ACS Publications. Most Trusted. Most Cited. Most Read
My Activity

Figure 1Loading Img

A Critical Review on the Multiple Roles of Manganese in Stabilizing and Destabilizing Soil Organic Matter

  • Hui Li
    Hui Li
    Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United States
    More by Hui Li
  • Fernanda Santos
    Fernanda Santos
    Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United States
  • Kristen Butler
    Kristen Butler
    Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United States
    Department of Earth and Planetary Sciences, College of Arts & Sciences, University of Tennessee, Knoxville, Tennessee 37996, United States
  • , and 
  • Elizabeth Herndon*
    Elizabeth Herndon
    Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United States
    Department of Earth and Planetary Sciences, College of Arts & Sciences, University of Tennessee, Knoxville, Tennessee 37996, United States
    *Phone: 865-341-0330; email: [email protected]
Cite this: Environ. Sci. Technol. 2021, 55, 18, 12136–12152
Publication Date (Web):September 1, 2021

Copyright © 2021 The Authors. Published by American Chemical Society. This publication is licensed under

CC-BY-NC-ND 4.0.
  • Open Access

Article Views





PDF (3 MB)
Supporting Info (3)»


Manganese (Mn) is a biologically important and redox-active metal that may exert a poorly recognized control on carbon (C) cycling in terrestrial ecosystems. Manganese influences ecosystem C dynamics by mediating biochemical pathways that include photosynthesis, serving as a reactive intermediate in the breakdown of organic molecules, and binding and/or oxidizing organic molecules through organo-mineral associations. However, the potential for Mn to influence ecosystem C storage remains unresolved. Although substantial research has demonstrated the ability of Fe- and Al-oxides to stabilize organic matter, there is a scarcity of similar information regarding Mn-oxides. Furthermore, Mn-mediated reactions regulate important litter decomposition pathways, but these processes are poorly constrained across diverse ecosystems. Here, we discuss the ecological roles of Mn in terrestrial environments and synthesize existing knowledge on the multiple pathways by which biogeochemical Mn and C cycling intersect. We demonstrate that Mn has a high potential to degrade organic molecules through abiotic and microbially mediated oxidation and to stabilize organic molecules, at least temporarily, through organo-mineral associations. We outline research priorities needed to advance understanding of Mn–C interactions, highlighting knowledge gaps that may address key uncertainties in soil C predictions.

This publication is licensed under

CC-BY-NC-ND 4.0.
  • cc licence
  • by licence
  • nc licence
  • nd licence

1. Introduction

Jump To

Soils are a dynamic global C reservoir that contain at least three times more C than the atmosphere and four times more C than terrestrial plants. (1,2) Even small changes in the stability and bioavailability of soil C can alter atmospheric CO2 concentrations and further influence global climate. Indeed, approximately ∼14% of anthropogenic CO2 released to the atmosphere over the past decade is attributed to land use changes that destabilized soil carbon. (3) Organic matter (OM) that is intimately associated with minerals constitutes an important soil C stock. (4,5) While an extensive literature exists on biotic and abiotic factors controlling inputs and outputs of soil C in terrestrial ecosystems, (6−8) there have been indications that C storage depends on interactions between climatic and geochemical factors. (9) Predicting the magnitude and direction of changes in global C cycling with climate change requires a fundamental understanding of how intersecting climatic, biological, and geochemical factors influence soil C storage.
Manganese is a highly reactive soil constituent that is known to interact with organic compounds but remains one of the least understood factors influencing soil C storage. This critical review provides an ecosystem-level examination of interactions between Mn and OM that regulate soil C storage. This topic, despite growing interest and increasingly recognized relevance to C cycling, has never been reviewed. Remucal and Ginder-Vogel (10) provide a comprehensive assessment of how a broad suite of organic contaminants react with Mn-oxides but focus on contaminant transformation rather than Mn–C interactions in an ecosystem context. Other recent reviews have focused on properties of Mn-oxides, (11) mechanisms of Mn-oxide reduction by inorganic species, (12) microbial transformation of Mn in marine systems, (13) and mineral–organic associations that mostly exclude Mn-oxides. (4) Additionally, Berg et al. (2015) (14) discuss the role of enzymatically produced Mn(III) in driving broad trends in litter decomposability. Our aim is to synthesize existing knowledge and identify critical gaps in understanding of coupled Mn–C cycling. Our ultimate objective is to evaluate in what circumstances Mn serves to increase C storage in soils through mineral protection or decrease C storage through oxidative degradation of organic compounds. To address this unknown, we review mechanisms underlying Mn–C interactions that stabilize or destabilize organic compounds and place these mechanisms within the broader context of terrestrial environments.

2. Manganese in Terrestrial Ecosystems: Occurrence and Characterization

Jump To

2.1. Manganese Geochemistry in Soils

Manganese is a redox-sensitive element that comprises about 0.1% of weight of crustal rocks (15) and is commonly found in air, soils and sediments, and surface and groundwater. (16−18) Mn(II) that is released during weathering of bedrock minerals is readily oxidized to Mn(III/IV) to form more than 30 known Mn (hydr)oxide (i.e., Mn-oxides) minerals with more than 15 phases known in soils (e.g., birnessite, manganite, and hausmannite; Supporting Information (SI) Table S1). (19,20) Manganese oxides are ubiquitous in soils and sediments as fine-grained aggregates, veins, nodules and concretions, crusts, dendrites, and coatings on rocks or other mineral particles. (19) Due to their high oxidation potentials and relatively large surface areas, Mn-oxides exert strong controls over the distribution, transport, and transformation of other elements, far out of proportion to their natural abundance. Here, we briefly introduce Mn redox cycling and the formation, properties, and transformation of Mn-oxide minerals and defer to existing reviews (10,13,19,21−23) for more detailed processes.
The majority of naturally present Mn-oxides are formed during biotic Mn(II) oxidation to a poorly crystalline phyllomanganate phase that subsequently transforms to more ordered phases (SI Figure S1). (21,24−26) Mn(II) is oxidized by a wide range of phylogenetically diverse microorganisms (bacteria, fungi, algae, and diatoms) using multiple enzymatic and nonenzymatic pathways. (11,13,21,27−29) Various microorganisms use multicopper oxidases and/or animal haem peroxidases (AHPs) to directly mediate a one-electron transfer from Mn(II) to generate Mn(III), which may be followed by disproportionation to Mn(II) and Mn(IV) (13,21) or a second electron transfer reaction to form Mn(IV). (30) Biotic Mn(II) oxidation also occurs through interaction with reactive oxygen species (ROS) produced by bacteria (e.g., Roseobacter (27,31)), Ascomycete fungi, (32) and phototrophic algae and diatoms. (28) The physiological function of Mn(II) oxidation remains unclear since most known organisms do not obtain energy from the reaction, and Mn(II) oxidation has only been demonstrated as a viable energy-yielding metabolism in select bacteria. (33,34)
Abiotic Mn(II) oxidation is another important and well-studied pathway for Mn-oxide formation. Homogeneous oxidation of aqueous Mn(II) by dissolved O2 is thermodynamically unfavorable (at pH < 8) and kinetically hindered due to high activation energy of the reaction, (12,18,35) yet there are many physical and chemical factors that drive abiotic formation and transformation of Mn-oxides. Mn(II) can be oxidized to Mn(III) by ROS (e.g., O2•–, 1O2, OH) produced through abiotic pathways including Fe(II) oxidation, (36) nitrate photolysis, (37,38) illumination of humic substances, (39) and illumination of metal oxides in desert varnish. (40) Further Mn(III) oxidation by ROS and/or disproportionation results in rapid Mn(IV)-oxide formation. (37,38) Mineral surfaces, such as Fe-oxides and Mn-oxides, also catalyze Mn(II) oxidation via interfacial catalysis and/or electrochemical reactions (41−43) at rates equivalent to or faster than biological oxidation. (44,45) Products of Mn(II) oxidation and Mn-oxide phase transformations are highly dependent on pH conditions (e.g., SI Figure S1) and mineral surface properties (e.g., semiconductivity, particle size, ions). (35,42,46−48) For instance, the presence of transition metals (e.g., Co and Ni), (49) high concentrations of uranyl, (50) and thallium (51) have all been shown to mediate changes in Mn-oxide crystallinity or phase transformation.
Microbial Mn(IV) reduction, well-understood as a primary driver of Mn-oxide dissolution under reducing conditions, is attributed primarily to bacteria with lesser contributions from fungi and archaea. (13) Mn(IV)-reducing bacteria respire Mn(IV) using H2 and various organic compounds as electron donors, but this process may be hindered in the presence of other electron acceptors such as oxygen, nitrate, and sulfur compounds. (34) Microbial Mn(IV) reduction initiates via a one-electron transfer step to form soluble Mn(III) as intermediate product. (12,52) Mn(III,IV)-oxides are also abiotically reduced by a wide range of organic (53,54) and inorganic compounds, (12) for example, Fe(II) and sulfide. Mn reduction rates vary with environmental factors including soil temperature, water potential, pH, and redox conditions. (55)

