common co-occurrence in many contaminated environments. In this study, we investigated the co-sorption kinetics of 250 μM As(V) and zinc [Zn(II)] in 10, 100, and 1000 mg goethite L⁻¹ 0.01 M NaCl solution at pH 7, collected complementary As and Zn K-edge extended X-ray absorption fine structure (EXAFS) data after various aging times, and performed a replenishment desorption/dissolution study at pH 4 and 5.5 after 6 months of aging time. Arsenate and Zn(II) formed adamite-like and koritnigite-like precipitates on goethite in 100- and 10-ppm goethite suspensions, respectively, whereas in 1000-ppm goethite suspensions, As(V) formed mostly double-corner sharing complexes and Zn(II) formed a solid solution on goethite according to EXAFS spectroscopic analyses. In all goethite suspension densities, surface adsorption reactions were part of the initial reaction processes. In 10- and 100-ppm goethite suspensions, a heterogeneous nucleation reaction occurred in which adamite-like precipitates began to form 48 h earlier than koritnigite-like surface precipitates. Arsenate and Zn(II) uptake from solution decreased after 4 weeks. Replenishment desorption studies showed that the precipitates and surface adsorbed complexes on goethite were susceptible to proton-promoted dissolution resulting in many cases in more than 80% loss of Zn(II) and ∼ 60% to 70% loss of arsenate. The molar Zn:As dissolution ratio was dependent on the structure of the precipitate and was cyclic for the adamite and koritnigite-like surface precipitates, reflecting the concentric and plane-layered structures of adamite and koritnigite, respectively.     

Effect of hydroxamate siderophores on Fe release and Pb(II) adsorption by goethite

Hydroxamate siderophores are biologically-synthesized, Fe(III)-specific ligands which are common in soil environments. In this paper, we report an investigation of their adsorption by the iron oxyhydroxide, goethite; their influence on goethite dissolution kinetics; and their ability to affect Pb(II) adsorption by the goethite surface. The siderophores used were desferrioxamine B (DFO-B), a fungal siderophore, and desferrioxamine D₁, an acetyl derivative of DFO-B (DFO-D1). Siderophore adsorption isotherms yielded maximum surface concentrations of 1.5 (DFO-B) or 3.5 (DFO-D1) μmol/g at pH 6.6, whereas adsorption envelopes showed either cation-like (DFO-B) or ligand-like (DFO-D1) behavior. Above pH 8, the adsorbed concentrations of both siderophores were similar. The dissolution rate of goethite in the presence of 240 μM DFO-B or DFO-D1 was 0.02 or 0.17 μmol/g hr, respectively. Comparison of these results with related literature data on the reactions between goethite and acetohydroxamic acid, a monohydroxamate ligand, suggested that the three hydroxamate groups in DFO-D1 coordinate to Fe(III) surface sites relatively independently. The results also demonstrated a significant depleting effect of 240 μM DFO-B or DFO-D1 on Pb(II) adsorption by goethite at pH > 6.5, but there was no effect of adsorbed Pb(II) on the goethite dissolution rate.     

Structural constraints of ferric (hydr)oxides on dissimilatory iron reduction and the fate of Fe(II)