2.2. Ecological Function of Mn as a Micronutrient and in Relation to Carbon Storage

Manganese plays essential yet complex roles in influencing soil C storage. Specific mechanisms influencing C dynamics include Mn’s key function as a biotic and abiotic oxidizer of soil organic matter (SOM), as discussed in more detail in following sections. In this subsection, we introduce its broader roles as a micronutrient and in microbial decomposition.
Manganese is a critical micronutrient for plants and can impair growth when present in either limiting or excess quantities. In acidic soils where Mn is readily soluble, cycling through vegetation (i.e., uptake and litterfall) results in Mn accumulation in surface soils after a few decades. (56,57) Dissolved Mn(II) in soil solution is moved through roots and transported primarily to leaves where it is utilized in the water-splitting complex of photosystem II and as a cofactor for superoxide dismutase, an enzyme that detoxifies byproducts of photosynthesis. (58) Manganese accumulates in leaves as aqueous or organic-bound Mn(II) (59,60) and is not translocated to other plant tissues or reabsorbed during senescence; consequently, Mn is largely returned to soil each year in throughfall or litterfall. (61) Mn(II) species in litter are oxidized to insoluble Mn(III/IV)-oxides during decomposition, facilitating their accumulation in surface soils. (59,62,63) Many plants accumulate Mn in linear proportion to the quantity of soluble and exchangeable Mn(II) in soil. (64) As such, foliar Mn concentrations often exceed nutritional requirements, resulting in oxidative stress and impaired photosynthesis in sensitive plant species (e.g., sugar maples). (58,65−67) At the ecosystem-scale, Mn toxicity can lead to declines in certain plant types where soil Mn is highly phytoavailable. For example, increases in Mn(II) solubility driven by chronic nitrogen deposition led to the decline of forbs in a temperate steppe ecosystem. (68) The resulting shift to an exclusively grassland system was attributed to the greater sensitivity of forbs to reduced photosynthetic rates and growth driven by high Mn(II) concentrations. Decreases in plant biomass and shifts in plant communities driven by Mn bioavailability may in turn affect the quantity and composition of C inputs to soil.
During SOM decomposition, facultative anaerobic microorganisms can utilize Mn(IV) as a terminal electron acceptor via dissimilatory Mn(IV) reduction, particularly when more thermodynamically favorable dissolved O2 and nitrate are absent or limited. (22,69) Microorganisms couple Mn(IV) reduction to their metabolic oxidation of organic compounds (e.g., acetate, sugars, amino acids, long chain fatty acids, and aromatics) to release CO2. (70,71) Dissimilatory Mn(IV) reduction depends on pH and redox potential and predominates where anoxic or suboxic conditions develop, such as pores within soil microaggregates and soil horizons that experience large fluctuations in moisture and/or are poorly drained. Specifically, Mn(IV)-oxides are reduced by microorganisms during periods of anoxia when O2 is depleted but regenerated during subsequent oxic periods. High concentrations of OM also promote reducing conditions that favor Mn reduction. As one example, applications of biodegradable OM to crop soils over 19 years coupled to high soil moisture and temperature promoted anaerobic respiration that mobilized high concentrations of Mn in subsoils. (72) Hence, Mn is expected to serve as an important terminal electron acceptor in soils, especially in organic-rich soils with absent or low levels of O2 and nitrate that are strongly affected by fluctuating water tables. Reducing conditions that drive Mn-reduction can also increase dissolved organic C (DOC) mobilization in soils, (73) potentially increasing OM bioavailability to microorganisms.
Manganese may serve other ecological functions that remain poorly recognized. For example, Macrotermitinae are a subfamily of termites that can depolymerize lignin in their guts (74) and cultivate specialized Basidiomycete fungi in their colonies to aid in digestion of lignocellulosic material. (75) Certain Basidiomycetes are highly effective wood degraders due in part to their production of Mn peroxidase enzymes that act as biocatalysts to degrade lignin. (76) Cultivated fungal combs can concentrate Mn to levels nearly 103× the surrounding soil, and termites that ingest these combs contain 100× more Mn than other insects and store unique Mn nodules in their abdomens. (77) Given that fungal-farming termites are important ecosystem engineers in many arid environments, Mn may be an unrecognized supporter of these ecosystems through its ability to enhance lignin degradation.

2.3. Mn Characterization Techniques and Limitations

Studies on Mn in soils are limited by its low abundance, monoisotopic nature, and the poor crystallinity of a diverse array of Mn-oxides. Poorly crystalline minerals are difficult to identify and characterize due to limitations of conventional techniques. Techniques used to study Fe-oxides and Fe–C interactions, for example, Fe stable isotope labeling and fractionation and Mössbauer spectroscopy, cannot be applied to Mn-oxides. Furthermore, commonly used chemical extractions (e.g., dithionite) are nonspecific and coextract Mn-oxides with Fe-oxides. Even chemicals that selectively dissolve Mn-oxides (hydroxylamine hydrochloride) may extract small quantities of Fe-oxide (<5%) that are equal in concentration to extracted Mn. (78)
Despite these challenges, a suite of techniques applied to natural and synthetic Mn-oxides have provided insight into their formation and properties. Synchrotron-source X-ray absorption spectroscopy (XAS) remains one of the key techniques for characterizing Mn within complex environmental samples, (21) including both minerals (25,79,80) and organic materials (59,62) in bulk or at the micrometer-scale, by probing its oxidation state and local bonding environment. (81) X-ray diffraction (XRD) has also been widely used to characterize synthetic, biogenic, and natural Mn-oxide structures, (11,19) but application of even synchrotron-source XRD is challenging for natural samples that are low in Mn-oxide abundance or lack long-range crystalline order. (25,80) Where XRD is limiting, Fourier transform infrared (FTIR) spectroscopy can probe short-range order by measuring Mn–O and O–H bond vibrations, yielding information on vacancies and proportions of distinct phases within mixtures (e.g., triclinic and hexagonal birnessite). (80,82) Additional phase identification has been performed using nondestructive Raman spectroscopy, although application of this technique has been limited by low Raman activity and experimental artifacts. (83,84) X-ray photon spectroscopy (XPS) provides key information on Mn oxidation states at mineral surfaces that mediate geochemical reactions but may differ from bulk composition. (85) Finally, new insights into fractionation of stable oxygen isotopes (δ18O) during biogenic Mn(II) oxidation indicate the potential for δ18OMnOx to record oxygen sources and bio-oxidation pathways. (86,87) These techniques, particularly in combination, can provide new insights into coupled Mn–C interactions in natural systems.

3. Mn-Promoted Protection of Organic Compounds

Jump To

Organic compounds are known to directly interact with mineral surfaces through various sorption reactions (88) or can be stabilized through coprecipitation and aggregate formation. The mineral–OM association is one essential mechanism in stabilizing and protecting SOM against microbial metabolism. (4,89) Iron oxides are common soil minerals that have been extensively studied for their abilities to bind OM through mechanisms such as anion exchange, ligand exchange/surface complexation, hydrophobic interaction, entropic effects, hydrogen bonding, and cation bridging. (90) Mn-oxides possess many similar properties to Fe-oxides but are less well studied for their interactions with OM. However, Mn-oxides, particularly poorly crystalline phases with relatively large specific surface area (SSA), have unusually high adsorption capacities and scavenging capabilities for organic (91,92) and inorganic compounds (e.g., heavy metals), (20) thereby potentially protecting OM against decay. This section explores mechanisms for OM stabilization by minerals including Mn-oxides and discusses current evidence for these associations in terrestrial (91,93) and engineered systems. (94) A summary of Mn–C interactions, both stabilizing reactions discussed here and destabilizing reactions reviewed in Section 4, is presented in Figure 1. Our aim is to present an overview of these processes, while more detailed reaction mechanisms are reviewed elsewhere. (10)

Figure 1

Figure 1. Conceptual diagram summarizing stabilizing and destabilizing interactions between Mn(III,IV)-oxides and organic compounds. Organic compounds are represented as unspecified organic matter (OM) or by specific functional groups and/or compounds. The processes shown here serve as examples for each category and are streamlined representations of detailed mechanisms. Stabilization reactions include outersphere interactions such as cation bridging (103) and electrostatic interactions, inner-sphere ligand exchange reactions that release hydroxyl ions, (94,106) physical trapping between Mn-oxide layers, (94) coprecipitation within the oxide structure, and polymerization of degraded phenolic structures. (116) Destabilization reactions include reductive dissolution of Mn-oxides through abiotic interactions with reducing compounds (e.g., catechol degradation (116)) or microbially mediated dissimilatory Mn reduction (22), fragmentation of phenolic radicals, (116) ligand-promoted extraction of Mn(III) (e.g., phosphonoformic acid), (54) and Mn(III)-promoted degradation of organic chelators through internal electron transfer (e.g., NTMP degradation (146)). Degradation reactions often release byproducts such as carbon dioxide or phosphate through cleavage of functional groups.

3.1. Mechanisms for Organic Matter Protection

Soil OM is protected from microbial degradation in part through mineral associations. (4) Adsorption of organic compounds to mineral surfaces involves physical forces such as hydrophobic partitioning, electrostatic outer-sphere complexation, and cation bridging, and chemical interactions including inner-sphere complexation (ligand exchange). (95) Organic matter can also be protected when it is incorporated into (coprecipitation) or encapsulated by (physical trapping) precipitating minerals. Polymerization of small organic compounds into more complex structures may increase chemical recalcitrance.

Hydrophobic Partitioning

Hydrophobic compounds partition from aqueous solution onto Mn-oxides due to incompatibility of nonpolar compounds with water. (96) Organic compounds with hydrophobic moieties (e.g., −OCH3, −CH3, −CF3, −CN) have shown a higher tendency to precipitate on Mn-oxides surface than compounds with hydrophilic moieties. (96) As one example, during adsorption of several phenylarsonic acid analogues on birnessite, molecules with higher hydrophobicity showed higher adsorption rates. (97)

Electrostatic Interaction

pH-dependent electrostatic interaction is one of the most important mechanisms driving OM adsorption to soil minerals. (98,99) Minerals and OM are net positively charged at lower pH than their isoelectric points (point of zero charge = PZC) due to surface protonation but net negatively charged at higher pH due to deprotonation. (100) Most Mn-oxides are net negatively charged on the surface at neutral pH (20) (SI Table S1) and have high potential to sorb positively charged ions and functional groups. For example, protein adsorption on soil minerals surface normally initiates through electrostatic interaction, during which positively charged functional groups of proteins complex with negatively charged mineral surfaces. (100,101) However, this weak interaction is generally reversible. (101) Similar to Mn-oxides, natural organic matter (NOM) is generally net negatively charged at circumneutral pH, (95) limiting electrostatic interactions between OM and Mn-oxides under many common soil conditions. However, electrostatic interactions between Mn-oxides and organic compounds increase with decreasing pH as organic compounds protonate and associate with Mn-oxide surfaces. (97,99)

Cation Bridging

Polyvalent cations can balance negative charges on mineral surfaces and OM functional groups, thus acting as a bridge to connect them. (102) For example, adsorption of Mn2+ ions to Mn-oxide surfaces can partially neutralize negative surface charges and facilitate adsorption of complex humic-like compounds. (103) Dissolved Ca2+ has similarly been shown to promote adsorption of river-isolated NOM to poorly crystalline and crystalline Mn-oxides (lithiophorite, birnessite, and cryptomelane). (96) This mechanism has been frequently observed for clay minerals, which remain negatively charged over a wide pH range. For example, metal cation-saturated montmorillonites can adsorb up to 5× more DNA than pure montmorillonites, (102) and divalent cations (e.g., Ca2+ and Mg2+) are generally more efficient in promoting adsorption than monovalent cations (e.g., Na+). (104)

Ligand Exchange

Ligand exchange is a strong adsorption mechanism that occurs between hydroxyl (−OH) functional groups on mineral surfaces and carboxyl or phenolic −OH groups of OM. (90,105) Significant −OH release during NOM adsorption to birnessite serves as an indicator of this ligand exchange reaction. (106) It is likely that carboxylate groups form inner-sphere complexes with birnessite via ligand exchange with −OH on the Mn-oxide surface. (94,106) Ligand exchange is particularly favored for NOM with high molecular weight, acidity, and aromaticity. (90,106) Unlike weaker electrostatic and cation bridging reactions, specific adsorption through ligand exchange is not readily reversible.


Under aeration, Mn(II) can be rapidly oxidized to Mn(III/IV)-oxides that coprecipitate with dissolved OM. (107) Coprecipitation with OM, primarily studied in Fe-oxides and Al-oxides, leads to quantities and compositions of stabilized OM that differ from adsorption. (108−110) Coprecipitation of SOM with Fe-oxides has resulted in higher C stabilization (108) and higher enrichment of aromatic components (110) than surface complexation (adsorption). Manganese widely coexists with Fe in soils, though at lower abundance, and precipitates under similar conditions. It is likely that SOM coprecipitates with Mn-oxides to generate C stabilization similar to Fe-oxides.

Physical Trapping

Mn-oxides can physically protect OM by coating organic structures or serving as cementing agents that stabilize OM in microaggregates. (111) Mn(II)-oxidizing bacteria (e.g., Leptothrix discophora and Pseudomonas putida), fungi (e.g., Acremonium and Ascomycete) and phototrophic organisms (e.g., algae and diatoms) are often encrusted in the Mn(III,IV)-oxides generated during microbial Mn(II) oxidation, (21,28,32,112,113) which can trap cellular components or whole cells. Dissolved organic compounds that bind on Mn-oxide surfaces can also be trapped when a new layer of Mn-oxide precipitates. (94)


Polymerization reactions stabilize soil organic carbon either physically, ascribed to increased size and complexity of micro/macro-aggregates, or chemically, via increasing the energy threshold to be overcome in microbial respiration. (114) Mn(IV)-oxides are Lewis acids that serve as efficient oxidants that can degrade organic compounds into smaller molecules, as discussed in the next section, or accelerate oxidative polymerization of polyphenols under common soil pH conditions (e.g., 4–8). The ability of Mn-oxides to drive polyphenol polymerization, sometimes referred to as “browning”, far exceeds similar effects of Fe-, Al-, or Si-oxides. (115) This accelerating effect is more efficient at near neutral pH than acidic conditions. Birnessite rapidly oxidizes hydroquinone to 1,4-benzoquinone at pH 7, which then undergoes dimerization and/or ring cleavage reactions that ultimately polymerize degradation products. (116) Associated decarboxylation reactions produce CO2, which can be released as a gas or react with aqueous Mn(II) to form rhodochrosite (MnCO3) at neutral pH. (116) In one case, NOM adsorbed to Mn-oxide largely via Mn-carboxylate bonds; subsequently, phenol groups within the bound OM were oxidatively transformed to phenoxy radicals that further polymerized (e.g., converted to quinones or produced dimers or polymeric products) in the interlayers of Mn-oxides. (94)

3.2. Evidence for Organic Matter Protection in Natural Systems

While many studies have investigated how individual organic compounds adsorb to and react with Mn-oxides, few have examined these interactions in natural or engineered systems with environmentally formed minerals and/or complex mixtures of organic compounds. In two studies, DOM leached from forest organic soils was reacted with synthetic goethite and Mn-oxides (birnessite and hydrous Mn-oxide). (92,106) Birnessite adsorbed low amounts of C relative to goethite but more effectively oxidized and transformed OM. (106) In contrast, hydrous Mn-oxide had a greater capacity for OM sorption than goethite, most likely due to its high surface area. However, OM bound to hydrous Mn-oxide was also more readily desorbed, possibly because goethite formed carboxylic-metal bonds with OM while bond with hydrous Mn-oxide were weaker. (92)
Interactions between Mn-oxides and OM have been reported for wastewater treatment systems where Mn-oxides are used as sorbents in the treatment process. Substantial OM was associated with birnessite-coated sand obtained from a water treatment filter bed. (94) Organic compounds that adsorbed to birnessite surfaces were physically trapped under progressive layers of freshly precipitated birnessite. Trapped OM was more aromatic than the initial OM, indicating the potential for Mn-mediated polymerization of OM within the oxide layers. In another study, (96) NOM was reacted with both synthetic birnessite and natural Mn-oxides (mixture of lithiophorite, birnessite, and cryptomelane) used for drinking water treatment. Organic matter adsorption increased as pH decreased and was facilitated by hydrophobic partitioning. High molecular weight molecules with high aromaticity preferentially adsorbed to the minerals and were partially oxidized to low molecular weight compounds.
There is also evidence that Mn-oxides serve as meaningful reservoirs for OM in natural environments that include brackish pond water, ferromanganese cave deposits, and soils. (91,93) Estes et al. (2017) (91) examined how Mn-associated OM in cave deposits and pond water compared to Mn–C associations formed in cultures of Mn-oxidizing terrestrial fungi and marine bacteria. In pond water and cave deposits, Mn-oxides trapped heterogeneous mixtures of OM that reflected the natural complexity of OM in those systems. In microbial cultures, Mn-oxides hosted relatively homogeneous OM enriched in proteinaceous and/or lipopolysaccharide compounds (amide, carboxylic, and o-alkyl functional groups), likely derived from the microbial components used in Mn-oxide formation.
In soils, the contribution of Mn-oxides to OM preservation may be historically underestimated by bulk techniques that fail to differentiate between Mn- and Fe-oxides; however, spatial associations between Fe, Mn, and OM are important to consider. (93) Within a soil ferromanganese nodule, Fe- and Mn-oxides were shown to form distinct accumulation zones following cycles of mobilization and reprecipitation during alternating wetting and drying conditions. (93) These oxide accumulations preserve OM by acting as cements that limit microbial access to small pores. Organic C was highly enriched where Mn-oxides accumulated but relatively depleted in Fe-rich zones, indicating an important role for Mn-oxides in preserving OM in these nodules.

4. Mn-Promoted Degradation of Organic Compounds

Jump To

4.1. Oxidation of Organic Compounds by Mn-Oxides

Mn-oxides are among the strongest natural oxidants in soils (95) and are capable of oxidizing OM (53,106,117) and inorganic compounds. (20,118) Mn-oxides can decompose OM via oxidative reactions into smaller compounds that can ultimately be transformed to CO2. (119) Mn-promoted OM oxidation also directly produces CO2 through decarboxylation reactions. (116) Although extensive research has addressed oxidation of organic contaminants by Mn-oxides, (10) particularly as a remediation strategy, the role of Mn-oxides in degrading soil OM remains unresolved. In this section, we discuss Mn-oxide dissolution by organic compounds via proton-promoted (organic acids) and ligand-assisted (e.g., oxalate, citrate) mechanisms as well as the oxidation and degradation of organic compounds (53) by reductive dissolution (e.g., catechol, phosphonoformic acid, oxalic acid) (Figure 1). (54,120)

Proton-Promoted Reaction

Decreases in pH accelerate mineral dissolution and transformation. (46) Organic acids are a potential source of protons that bind to oxygen atoms on the mineral surface, weaken their bonds with metal cations, (95) and facilitate Mn-oxide dissolution. For example, manganite (MnIIIOOH) dissolves via proton-promoted disproportionation at pH < 6 to produce MnIVO2 solids and dissolved Mn(II). (121) Similarly, disproportion of hausmannite (Mn3O4) is initiated by proton-promoted dissolution at pH < 7 and accelerates with decreasing pH. (46)

Ligand-Assisted Reaction

Ligands that complex onto structural Mn centers via exchange for hydroxyl functional groups can change mineral surface coordination, polarize and weaken the Mn–OH bonds, and cause detachment of the Mn cation into solution. (95,105) MnOOH dissolution by citrate, pyrophosphate (for β-MnOOH), (105) and phosphonoformic acid (for γ-MnOOH) (54) initiates via ligand-assisted dissolution reactions that form aqueous Mn(III)-complexes.

Redox Reaction

Mn(IV) and Mn(III) in Mn-oxides are both capable of accepting electrons from organic molecules and being reduced to lower valence states, for example, to Mn(III) and Mn(II). The reducing organic compound is concurrently oxidized to produce various degradation products. For example, in contrast to manganite, dissolution of birnessite by phosphonoformic acid (H2PO3COOH) occurs via reductive dissolution to generate aqueous Mn(II), orthophosphate, and CO2. (54) Birnessite efficiently oxidizes catechol into benzoquinone and eventually to CO2, forming three moles of Mn(II) for each mole of CO2 released. (122) Oxidative transformation of organic compounds by Mn-oxides is also evident for complex organic mixtures. Dissolved OM from forest floor leachate reacted with Mn-oxide to release aqueous Mn(II), acetic acid, and formic acid as major products. (106)
Reductive dissolution is the predominant mineral dissolution mechanism induced by most organic reactants; (54) however, many SOM compounds interact with Mn-oxides via a combination of the reaction mechanisms discussed above. For example, oxalic acid, phosphonoformic acid, and citric acid strongly adsorb to several Mn-oxides. (54,120) These organic compounds complex with Mn-oxides through ligand exchange, initiate ligand-assisted dissolution, and then undergo intramolecular electron transfer in the solubilized Mn(III)–organic complex to generate aqueous Mn(II) and oxidized organic compounds.
Plant and microbial siderophores, organic chelating compounds produced by soil organisms to acquire Fe, also effectively dissolve and are degraded by Mn-oxides through a combination of dissolution mechanisms. Siderophores bind structural Mn(III) atoms to functional groups such as hydroxyamate, catecholamide, hydroxycarboxylate, and/or catecholate. (123,124) At circumneutral to high pH, siderophores extract Mn(III) and drive ligand-promoted dissolution. (125−127) At circumneutral to low pH, siderophores and their subsequent oxidation products reduce Mn(III) to Mn(II), either at the mineral surface or in solution following Mn(III) extraction. (124−127) In Mn(II)-bearing oxides such as hausmannite, depletion of Mn(III) ultimately destabilizes the mineral structure and induces additional release of Mn(II) ions into solution. (127) These compounds may act in concert with low molecular weight organic compounds to dissolve Mn-oxides in soils, (128) with the added significance that Mn-oxides may compete with Fe-oxides and act as an important soil sink for siderophores.