Due to the strong reducing capacity of ferrous Fe, the fate of Fe(II) following dissimilatory iron reduction will have a profound bearing on biogeochemical cycles. We have previously observed the rapid and near complete conversion of 2-line ferrihydrite to goethite (minor phase) and magnetite (major phase) under advective flow in an organic carbon-rich artificial groundwater medium. Yet, in many mineralogically mature environments, well-ordered iron (hydr)oxide phases dominate and may therefore control the extent and rate of Fe(III) reduction. Accordingly, here we compare the reducing capacity and Fe(II) sequestration mechanisms of goethite and hematite to 2-line ferrihydrite under advective flow within a medium mimicking that of natural groundwater supplemented with organic carbon. Introduction of dissolved organic carbon upon flow initiation results in the onset of dissimilatory iron reduction of all three Fe phases (2-line ferrihydrite, goethite, and hematite). While the initial surface area normalized rates are similar (∼10⁻¹¹ mol Fe(II) m⁻² g⁻¹), the total amount of Fe(III) reduced over time along with the mechanisms and extent of Fe(II) sequestration differ among the three iron (hydr)oxide substrates. Following 16 d of reaction, the amount of Fe(III) reduced within the ferrihydrite, goethite, and hematite columns is 25, 5, and 1%, respectively. While 83% of the Fe(II) produced in the ferrihydrite system is retained within the solid-phase, merely 17% is retained within both the goethite and hematite columns. Magnetite precipitation is responsible for the majority of Fe(II) sequestration within ferrihydrite, yet magnetite was not detected in either the goethite or hematite systems. Instead, Fe(II) may be sequestered as localized spinel-like (magnetite) domains within surface hydrated layers (ca. 1 nm thick) on goethite and hematite or by electron delocalization within the bulk phase. The decreased solubility of goethite and hematite relative to ferrihydrite, resulting in lower Fe(III)_aq and bacterially-generated Fe(II)_aq concentrations, may hinder magnetite precipitation beyond mere surface reorganization into nanometer-sized, spinel-like domains. Nevertheless, following an initial, more rapid reduction period, the three Fe (hydr)oxides support similar aqueous ferrous iron concentrations, bacterial populations, and microbial Fe(III) reduction rates. A decline in microbial reduction rates and further Fe(II) retention in the solid-phase correlates with the initial degree of phase disorder (high energy sites). As such, sustained microbial reduction of 2-line ferrihydrite, goethite, and hematite appears to be controlled, in large part, by changes in surface reactivity (energy), which is influenced by microbial reduction and secondary Fe(II) sequestration processes regardless of structural order (crystallinity) and surface area.     

Reversible surface-sorption-induced electron-transfer oxidation of Fe(II) at reactive sites on a synthetic clay mineral

The sorption of ⁵⁷Fe(II) onto an Fe-free, mineralogically pure and Ca-saturated synthetic montmorillonite sample (structural formula: Ca_0.15(Al_1.4Mg_0.6)(Si₄)O₁₀(OH,F)₂), was studied as a function of pH under strictly anoxic conditions (N₂ glove box atmosphere, O₂ content <1 ppm), using wet chemistry and cryogenic (T = 77 K) ⁵⁷Fe Mössbauer spectrometry. No Fe(III) was detected in solution at any pH. However, in pH conditions where Fe(II) is removed from solution, a significant amount of surface-bound Fe(III) was produced, which increased with pH from 0% to 3% of total Fe in a pre-sorption edge region (i.e. at pH < 7.5 where about 15% of total Fe is sorbed) to 7% of total Fe when all Fe is sorbed. At low pH, where the pre-sorption edge plateau occurs (2 < pH < 7.5), the total sorbed-Fe amount remained constant but, within this sorbed-Fe pool, the Fe(III)/Fe(II) ratio increased with pH, from 0.14 at pH 2 up to 0.74 at pH 7. The pre-sorption edge plateau is interpreted as cation exchange on interlayer surfaces together with a sorption phenomenon occurring on highly reactive (i.e. high affinity) surface sites. As pH increases and protons are removed from the clay edge surface, we propose that more and more of these highly reactive sites acquire a steric configuration that stabilizes Fe(III) relative to Fe(II), thereby inducing a Fe to clay particle electron transfer. A sorption model based on cation exchange combined with surface complexation and electron transfers reproduces both wet chemical as well as the Mössbauer spectrometric results. The mechanism is fully reversible: sorbed-Fe is reduced as pH decreases (Mössbauer solid-state analyses) and all Fe returned to solution is returned as Fe(II) (solution analyses). This would not be the case if the observed oxidations were due to contaminant oxidizing agents in solution. The present work shows that alternating pH may induce surface redox phenomena in the absence of an electron acceptor in solution other than H₂O.     