4.2. Oxidation of Organic Compounds by Mn(II)/Mn(III) Cycling

4.2.1. Fungal-Mediated Mn Cycling and Litter Decomposition

Manganese is a critical component of the microbial decomposition of lignin compounds and has been identified as a key factor influencing litter decomposition rates. Much of what is known about the mechanisms for Mn-promoted lignin degradation is based on pure culture studies of lignin-degrading fungi and their enzymes. (76) However, correlational relationships at the ecosystem-scale suggest positive relationships between foliar Mn and both litter mass loss (14,129) and the activities of lignin-degrading enzymes. (130) Similarly, exchangeable soil Mn(II) is negatively correlated with soil C storage across multiple biomes. (131−133)
During litter decomposition, a select group of saprotrophic fungi (Agaricomycetes within the phylum Basidiomycota) release the extracellular enzyme Mn peroxidase (MnP) which degrades lignin by oxidizing its phenolic bonds. (134) Other ligninolytic enzymes (e.g., lignin peroxidase, laccase) are less efficient lignin degraders because they only attack nonphenolic bonds. Specifically, MnP oxidizes aqueous Mn(II) to chelated-Mn(III), which then serves as a diffusible reactant that oxidizes lignin. (135) As the microbial oxidation of lignin proceeds, (hemi)cellulose bound within the lignocellulose matrix becomes readily accessible to soil microbes. Given that lignin is the next most abundant natural polymer on Earth after (hemi)cellulose, understanding how Mn(II) and Mn(III) cycling in soils contributes to litter degradation (130,136) is essential to determine its role in the global C cycle. It is important to note that MnP can oxidize more than lignin. For example, certain basidiomycetes can oxidize a suite of polycyclic aromatic hydrocarbons. (137−139) Chelated-Mn(III) can also oxidize unsaturated fatty acids to lipid radicals (140,141) which further oxidize and cleave nonphenolic lignin structures. (142)
These observations suggest that Mn availability is a key regulator of fungal-mediated litter decomposition and plays a relevant role in stimulating soil CO2 fluxes and influencing soil C storage. To assess net impacts of Mn enrichment on ecosystem C storage, we compiled available field and laboratory data reporting ecosystem response to simulated Mn addition (Figure 2; SI Table S2). Despite experimental differences in substrate composition, Mn addition, and incubation time, these studies demonstrate that Mn addition stimulates production of ligninolytic enzymes including MnP (+20 ± 6% change relative to control), depletes lignin in decomposing litter (−14 ± 5%), and increases C loss as both DOC (+17 ± 12%) and CO2 (+14 ± 3%). As one example, litter and organic horizons that received Mn additions were reported to have 10–35% higher rates of soil-respired CO2 effluxes (143,144) and produce higher DOC concentrations (145) than control treatments. The mass of leaf litter that remained following decomposition also generally decreased in response to Mn addition over the incubation time (−5 ± 6%), but this effect was less clear across treatments.

Figure 2

Figure 2. (top) Mn-mediated enzymatic oxidation of lignin by Agaricomycete fungi (phylum Basidiomycota). Fungi produce manganese peroxidase (MnP) that reacts with hydrogen peroxide (H2O2) to convert Mn2+ into reactive Mn3+, which in turn is stabilized by small organic molecules (i.e., organic acids). The attack of phenolic lignin structures by chelated Mn3+ leads to the production of CO2 and DOC. (bottom) Responses (in % difference relative to control) of ligninolytic enzyme activity, leaf litter mass, lignin content, CO2 fluxes (or respiration), and dissolved organic carbon to Mn additions in studies that simulated Mn addition to litter and/or soils (see data in SI Table S2). Data in the violin plots include the median (white dot), interquartile range (horizontal black bar), lower and upper adjacent values (black whiskers), outliers (black crosses), and probability density (colored area).

4.2.2. Organic–Mn(III) Complexes in Soil Solution

Organic–Mn(III) complexes formed through abiotic processes have also been implicated as a driver of OM degradation. (43,146) Chelated Mn(III) can form through either oxidation of aqueous Mn(II), extraction of solid-phase Mn(III) by ligands, or surface-promoted oxidation of sorbed Mn(II). For the latter, minerals such as goethite facilitate the oxidation of aqueous Mn(II) to reactive Mn(III) species that subsequently oxidize organic molecules to regenerate Mn(II). (43)
Aqueous Mn(II) ions form complexes with a variety of organic molecules that accelerate their oxidation to Mn(III). For example, phosphonic acids such as nitrilostris(methylene-phosphonic acid) (NTMP) form Mn(II)–NTMP complexes that oxidize to Mn(III)–NTMP in air. (146) Mn(III) subsequently extracts an electron from the compound and oxidizes NTMP to form degradation products. A diverse suite of siderophores, including desferrioxamine B (DFOB), rhizoferrin, and protochelin, similarly bind aqueous Mn(II) and facilitate air oxidation to Mn(III)–siderophore complexes. (123,147,148) These Mn(III) complexes are stable at high pH (∼7–11) but undergo internal electron transfer to generate aqueous Mn(II) and oxidized degradation products at pH < 7. (147) Aqueous organic-Mn(III) complexes can also form when ligands (e.g., phosphonates, siderophores) extract Mn(III) from Mn-oxides. (149) Siderophores drive ligand-promoted dissolution of hausmannite, (127) manganite, (125) Mn(III/IV)-oxides, (126) and δ-MnO2 (128) under circumneutral to alkaline pH by either directly extracting Mn(III) from Mn(III)-bearing minerals or reducing Mn(IV) atoms to Mn(III) prior to extraction.
Although the occurrence of Mn(III) species at redox interfaces is increasingly recognized in aqueous and terrestrial systems, (150,151) the environmental interactions of these Mn(III) species are relatively unexplored. Mn(III) can form complexes with organic ligands under oxic and anoxic conditions, (123,148,152) and evidence regarding their subsequent reactivity is newly emerging. For example, Mn(III) catalyzes N-dealkylation of atrazine and can also oxidatively degrade 4-chlorophenol through one electron transfer. (153) Mn(III) species formed by Mn(II) oxidation at redox interfaces also oxidize and depolymerize OM in soils. (154) Indeed, oxidative capability was shown to be strongly correlated with Mn(III) abundance in the litter layer of a temperate deciduous forest, from which Jones et al. (2020) (155) contend that Mn(III)-mediated reactions may dominant oxidative processes in litter and organic soils. Mn(III)–complexes can also oxidize nitrite to nitrate under anoxic conditions, (156) indicating an important role for Mn in driving oxidation of both organic and inorganic species.

5. Intrinsic and Environmental Factors Influencing Mn-Oxide–C Interactions

Jump To

This section discusses the roles of intrinsic characteristics of Mn-oxides and SOM and external environmental conditions on regulating Mn-oxide–C interactions.

5.1. Characteristics of Mn-Oxides That Influence Organo–Mineral Interactions

Biogenic and abiotic Mn-oxides that form in soils (157) transform to more stable and oxidized crystalline phases over time (SI Figure S1). (157,158) The properties of these Mn-oxide minerals are important for regulating reactivity with organic compounds. For example, the oxidation state, SSA, redox potential, and structure of Mn-oxides determine OM adsorption capacity and oxidation efficiency. (12,54) Here, we discuss how characteristics of Mn-oxides control their stability and reactivity. (157)
Specific surface area varies from <10 to >200 m2 g–1 among Mn-oxides that differ in structure, particle sizes, and crystallinity (SI Table S1). (20,159,160) In general, Mn-oxides with smaller particle sizes and poor crystallinity exhibit higher SSA and more exposed reactive surface sites. Biogenic Mn-oxides that consist of phyllomanganates with dominant structural Mn(IV) generally have higher SSA than more crystalline synthetic Mn-oxides. (20,160,161) High SSA and disordered crystalline structure generally confer higher reactivity, as has been demonstrated with poorly crystalline Fe-minerals that are more susceptive to reductive dissolution and subsequent mineralization of associated OM than crystalline Fe-oxide phases. (162,163)
Different Mn-oxides also possess different types and abundances of reactive surface sites, for example, planar octahedral vacancy sites and edge sites, (19,26,164) which provide different levels of binding energies. Edge sites contain unsaturated singly and doubly coordinated oxygen groups that confer a large fraction of the total surface charge and have high ion adsorption capacity. (26) In low-vacancy (<10%) δ-MnO2, edge sites were shown to contribute ≥50% of the total reactive sites for adsorbing Pb(II). (165) Edge sites also dominate electron transfer reactions for Mn-oxides such as birnessite. (166) In birnessite, edges serve as main adsorption sites for ligands such as fulvic acid (164) and for oxyanions such as As(III) and As(V) while vacancies have higher affinity for metal ions (e.g., Mn(II), Mn(III), Zn2+). (166)
Oxidation state largely determines the forms of Mn present and their environmental interactions. Mn(III) and Mn(IV) precipitate in Mn(III,IV)-oxides, which have among the highest standard reduction potentials of natural minerals (e.g., E° = 1.29 V for the synthetic vernadite-analog MnO2) and a higher capacity to oxidize organic compounds than MnIIIOOH. (54,95) Birnessite, the most dominant class of MnO2 minerals in soils, oxidizes a wide range of organic compounds. (54) Within a single MnO2 phase (e.g., birnessite or vernadite), higher Mn(IV) content generally results in stronger capacity to oxidize reduced species (e.g., As(III) and Cr(III)) (167,168) and organic compounds (e.g., hydroquinone). (169) The relative oxidizing activity between Mn(III) and Mn(IV) within Mn-oxides may vary due to multiple mechanisms. For example, Mn(III) sites are less effective at oxidizing As(III) due to their lower affinity for As(III) adsorption onto δ-MnO2 (167) and Cr(III) due to surface site occupation by Mn(III) on vernadite. (168) Although Mn(IV) sites confer higher oxidation capacity, Mn(III) undergoes more rapid ligand exchange. Surface functional groups OH and H2O easily dissociate from relatively labile Mn(III) due to Jahn–Teller distortion but barely from Mn(IV). (12) Mn(III) in Mn-oxides can accept one electron from OM or inorganic reactants and be reduced to Mn(II). Mn(II) is present in certain Mn-oxides (e.g., hausmannite) but is highly soluble and often found as an aqueous ion. (120) The presence of Mn(II) in solution improves OM adsorption to mineral surfaces via cation bridging. Mn(II) also complexes with humic acid to promote OM aggregation and precipitation, although this effect is more strongly induced by Mn(III). (43,152)
The structure, phase transformation, and morphology of Mn-oxides also regulate their oxidizing reactivity. Biogenic Mn-oxides typically undergo transformation from hexagonal to triclinic birnessite which decreases their oxidizing capability, as observed for Cr(III) oxidation in the presence of OM and light that drove photoinduced regeneration of hexagonal birnessite. (170) Rates that various MnO2 structures were able to oxidize bisphenol A decreased in the order δ-MnO2 > α-MnO2≈ γ-MnO2 > λ-MnO2 > β-MnO2. (171) Similarly, the rate that α-MnO2 oxidized the anti-inflammatory compound naproxen was shown to increase from nanorod to flower-like nanostructure to nanoparticle morphologies. (172) Reactions between Mn-oxides and organic compounds can also induce structural changes, for example, a buildup of interlayer Mn(II/III), that reduce reactivity over time. (173)

5.2. Characteristics of Organic Compounds That Influence Organo–Mineral Interactions

Molecular properties of OM, including molecular size, polarity, composition and position of various functional groups, also control OM affinity to Mn-oxides and their reactivity. (88,97) In this section, we discuss how intrinsic OM characteristics and various complexation mechanisms (e.g., inner-sphere and outer-sphere) determine reactivity with Mn-oxides. (53)
Soil OM is comprised of various functional groups, including acidic (e.g., carboxyl), neutral (e.g., alcohol), and basic groups (e.g., amine), and is, in general, negatively charged at near neutral pH. (95) Functional groups exert a strong control over how organic molecules adsorb to and react with mineral surfaces. The affinities of functional groups for metal ions vary, (95,97) generally decreasing in the order –O > −NH2 > −N = N >COO > −C=O. (95) Most DOC binds on Fe-oxide and clay minerals (e.g., goethite and montmorillonite) preferentially via carboxylate moieties, but N-rich organic compounds may have higher adsorption affinity to Mn-oxides. (94) Phenolic groups are particularly susceptible to oxidation by Mn-oxides. (43)
The effects of functional groups on organo-mineral interactions have been assessed by comparing adsorption and oxidation of several phenylarsonic acid analogues on birnessite. It has been observed that −OH and −NH2 groups promote adsorption via deprotonation and H-bonding and also donate electrons to birnessite to generate radicals, with −NH2 demonstrating a higher electron-donating ability than −OH. (97,99) In contrast, the electron-withdrawing −NO2 group decreases compound adsorption due to steric hindrance and also inhibits electron transfer and radical formation. (97) The position of these functional moieties also influences the reaction. For instance, −NH2 and −OH groups at ortho and para positions of the aromatic ring, which are the dominant positions of functional moieties, increase electron density at C1 and accelerate electron transfer from OM to birnessite. (97)
Differences in preferential adsorption between various organic compounds and minerals can lead to certain types of organic molecules being selectively removed from complex mixtures. For example, catechol forms predominately outer-sphere complexes on MnO2 but inner-sphere complexes on Fe2O3, TiO2, and Cr2O3 minerals, which may be ascribed to their different surface charges (negative vs positive, respectively). (174) Goethite preferentially adsorbs high molecular weight (MW) aromatic components from natural OM, while montmorillonite adsorbs low MW constituents without a preference toward aromatic or aliphatic OM. (106) In contrast, birnessite preferentially adsorbs aromatic moieties with no significant fractionation effects on the MW. (98,106)

5.3. Environmental Factors That Influence Organo–Mineral Interactions

Environmental factors influence the properties of both Mn-oxides and organic compounds, and consequently their interaction mechanisms and reaction rates. Here, we briefly elaborate on several environmental factors that influence organo-mineral interactions, including soil and soil solution pH (which influences surface charges, adsorption mechanisms, and redox potential), (95) redox conditions, (11) C:Mn ratios, (92) and pre-existing sorbed OM (175) and cations. (43)
Environmental pH alters surface charges of minerals and OM, consequently influencing adsorption mechanisms, controlling redox potential of Mn-oxides, and regulating interaction rates. (95) Oxygen atoms on Mn-oxide surfaces present as > Mn–O under neutral or alkaline conditions but > Mn–OH or > Mn–OH2+ under acidic conditions. (12) In general, Mn-oxides present net negative surface charges that electrostatically repel negatively charged moieties of OM at circumneutral to high pH, resulting in decreased OM adsorption capacity as pH increases. (96,97,99) In addition to charge effects, the redox potential of birnessite decreases as pH increases following the Nernst equation, leading to decreases in reaction rates with increasing pH. (176)
Redox conditions control the phases as well as reactivity of Mn-oxides. (11) Under oxic conditions, Mn(II) produced through reductive dissolution can sorb to Fe- or Mn-oxide surfaces and be reoxidized to Mn(IV), (43,106,177) accelerating degradation of organic compounds. (176) Mn-oxides can also degrade organic compounds under anoxic conditions, although possibly through different mechanisms than observed under oxic conditions. For example, in the presence of O2, NTMP degrades through intramolecular electron transfer following O2-promoted oxidation of NTMP–Mn(II) to NTMP–Mn(III). (146,178) Under anoxic conditions, NTMP directly transfers an electron to Mn(III) centers in MnOOH. (178)
C:Mn molar ratios determine the complexation capacity of Mn-oxides for OM. (92) For instance, sorption of DOC, leached from the O horizon of an Ultisol forest soil, to hydrous Mn-oxides was shown to increase with increasing C:Mn molar ratios (up to 10), ascribed to higher affinity of the oxide for organic C at higher C:Mn molar ratios. (92) Increasing C:Mn ratios were observed to slow adsorption kinetics but increase total adsorption of fulvic acid on birnessite, consequently increasing reductive dissolution. (98) For NOM–Mn(III) complexes, lower C:Mn ratios promote aggregation and removal from solution whereas higher C:Mn ratios (10–50) enhance the stability of these complexes in aqueous systems. (152)
Mineral coatings and sorbed compounds can also influence OM complexation by altering the surface properties of Mn-oxides. (106,175) Surface complexation of humic acid was shown to strongly influence products of Mn(II) oxidation on goethite surfaces, inhibiting β-MnOOH generation but not affecting γ-MnOOH and Mn3O4 formation. (43) Metal cation (e.g., Mn2+, Ca2+) sorption can balance negative surface charges and facilitate OM complexation. (96,103) For example, Zn2+, Mg2+, and Ca2+ sorption on birnessite surfaces enhances fulvic acid adsorption to birnessite and increases reductive dissolution. (164) Metal ion adsorption to mineral surfaces also affects Mn-oxide transformation, thus influencing the rates and mechanisms of interactions between SOM and Mn-oxides. Mn(II) adsorption to octahedral vacancies in phyllomanganates induces disproportionation with structural Mn(IV) that drives mineral transformation under acidic conditions, with more subtle effects observed at neutral pH. (179) Moreover, the transformation of Mn(II)-bearing birnessite to diverse sheet and tunneled Mn-oxide structures (γ-MnO2, δ-MnO2, triclinic birnessite, 4 × 4 tunneled Mn-oxides) is regulated by the presence of specific ions (e.g., Cu2+, Mg2+, Ca2+, Li+, Na+, or K+). (180)
Although many studies investigate rates and mechanisms of reactions between single phase Mn-oxides and individual organic compounds, (10,53,105,116,120) few studies examine interactions with NOM or organic mixtures, for example, preferential adsorption of high versus low molecular weight compounds or aliphatic versus aromatic structures. (96,106) Given the complexity of organic molecules in soil solution, more efforts are needed to determine how Mn-oxides interact with specific organic compounds in complex OM mixtures under varied environmental conditions.

6. Organic Matter Stabilization and Destabilization Potential in Soils

Jump To

Major soil minerals (clays, Fe-oxides, Al-oxides) are well-studied for their interactions with OM (4,181−184) relative to Mn-oxides. Here, we synthesize data quantifying the ability of Mn(III/IV)-oxides and Fe(III)-oxides to stabilize (immobilize) or destabilize (degrade) organic molecules. We compile existing data from laboratory studies that span a range of mineralogies, reactions conditions, and organic constituents (SI Table S3). We also evaluate quantities of organic carbon associated with metal oxides in soils and sediments.
Given the nonspecificity and heterogeneity in data availability (see SI for further discussion), our strategy is to present information on “stabilization” and “destabilization” of organic compounds (Figure 3; SI Table S3). Stabilization (S, μmol C g-oxide–1) refers to removal of organic C from solution during reaction with the oxide but does not consider underlying mechanisms (e.g., sorption, polymerization into insoluble phases) and does not reflect any potential transformation that occurs during stabilization. These values represent interactions between either individual compounds or NOM with either synthetic or natural oxides. Destabilization (D, μmol C g-oxide–1 h–1) refers to the transformation of organic molecules during reaction with the oxide, that is, OM oxidation coupled to reductive dissolution of the oxide. These values primarily reflect interactions between individual compounds and synthetic minerals but largely exclude reactions with NOM due to difficulties quantifying transformation in these complex systems. A broad suite of compounds that largely consists of small organic molecules is considered.