Influence of microbial hydroxamate siderophores on Pb(II) desorption from α-FeOOH

Siderophores are low-molecular weight organic molecules secreted by plants and micro-organisms in response to Fe stress. With stability constants commonly exceeding 10³⁰, siderophores are considered to have higher affinities for Fe(III) than for any other major or trace element dissolved in soil solution. However, several siderophores have affinities for trace metals that approach those for Fe(III), and certain actinides form siderophore complexes of surprisingly high stability. The purpose of this study was to examine the role of hydroxamate siderophores in controlling Pb sorption to an Fe(III) oxide adsorbent. Goethite [α-FeOOH], prepared by standard methods and identified by X-ray diffraction, gave a specific surface of 36 m² g⁻¹ as determined by N₂ multipoint BET analysis. Adsorption experiments were performed aseptically using a batch method with a goethite concentration of 1.0 g l⁻¹ and an ionic strength of 0.01 M NaClO₄. Soluble Pb and Fe were measured between pH 3 and 8 by first adding Pb (10 μM) and then siderophore (10, 20, or 40 μM) to the goethite suspension. Three hydroxamate siderophores were employed: desferrioxamine B (DFB), ferrichrome (FC), and rhodotorulic acid (RA). Following 20 h reaction, Pb and Fe in solution were measured by ICP–MS and ICP–AES, respectively. The efficacy of siderophore-mediated Pb desorption varied with siderophore type and generally increased with pH and siderophore/Pb molar ratio. Desferrioxamine B, at pH 6.5 and a DFB/Pb molar ratio of 4, solubilised nearly 25% of the total sorbed Pb. In the presence of 10 μM FC, Pb adsorption largely mimicked that for the siderophore-free system, whereas significant amounts of Pb were desorbed with 20 μM FC at pH >5.5. The dihydroxamate siderophore, RA, was the least effective Pb chelator, requiring 20 μM to desorb detectable amounts of Pb.     

Mineralogical and morphological constraints on the reduction of Fe(III) minerals by Geobacter sulfurreducens

The biologically-mediated reduction of synthetic samples of the Fe(III)-bearing minerals hematite, goethite, lepidocrocite, feroxhyte, ford ferrihydrite, akaganeite and schwertmannite by Geobacter sulfurreducens has been investigated using microbiological techniques in conjunction with X-ray Diffraction (XRD), Transmission Electron Microscopy (TEM) and X-ray Photoelectron Spectroscopy (XPS). This combination of approaches offers unique insights into the influence of subtle variations in the crystallinity of a given mineral on biogeochemical processes, and has highlighted the importance of (oxyhydr)oxide crystallite morphology in determining the changes occurring in a given mineral phase. Problems arising from normalising the biological Fe(III) reduction rates relative to the specific surface areas of the starting materials are also highlighted. These problems are caused primarily by particle aggregation, and compounded when using spectrophotometric assays to monitor reduction. For example, the initial rates of Fe(III) reduction observed for two synthetic feroxyhytes with different crystallinities (as shown by XRD and TEM studies) but almost identical surface areas, differ substantially. Both microbiological and high-resolution TEM studies show that hematite and goethite are susceptible to limited amounts of Fe(III) reduction, as evidenced by the accumulation of Fe(II) during incubation with G. sulfurreducens and the growth of nodular structures on crystalline goethite laths during incubation. Lepidocrocite and akaganeite readily transform into mixtures of magnetite and goethite, and XRD data indicate that the proportion of magnetite increases within the transformation products as the crystallinity of the starting material decreases. The presence of anthraquinone-2,6-disulfonate (AQDS) as an electron shuttle increases both the initial rate and longer term extent of biological Fe(III) reduction for all of the synthetic minerals examined. High-resolution XPS indicates subtle but measurable differences in the Fe(III):Fe(II) ratios at the mineral surfaces following extended incubation. For example, for a poorly crystalline schwertmannite, deconvolution of the Fe2p_3/2 peak suggests that the Fe(III):Fe(II) ratio of the near-surface regions varies from 1.0 in the starting material to 0.9 following 144 h of incubation with G.sulfurreducens, and to 0.75 following the same incubation period in the presence of 10 μM AQDS. These results have important implications for the biogeochemical cycling of iron.     