Figure 3

Figure 3. Compilation of (de)stabilization potentials of a broad suite of organic compounds by Mn-oxides (blue boxes) and Fe-oxides (orange boxes) (SI Table S3). (a) Destabilization rates (log-scale) are presented per mass oxide (μg C g-oxide–1 h–1) and per mass soil (μg C g-soil–1 h–1). (b) Comparison of destabilization rates (μg C g-oxide–1 h–1) of individual organic compounds by Mn- and Fe-oxides. Compound structures are shown below each name. (c) Stabilization (log-scale) for Mn-oxides and Fe-oxides are presented per mass oxide (μg C g-oxide–1) and per mass soil (μg C g-soil–1) and are compared with C stabilization by reactive minerals (dithionite-soluble) in soils and sediments. Differences between Mn-oxides and Fe-oxides within each group are shown as significant (* = p < 0.05) or highly significant (*** = p < 0.001) based on oneway ANOVA.

To encompass the wide variety of experimental conditions, we report the highest rates of stabilization or destabilization from each study and compile these across commonly used Mn(III/IV)-oxides (e.g., birnessite, Manganite) and Fe(III)-oxides (e.g., ferrihydrite, goethite, hematite). For all minerals, adsorption and reaction rates tend to be higher at low pH (<4) where increasingly positive surface charge facilitates interactions with negatively charged organic compounds. (99,185) The focus on low pH systems also reduces complexities associated with the sorption or reoxidation of reduced metals given that Mn2+ and Fe2+ ions released through reductive dissolution remain in solution at low pH. (186) However, many studies conducted only at circumneutral pH (∼7) are still included with careful consideration for comparability. For both stabilization and destabilization, we excluded compounds that did not show measurable effects. For example, compounds that only adsorbed to mineral surfaces were excluded from compiled destabilization values unless directly comparing interactions of specific compounds. To scale these observations to soil, we also multiplied oxide stabilization (μg C g-oxide–1) and destabilization (μg C g-oxide–1 h–1) by approximate oxide concentrations in surface soils (Fe-oxide ∼20 000 μg g-soil–1 and Mn-oxide ∼2000 μg g-soil–1) to obtain soil stabilization (μg C g-soil–1) and soil destabilization rates (μg C g-soil–1 h–1), with the caveat that reaction rates reported for synthetic minerals are typically higher than those reported for environmental samples. (10)


Mn-oxides demonstrated a higher average oxide destabilization potential (log D = 4.08 ± 0.14 μg C g-oxide–1 h–1) and broader reactivity than Fe-oxides (log D = 3.52 ± 0.19 μg C g-oxide–1 h–1) (p < 0.001) (Figure 3a). In contrast, Fe-oxides showed higher average soil destabilization potential (log D = 1.82 ± 0.19 μg C g-soil–1 h–1) than Mn-oxides (log D = 1.38 ± 0.14 μg C g-soil–1 h–1) due to the higher abundance of Fe-oxides in soils, although this difference was not considered significant (p = 0.06). Importantly, destabilization rates reported for Mn-oxides largely represent direct oxidation of diverse organic compounds on the mineral surface. Mn(III/IV)-oxides are known to oxidize a large suite of organic molecules over a wide range of environmentally relevant pH conditions, whereas Fe(III)-oxides are thermodynamically constrained to interactions with strong reductants at low pH (<4). (187) As such, transformation rates reported here for Fe(III)-oxides are limited to select molecules (e.g., ascorbate, quinones, flavins) that serve as electron shuttles and are not degraded.
To more directly compare destabilization potentials between these minerals, we also examined a subset of organic molecules (catechol, glyphosate, hydroquinone, oxalic acid, pyrogallol, and resorcinol) for which reactions with both Fe(III)- and Mn(III/IV)-oxides are reported (Figure 3b; SI Table S3). All examined organic molecules reacted with Mn(III/IV)-oxides at rates ≥103 μg C g-oxide–1 h–1 (pH 4–7), whereas Fe(III)-oxides were only reactive with phenolic compounds at pH ≤ 4 with rates ≤103 μg C g-oxide–1 h–1.


Mn-oxides exhibited higher oxide stabilization potential (log S = 4.33 ± 0.21 μg C g-oxide–1) than Fe-oxides (log S = 3.69 ± 0.17 μg C g-oxide–1) (p = 0.02) but with a broader spread (Figure 3c). The potential for soil stabilization was similar for Mn-oxides (log S = 1.63 ± 0.21 μg C g-soil–1) and Fe-oxides (log S = 1.99 ± 0.17 μg C g-soil–1) (p = 0.20), indicating a role for C stabilization by Mn-oxides that is comparable to Fe-oxides. However, stabilization by Fe-oxides may persist for longer than Mn-oxides given that OM can readily desorb from Mn-oxides, (92) whereas release from Fe-oxide may depend on mineral dissolution.
We also examined concentrations of organic C associated with reactive minerals as reported for soils and sediments. (188−190) Reactive minerals represent those extracted from soils with either dithionite and/or hydroxylamine hydrochloride and are broadly interpreted to represent Fe-oxides; as such, we present oxide and soil concentrations for comparison with mineral studies. Soil concentrations (μg C g-soil–1) were reported in the literature while oxide concentrations (μg C g-oxide–1) were calculated based on concentrations of extractable Fe and the assumption that Fe-oxides are dominantly in the FeOOH stoichiometry.
Organic C stabilization by reactive minerals, calculated as log S = 5.5 ± 0.1 μg C g-oxide–1 and log S = 3.4 ± 0.1 μg C g-soil–1, was 10–100× higher than stabilization reported for individual minerals, representing an apparent ability for Fe-oxides to stabilize substantially more organic C in soil than in lab studies. However, it is likely that “Fe-bound” OM is also derived from Mn-oxides and potentially Al-oxides, which are also at least partially extracted by dithionite. (191) Thus, this stabilization may represent organic C bound to multiple minerals. Furthermore, lab studies are typically designed to probe sorption processes, whereas soil minerals are associated with organic C through multiple mechanisms. Sorption processes may represent only a small portion of the total C stabilization in soils, (192) and thus certain lab studies may underestimate stabilization potential.
Our analyses demonstrate that Mn- and Fe-oxides may serve as reservoirs for organic C in soils, whereas Mn-oxides also possess a high potential to oxidize and degrade organic molecules in a manner not yet understood in natural systems. Further characterization of oxide–OM interactions in soils is needed to validate the environmental relevance of these trends. Laboratory experiments summarized here indicate that Fe- and Mn-oxides can bind organic compounds and theoretically protect them from microbial degradation; however, the persistence of these associations over time are unknown.

7. Future Framework for Investigating Mn–C Interactions in Terrestrial Ecosystems

Jump To

Mn and C biogeochemistry intersect through diverse processes that degrade or protect organic molecules. Manganese demonstrates a potential to stabilize organic molecules via multiple organo-mineral associations (Figure 1) at scales on par with Fe (Figure 3c) but also acts as a transformer that processes organic molecules through biotic (Figure 2) and abiotic (Figure 3a,b) pathways. From these observations, we hypothesize that increasing Mn availability may decrease ecosystem C storage. However, the spatiotemporal variability of these interactions in soils and the magnitude of their effects on soil C storage remain unknown. Similar to Fe, (163) the ability of Mn to protect or degrade OM depends on the stability and reactivity of Mn(III) and Mn(IV) phases under particular environmental conditions at a given time. Although SOM is often considered to be stabilized by Fe-minerals, Fe cycling can also stimulate soil C loss, specifically in anoxic or redox-dynamic soils. Even under oxic conditions, Fe-driven Fenton reactions promote SOM mineralization. (163) Mn has similar interactions with OM: it can physically bind OM to form aggregates, adsorb or coprecipitate with SOM, and transfer electrons that destabilize and/or degrade SOM. The contribution of each role is highly dependent on environmental conditions (soil moisture, redox potential, oxygen availability) and the crystallinity and stability of Mn phases.
Here, we present open questions in Mn–C interactions and highlight opportunities to further explore how Mn regulates C cycling in terrestrial ecosystems. In particular, we focus on Mn–C interactions in soils, while also noting that plant–soil interactions are essential for placing soil processes within a broader ecosystem context. For example, to constrain ecosystem C dynamics, it is necessary to parameterize species-dependent abilities of plants to recycle and retain Mn within surface soils, the influence of varied Mn uptake on biochemical processes such as C fixation, and combined effects of lignin and Mn content on litter decomposition.

Over What Time Scales do Mn-Oxides Stabilize Organic Matter?

It is necessary to establish how long and under what environmental conditions Mn-oxide minerals serve as carbon reservoirs. Although other minerals (e.g., clays, Al-oxides, Fe-oxides) may store organic C for decades to millennia, (6) the magnitude and persistence of OM stabilization by Mn-oxides remains unknown. It is possible that these associations are relatively fleeting. Organic compounds that sorb onto a Mn-oxide surface may undergo redox transformation or be desorbed back into solution. Reductive dissolution of Mn-oxides may release associated organic molecules, as has been widely observed for Fe-oxides. (163) Isotopic techniques have been used to constrain ages (14C) and sources (13C) of organic molecules associated with other minerals (6,188) but have not been applied to Mn-oxides. Indeed, isolating Mn-oxides from soil and characterizing associated OM is technically challenging given its similarities with Fe-oxides. In order to assess the contribution of Mn-oxides to carbon storage in soils, it is necessary to quantify the amount, composition, and age of organic molecules associated with Mn-oxides across diverse environments.

How Do Mn–C Interactions Vary in Space and Time?