Formation of Fe(III) oxyhydroxide colloids in freshwater and brackish seawater, with incorporation of phosphate and calcium

The formation of Fe(III) oxyhydroxide colloids by oxidation of Fe(II) and their subsequent aggregation to larger particles were studied in laboratory experiments with natural water from a freshwater lake and a brackish coastal sea. Phosphate was incorporated in the solid phase during the course of hydrolysis of iron. The resulting precipitated amorphous Fe(III) oxyhydroxide phases were of varying composition, depending primarily on the initial dissolved Fe/P molar ratio, but with little influence by salinity or concentration of calcium ions. The lower limiting Fe/P ratio found for the solid phase suggests the formation of a basic Fe(III) phosphate compound with a stoichiometric Fe/P ratio of close to two. This implies that an Fe/P stoichiometry of ≈2 ultimately limits the capacity of precipitating Fe(III) to fix dissolved phosphate at oxic/anoxic boundaries in natural waters. In contrast to phosphorus, the uptake of calcium seemed to be controlled by sorption processes at the surface of the iron-rich particles formed. This uptake was more efficient in freshwater than in brackish water, suggesting that salinity restrains the uptake of calcium by newly formed Fe(III) oxyhydroxides in natural waters. Moreover, salinity enhanced the aggregation rate of the colloids formed. The suspensions were stabilised by the presence of organic matter, although this effect was less pronounced in seawater than in freshwater. Thus, in seawater of 6 to 33 ‰S, the removal of particles was fast (removal half time < 200 h), whereas the colloidal suspensions formed in freshwater were stable (removal half time > 900 h). Overall, oxidation of Fe(II) and removal of Fe(III) oxyhydroxide particles were much faster in seawater than in freshwater. This more rapid turnover results in lower iron availability in coastal seawater than in freshwater, making iron more likely to become a limiting element for chemical scavenging and biologic production.     

An iron isotope signature related to electron transfer between aqueous ferrous iron and goethite

We measured the Fe isotope fractionation during the reactions of Fe(II) with goethite in the presence and absence of a strong Fe(III) chelator (desferrioxamine mesylate, DFAM). All experiments were completed in an O₂-free glove box. The concentrations of aqueous Fe(II) ([Fe(II)_aq]) decreased below the initial total dissolved Fe concentrations ([Fe(II)_total], 2.15 mM) due to fast adsorption within 0.2 day. The concentration of adsorbed Fe(II) ([Fe(II)_ads]) was determined as the difference between [Fe(II)_aq] and the concentration of extracted Fe(II) in 0.5 M HCl ([Fe(II)_extr]) (i.e., [Fe(II)_ads] = [Fe(II)_extr] − [Fe(II)_aq]). [Fe(II)_ads] also decreased with time in experiments with and without DFAM, documenting that fast adsorption was accompanied by a second, slower reaction. Interestingly, [Fe(II)_extr] was always smaller than [Fe(II)_total], indicating that some Fe(II) was sequestered into a pool that is not HCl-extractable. The difference was attributed to Fe(II) incorporated into goethite structure (i.e., [Fe(II)_inc] = [Fe(II)_total] −[Fe(II)_extr]). More Fe(II) was incorporated in the presence of DFAM than in its absence at all time steps. Regardless of the presence of DFAM, both aqueous and extracted Fe(II) (δ^56/54Fe(II)_aq and δ^56/54Fe(II)_extr) became isotopically lighter than or similar to goethite (− 0.27‰) at day 7, implying that the isotope exchange occurred between bulk goethite and aqueous Fe. Consistently, the mass balance indicated that the incorporated Fe is isotopically heavier than extracted Fe. These observations suggested that (i) co-adsorption of Fe(II) with DFAM resulted in more pervasive electron transfer, (ii) the electron transfer from heavy Fe(II) in the adsorbed Fe(II) to light Fe(III) in goethite results in the fixation of heavy adsorbed Fe(III) on the surface and accumulation of Fe(II) within the goethite, and (iii) desorption of the reduced, light Fe from goethite does not necessarily occur at the same surface sites where adsorption occurred.     

Adsorption of aromatic carboxylate compounds on the surface of synthesized iron oxide-coated sands

The complex mineral assemblages of silica and Fe minerals play a significant role in the transport of compounds in soils and sediments. Five coated sands including Goethite, Lepidocrocite, Ferrihydrite, Hematite and Magnetite were synthesized by a heterogeneous suspension method and characterized by FTIR spectroscopy, XRD, BET surface area and chemical analyses. The synthesis results showed that the degree of coating (mg Fe/g sand) varied with the mineralogy of Fe coating phases, which may have different affinities towards the silica surface. Batch experiments were conducted with two compounds (2,5-dihydroxybenzoic acid and 1-hydroxy-2-naphthoic acid) to quantify the contributions to adsorption from different oxide coatings and compare adsorption characteristics of selected organic acids. Sorption of these compounds to coated sands was examined versus a wide range of conditions (time, pH, ionic strength and sorbate concentration). Because of the attachment of Fe oxide, the coated sand had higher specific surface area, involving a better adsorption efficiency of organic compounds. Mineral surface charge and pH proved to be important for the adsorption of these compounds. The batch results indicated that the degree of coating was the most significant factor enhancing the sorption of aromatic compounds on the surface of sand and the mineralogy of the Fe phase was of less importance.     

Reactivity at (nano)particle-water interfaces,redox processes,and arsenic transport in the environment

Massive deleterious impacts to human health are resulting from the use of arsenic-bearing groundwaters in South-East Asia deltas and elsewhere in the world for drinking, cooking and/or irrigation. In Bangladesh alone, a fifth of all deaths are linked to arsenicosis. In the natural and engineered subsurface environment, the fate of arsenic is, to a large extent, controlled by redox potential, pH, as well as total iron, sulfur and carbonate content, via sorption and coprecipitation on a variety of natural and engineered (nano)particles. In the present article, we address: (1) new insights in the sorption mechanisms of As on Fe(II) and Fe(III) nanophases recognized to play an important role in the microbial cycling of As and Fe; (2) artifacts often encountered in field and laboratory studies of As speciation due to the extreme redox sensitivity of the Fe-As-O-H phases; and (3) as a conclusion, the implications for water treatment. Indeed the specific reactivity of nanoparticles accounts not only for the As bioavailability within soils and aquifers, but also opens new avenues in water treatment.     

Mössbauer spectroscopic study of iron in Victorian brown coal

The forms of non-pyritic Fe in a suite of Victorian brown coals have been determined by ⁵⁷Fe Mössbauer analysis. The dominant Fe phase is a poorly-ordered ferric oxyhydroxide with a magnetic ordering temperature of (530 ± 50) K and particle size of approximately 50 Å. Upon exposure of the coal to air, this phase slowly crystallises to goethite. Most of the remaining Fe occurs as a high-spin Fe(II) species attributed to dissolved and hence mobile, Fe(II) humate, which precipitates as the ferric oxyhydroxide to an extent determined by the pH. A third species, present in a much lower concentration, appears to exhibit a transition from low-spin to high-spin Fe(II) as water is removed from the coal.     