Many studies have focused on abiotic transformation of synthetic Mn-oxide phases under specified pH, temperature, and/or solution conditions. (35,46,48,121,193) However, the stability and phase transformation of various Mn-oxides in soils remain poorly studied. Long-term weathering of Mn-oxides in complex soil environments likely results in Mn translocation and mineral phase changes that modify mineral surface properties and OM complexation capacity. Manganese is subject to redistribution at pore-scales (93) and across soil profiles as Mn-oxides dissolve and reprecipitate under fluctuating redox. Different Mn-oxides exhibit diverse properties (e.g., SSA, PZC, Mn oxidation states) (20,21) that strongly regulate their interactions with OM. Thus, it is necessary to evaluate mineral transformation in soil environments to determine the ability of Mn-oxides to stabilize or destabilize OM across wide-ranging ecosystems. Additionally, Mn-oxides strongly adsorb a large variety of metal cations (e.g., Ca2+, Pb2+, Zn2+, Co2+) (49,118,161) and oxyanions (e.g., AsO33–, PO43–) (161,168,194) that compete with organic compounds for reactive surface sites. Further research will be necessary to determine practical reactivity of Mn-oxides in complex soil environments when interacting with SOM.

What Role Does Mn(III) Play in Oxidizing Organic Compounds in Soils?

Although dissolved Mn(III) has long been considered a negligible species in the environment, several recent studies have identified abundant dissolved Mn(III) in marine, estuarine, sediment, and sediment porewater systems. (150,151,195) Diverse compounds including small organic acids, siderophores, and humic acids can form Mn(III)-complexes that are stable relative to the aqueous ion. These Mn(III)-complexes can in turn oxidize organic and inorganic species. (76,123,148,155,195) However, despite their known occurrence and reactivity, the formation, persistence, and reactivity of Mn(III)-complexes in soil environments are not established.

How Do Microbial Mn–C Interactions Influence Soil C Processing?

Increased Mn(II) availability reportedly increases MnP activity and decreases lignin content in litter and organic soils, (130,144,196) demonstrating the importance of Mn(II) in stimulating MnP-mediated lignin oxidation (Figure 2, SI Table S2). However, it remains unclear how relationships among Mn(II) supply, MnP expression and activity, and microbial community composition vary across biomes. Accounting for MnP activity and microbial communities capable of using MnP to degrade complex OM is an important step toward elucidating mechanisms driving the negative relationship between foliar Mn and litter mass loss reported by multiple studies. Furthermore, various microorganisms use extracellular multicopper oxidases and reactive oxygen species to oxidize Mn(II) to reactive Mn(III) species that may be important SOM oxidizers independent of MnP activity.

How Do Mn–C Interactions Influence Soil C Turnover and Storage?

Current knowledge of the influence of Mn on C turnover is limited to Mn manipulation studies conducted in limited laboratory and field studies. These studies show increased C losses (as CO2 and DOC) in response to Mn addition to soils during litter decomposition (Figure 2, SI Table S2), (143−145) suggesting that increased Mn supply to soil would lead to decreased soil C storage. It is possible that these effects are restricted to surface soil where biological activity and the supply of OM and Mn(II) from litterfall are high compared with subsoils. (155) However, the net effect of Mn on C storage has not been validated experimentally. Progress toward a better understanding of Mn effects on C persistence in soils would be enhanced with studies designed to capture changes in C stocks over time and to account for other environmental variables that are likely to influence Mn and C cycling, such as temperature, precipitation, and the plant community.
These individual questions provide context for a larger objective: quantifying the influence of Mn-mediated processes on C storage across diverse ecosystems. Addressing this objective will require detailed investigation of complex soil processes as well as coupled modeling studies that can extend observations through space and time. These efforts potentially fill a knowledge gap in our understanding of processes that regulate ecosystem C dynamics. Earth system models assessed by the Coupled Model Intercomparison Project Phase 5 (CMIP5) present large uncertainty with respect to current soil C storage and turnover rates, limiting our ability to predict how soil C will respond to changing climate. (197) Improved understanding and quantification of key geochemical processes, such as the coupled Mn–C interactions presented here, may better inform and constrain predictions into the future.

Supporting Information

Jump To

The Supporting Information is available free of charge at

Terms & Conditions

Most electronic Supporting Information files are available without a subscription to ACS Web Editions. Such files may be downloaded by article for research use (if there is a public use license linked to the relevant article, that license may permit other uses). Permission may be obtained from ACS for other uses through requests via the RightsLink permission system:

Author Information

Jump To

  • Corresponding Author
    • Elizabeth Herndon - Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United StatesDepartment of Earth and Planetary Sciences, College of Arts & Sciences, University of Tennessee, Knoxville, Tennessee 37996, United StatesOrcid Email: [email protected]
  • Authors
    • Hui Li - Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United States
    • Fernanda Santos - Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United States
    • Kristen Butler - Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 37831, United StatesDepartment of Earth and Planetary Sciences, College of Arts & Sciences, University of Tennessee, Knoxville, Tennessee 37996, United States
  • Notes
    The authors declare no competing financial interest.


Jump To

This work was sponsored by the Laboratory Directed Research and Development Program of Oak Ridge National Laboratory, managed by UT-Battelle, LCC for the U.S. Department of Energy under contract DE-AC05-00OR22725. We gratefully acknowledge Nathan Armistead (ORNL) for graphics development and five anonymous reviewers for their constructive comments.


Jump To

This article references 197 other publications.