Hydrous ferric oxide precipitation in the presence of nonmetabolizing bacteria: Constraints on the mechanism of a biotic effect

We have used room temperature and cryogenic ⁵⁷Fe Mössbauer spectroscopy, powder X-ray diffraction (pXRD), mineral magnetometry, and transmission electron microscopy (TEM), to study the synthetic precipitation of hydrous ferric oxides (HFOs) prepared either in the absence (abiotic, a-HFO) or presence (biotic, b-HFO) of nonmetabolizing bacterial cells (Bacillus subtilis or Bacillus licheniformis, ∼10⁸ cells/mL) and under otherwise identical chemical conditions, starting from Fe(II) (10⁻², 10⁻³, or 10⁻⁴ mol/L) under open oxic conditions and at different pH (6-9). We have also performed the first Mössbauer spectroscopy measurements of bacterial cell wall (Bacillus subtilis) surface complexed Fe, where Fe(III) (10^−3.5-10^−4.5 mol/L) was added to a fixed concentration of cells (∼10⁸ cells/mL) under open oxic conditions and at various pH (2.5-4.3). We find that non-metabolic bacterial cell wall surface complexation of Fe is not passive in that it affects Fe speciation in at least two ways: (1) it can reduce Fe(III) to sorbed-Fe²⁺ by a proposed steric and charge transfer effect and (2) it stabilizes Fe(II) as sorbed-Fe²⁺ against ambient oxidation. The cell wall sorption of Fe occurs in a manner that is not compatible with incorporation into the HFO structure (different coordination environment and stabilization of the ferrous state) and the cell wall-sorbed Fe is not chemically bonded to the HFO particle when they coexist (the sorbed Fe is not magnetically polarized by the HFO particle in its magnetically ordered state). This invalidates the concept that sorption is the first step in a heterogeneous nucleation of HFO onto bacterial cell walls. Both the a-HFOs and the b-HFOs are predominantly varieties of ferrihydrite (Fh), often containing admixtures of nanophase lepidocrocite (nLp), yet they show significant abiotic/biotic differences: Biotic Fh has less intraparticle (including surface region) atomic order (Mössbauer quadrupole splitting), smaller primary particle size (magnetometry blocking temperature), weaker Fe to particle bond strength (Mössbauer center shift), and no six-line Fh (6L-Fh) admixture (pXRD, magnetometry). Contrary to current belief, we find that 6L-Fh appears to be precipitated directly, under a-HFO conditions, from either Fe(II) or Fe(III), and depending on Fe concentration and pH, whereas the presence of bacteria disables all such 6L-Fh precipitation and produces two-line Fh (2L-Fh)-like biotic coprecipitates. Given the nature of the differences between a-HFO and b-HFO and their synthesis condition dependences, several biotic precipitation mechanisms (template effect, near-cell environment effect, catalyzed nucleation and/or growth effect, and substrate-based coprecipitation) are ruled out. The prevailing present view of a template or heterogeneous nucleation barrier reduction effect, in particular, is shown not to be the cause of the large observed biotic effects on the resulting HFOs. The only proposed mechanism (relevant to Fh) that is consistent with all our observations is coprecipitation with and possible surface poisoning by ancillary bacteriagenic compounds. That bacterial cell wall functional groups are redox active and the characteristics of biotic (i.e., natural) HFOs compared to those of abiotic (i.e., synthetic) HFOs have several possible biogeochemical implications regarding Fe cycling, in the photic zones of water columns in particular.     

Adsorption of copper onto goethite in aqueous systems

The adsorption of copper(II) onto goethite was studied as a function of pH, total dissolved copper concentration, surface area of goethite, and ionic strength. The adsorption of copper was similar to that of other hydrolyzable metals. A tenfold increase in goethite surface area had a significant effect on the adsorption edge, but a tenfold increase in the ionic strength of the medium did not effect the adsorption edge. The distribution coefficients increase sharply with increase in pH and ranged from 10 to 60,000 ml/g over a range of two and half pH units, depending on the goethite surface area and copper concentration. A tenfold decrease in ionic strength as well as a tenfold increase in surface area of goethite did not have any effect on the magnitude of distribution coefficients. Distribution coefficients were used to calculate the number of protons released per mole of copper adsorbed during the adsorption process. The average number of protons released per mole of copper adsorbed was estimated to be 1.40 ± 0.10.Managed by Martin Marietta Energy System, Inc., for the U.S. Department of Energy under contract no. DE-AC05-840R21400.