  1. 1
    Lehmann, J.; Kleber, M. The contentious nature of soil organic matter. Nature 2015, 528, 6068,  DOI: 10.1038/nature16069
  2. 2
    Ciais, P.; Sabine, C.; Bala, G.; Bopp, L.; Brovkin, V.; Canadell, J.; Chhabra, A.; DeFries, R.; Galloway, J.; Heimann, M.; Jones, C. Carbon and other biogeochemical cycles. Climate change 2013: the physical science basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. 2013, 95123
  3. 3
    Friedlingstein, P.; Jones, M. W.; O’Sullivan, M.; Andrew, R. M.; Hauck, J.; Peters, G. P.; Peters, W.; Pongratz, J.; Sitch, S.; Le Quéré, C.; Bakker, D. C. E.; Canadell, J. G.; Ciais, P.; Jackson, R. B.; Anthoni, P.; Barbero, L.; Bastos, A.; Bastrikov, V.; Becker, M.; Bopp, L.; Buitenhuis, E.; Chandra, N.; Chevallier, F.; Chini, L. P.; Currie, K. I.; Feely, R. A.; Gehlen, M.; Gilfillan, D.; Gkritzalis, T.; Goll, D. S.; Gruber, N.; Gutekunst, S.; Harris, I.; Haverd, V.; Houghton, R. A.; Hurtt, G.; Ilyina, T.; Jain, A. K.; Joetzjer, E.; Kaplan, J. O.; Kato, E.; Klein Goldewijk, K.; Korsbakken, J. I.; Landschützer, P.; Lauvset, S. K.; Lefèvre, N.; Lenton, A.; Lienert, S.; Lombardozzi, D.; Marland, G.; McGuire, P. C.; Melton, J. R.; Metzl, N.; Munro, D. R.; Nabel, J. E. M. S.; Nakaoka, S.-I.; Neill, C.; Omar, A. M.; Ono, T.; Peregon, A.; Pierrot, D.; Poulter, B.; Rehder, G.; Resplandy, L.; Robertson, E.; Rödenbeck, C.; Séférian, R.; Schwinger, J.; Smith, N.; Tans, P. P.; Tian, H.; Tilbrook, B.; Tubiello, F. N.; van der Werf, G. R.; Wiltshire, A. J.; Zaehle, S. Global Carbon Budget 2019. Earth System Science Data 2019, 11 (4), 17831838,  DOI: 10.5194/essd-11-1783-2019
  4. 4
    Kleber, M.; Eusterhues, K.; Keiluweit, M.; Mikutta, C.; Mikutta, R.; Nico, P. Mineral-organic associations: Formation, properties, and relevance in soil environments. Adv. Agron. 2015, 130, 1140,  DOI: 10.1016/bs.agron.2014.10.005
  5. 5
    Cotrufo, M. F.; Ranalli, M. G.; Haddix, M. L.; Six, J.; Lugato, E. Soil carbon storage informed by particulate and mineral-associated organic matter. Nat. Geosci. 2019, 12 (12), 989994,  DOI: 10.1038/s41561-019-0484-6
  6. 6
    Torn, M. S.; Trumbore, S. E.; Chadwick, O. A.; Vitousek, P. M.; Hendricks, D. M. Mineral control of soil organic carbon storage and turnover. Nature 1997, 389 (6647), 170,  DOI: 10.1038/38260
  7. 7
    Lehmann, J.; Hansel, C. M.; Kaiser, C.; Kleber, M.; Maher, K.; Manzoni, S.; Nunan, N.; Reichstein, M.; Schimel, J. P.; Torn, M. S.; Wieder, W. R.; Kögel-Knabner, I. Persistence of soil organic carbon caused by functional complexity. Nat. Geosci. 2020, 13 (8), 529534,  DOI: 10.1038/s41561-020-0612-3
  8. 8
    Davidson, E. A.; Janssens, I. A. Temperature sensitivity of soil carbon decomposition and feedbacks to climate change. Nature 2006, 440 (7081), 165173,  DOI: 10.1038/nature04514
  9. 9
    Doetterl, S.; Stevens, A.; Six, J.; Merckx, R.; Van Oost, K.; Casanova Pinto, M.; Casanova-Katny, A.; Muñoz, C.; Boudin, M.; Zagal Venegas, E.; Boeckx, P. Soil carbon storage controlled by interactions between geochemistry and climate. Nat. Geosci. 2015, 8 (10), 780783,  DOI: 10.1038/ngeo2516
  10. 10
    Remucal, C. K.; Ginder-Vogel, M. A critical review of the reactivity of manganese oxides with organic contaminants. Environ. Sci. Process Impacts 2014, 16 (6), 12471266,  DOI: 10.1039/c3em00703k
  11. 11
    Spiro, T. G.; Bargar, J. R.; Sposito, G.; Tebo, B. M. Bacteriogenic manganese oxides. Acc. Chem. Res. 2010, 43, 29,  DOI: 10.1021/ar800232a
  12. 12
    Luther, G. W.; Thibault de Chanvalon, A.; Oldham, V. E.; Estes, E. R.; Tebo, B. M.; Madison, A. S. Reduction of Manganese Oxides: Thermodynamic, Kinetic and Mechanistic Considerations for One- Versus Two-Electron Transfer Steps. Aquat. Geochem. 2018, 24 (4), 257277,  DOI: 10.1007/s10498-018-9342-1
  13. 13
    Hansel, C. M. Manganese in Marine Microbiology. Adv. Microb. Physiol. 2017, 70, 3783,  DOI: 10.1016/bs.ampbs.2017.01.005
  14. 14
    Berg, B.; Erhagen, B.; Johansson, M.-B.; Nilsson, M.; Stendahl, J.; Trum, F.; Vesterdal, L. Manganese in the litter fall-forest floor continuum of boreal and temperate pine and spruce forest ecosystems - A review. For. Ecol. Manage. 2015, 358, 248260,  DOI: 10.1016/j.foreco.2015.09.021
  15. 15
    Turekian, K. K.; Wedepohl, K. H. Distribution of the elements in some major units of the Earth’s crust. Geol. Soc. Am. Bull. 1961, 72, 175192,  DOI: 10.1130/0016-7606(1961)72[175:DOTEIS]2.0.CO;2
  16. 16
    U.S. EPA, Health effects support document for manganese. U.S. Environmental Protection Agency, Office of Water, EPA-822-R-03-003; EPA: Washington, D.C., 2003.
  17. 17
    Hochella, M. F. J.; Kasama, T.; Putnis, A.; Putnis, C. V.; Moore, J. N. Environmental important, poorly crystalline Fe/Mn hydrous oxides: Ferrihydrite and a possibly new vernadite-like mineral from the Clark Fork River Superfund Complex. Am. Mineral. 2005, 90, 718724,  DOI: 10.2138/am.2005.1591
  18. 18
    Morgan, J. J. Kinetics of reaction between O2 and Mn(II) species in aqueous solutions. Geochim. Cosmochim. Acta 2005, 69, 3548,  DOI: 10.1016/j.gca.2004.06.013
  19. 19
    Post, J. E. Manganese oxide minerals: Crystal structures and economic and environmental significance. Proc. Natl. Acad. Sci. U. S. A. 1999, 96, 34473454,  DOI: 10.1073/pnas.96.7.3447
  20. 20
    Feng, X. H.; Zhai, L. M.; Tan, W. F.; Liu, F.; He, J. Z. Adsorption and redox reactions of heavy metals on synthesized Mn oxide minerals. Environ. Pollut. 2007, 147, 366373,  DOI: 10.1016/j.envpol.2006.05.028
  21. 21
    Tebo, B. M.; Bargar, J. R.; Clement, B. G.; Dick, G. J.; Murray, K. J.; Parker, D.; Verity, R.; Webb, S. M. Biogenic mManganese Oxides: Properties and Mechanisms of Formation. Annu. Rev. Earth Planet. Sci. 2004, 32 (1), 287328,  DOI: 10.1146/
  22. 22
    Lovley, D. R.; Holmes, D. E.; Nevin, K. P. Dissimilatory Fe (III) and Mn (IV) reduction. Adv. Microb. Physiol. 2004, 49 (2), 219286,  DOI: 10.1016/S0065-2911(04)49005-5
  23. 23
    Brock, S. L. D. N.; Tian, Z. R.; Giraldo, O.; Zhou, H.; Suib, S. L. A review of porous manganese oxide materials. Chem. Mater. 1998, 10, 26192628,  DOI: 10.1021/cm980227h
  24. 24
    Learman, D. R.; Wankel, S. D.; Webb, S. M.; Martinez, N.; Madden, A. S.; Hansel, C. M. Coupled biotic-abiotic Mn(II) oxidation pathway mediates the formation and structural evolution of biogenic Mn oxides. Geochim. Cosmochim. Acta 2011, 75, 60486063,  DOI: 10.1016/j.gca.2011.07.026
  25. 25
    Santelli, C. M.; Webb, S. M.; Dohnalkova, A. C.; Hansel, C. M. Diversity of Mn oxides produced by Mn(II)-oxidizing fungi. Geochim. Cosmochim. Acta 2011, 75 (10), 27622776,  DOI: 10.1016/j.gca.2011.02.022
  26. 26
    Villalobos, M.; Lanson, B.; Manceau, A.; Toner, B.; Sposito, G. Structural model for the biogenic Mn oxide produced by pseudomonas putida. Am. Mineral. 2006, 91, 489502,  DOI: 10.2138/am.2006.1925
  27. 27
    Learman, D. R.; Voelker, B. M.; Vazquez-Rodriguez, A. I.; Hansel, C. M. Formation of manganese oxides by bacterially generated superoxide. Nat. Geosci. 2011, 4 (2), 9598,  DOI: 10.1038/ngeo1055
  28. 28
    Chaput, D. L.; Fowler, A. J.; Seo, O.; Duhn, K.; Hansel, C. M.; Santelli, C. M. Mn oxide formation by phototrophs: Spatial and temporal patterns, with evidence of an enzymatic superoxide-mediated pathway. Sci. Rep. 2019, 9 (1), 18244,  DOI: 10.1038/s41598-019-54403-8
  29. 29
    Bohu, T.; Santelli, C. M.; Akob, D. M.; Neu, T. R.; Ciobota, V.; Rosch, P.; Popp, J.; Nietzsche, S.; Kusel, K. Characterization of pH dependent Mn(II) oxidation strategies and formation of a bixbyite-like phase by Mesorhizobium australicum T-G1. Front. Microbiol. 2015, 6, 734,  DOI: 10.3389/fmicb.2015.00734
  30. 30
    Butterfield, C. N.; Soldatova, A. V.; Lee, S.-W.; Spiro, T. G.; Tebo, B. M. Mn(II,III) oxidation and MnO2 mineralization by an expressed bacterial multicopper oxidase. Proc. Natl. Acad. Sci. U. S. A. 2013, 110, 1173111735,  DOI: 10.1073/pnas.1303677110
  31. 31
    Andeer, P. F.; Leadbetter, J. R.; Mcllvin, M.; Dunn, J. A.; Hansel, C. M. Extracellular haem peroxidases mediate Mn(II) oxidation in a marine Roseobacter bacterium via superoxide production. Environ. Microbiol. 2015, 17, 39253936,  DOI: 10.1111/1462-2920.12893
  32. 32
    Tang, Y. Z. C. A.; Santelli, C. M.; Hansel, C. M. Fungal oxidative dissolution of the Mn(II)-bearing mineral rhodochrosite and the role of metabolites in manganese oxide formation. Environ. Microbiol. 2013, 15, 10631077,  DOI: 10.1111/1462-2920.12029
  33. 33
    Yu, H.; Leadbetter, J. R. Bacterial chemolithoautotrophy via manganese oxidation. Nature 2020, 583, 453458,  DOI: 10.1038/s41586-020-2468-5
  34. 34
    Hansel, C. M.; Learman, D. R. Chapter eighteen: Geomicrobiology of manganese. In Ehrlich’s Geomicrobiology; CRC Press: Boca Raton, FL, 2015 401 452
  35. 35
    Elzinga, E. J. Reductive transformation of birnessite by aqueous Mn(II). Environ. Sci. Technol. 2011, 45, 63666372,  DOI: 10.1021/es2013038
  36. 36
    van Genuchten, C. M.; Pena, J. Mn(II) Oxidation in Fenton and Fenton Type Systems: Identification of Reaction Efficiency and Reaction Products. Environ. Sci. Technol. 2017, 51 (5), 29822991,  DOI: 10.1021/acs.est.6b05584
  37. 37
    Jung, H.; Chadha, T. S.; Kim, D.; Biswas, P.; Jun, Y. S. Photochemically assisted fast abiotic oxidation of manganese and formation of delta-MnO2 nanosheets in nitrate solution. Chem. Commun. (Cambridge, U. K.) 2017, 53 (32), 44454448,  DOI: 10.1039/C7CC00754J
  38. 38
    Zhang, T.; Liu, L.; Tan, W.; Suib, S. L.; Qiu, G. Formation and transformation of manganese(III) intermediates in the photochemical generation of manganese(IV) oxide minerals. Chemosphere 2021, 262, 128082,  DOI: 10.1016/j.chemosphere.2020.128082
  39. 39
    Nico, P.; Anastasio, C.; Zasoski, R. Rapid photo-oxidation of Mn(II) mediated by humic substances. Geochim. Cosmochim. Acta 2002, 66 (23), 40474056,  DOI: 10.1016/S0016-7037(02)01001-3
  40. 40
    Xu, X.; Li, Y.; Li, Y.; Lu, A.; Qiao, R.; Liu, K.; Ding, H.; Wang, C. Characteristics of desert varnish from nanometer to micrometer scale: A photo-oxidation model on its formation. Chem. Geol. 2019, 522, 5570,  DOI: 10.1016/j.chemgeo.2019.05.016
  41. 41
    Lan, S.; Wang, X.; Xiang, Q.; Yin, H.; Tan, W.; Qiu, G.; Liu, F.; Zhang, J.; Feng, X. Mechanisms of Mn(II) catalytic oxidation on ferrihydrite surfaces and the formation of manganese (oxyhydr)oxides. Geochim. Cosmochim. Acta 2017, 211, 7996,  DOI: 10.1016/j.gca.2017.04.044
  42. 42
    Wang, X.; Lan, S.; Zhu, M.; Ginder-Vogel, M.; Yin, H.; Liu, F.; Tan, W.; Feng, X. The presence of ferrihydrite promotes abiotic formation of manganese (oxyhydr)oxides. Soil Sci. Soc. Am. J. 2015, 79, 12971305,  DOI: 10.2136/sssaj2014.12.0502
  43. 43
    Ma, D.; Wu, J.; Yang, P.; Zhu, M. Coupled Manganese Redox Cycling and Organic Carbon Degradation on Mineral Surfaces. Environ. Sci. Technol. 2020, 54 (14), 88018810,  DOI: 10.1021/acs.est.0c02065
  44. 44
    Learman, D. R.; Wankel, S. D.; Webb, S. M.; Martinez, N.; Madden, A. S.; Hansel, C. M. Coupled biotic-abiotic Mn(II) oxidation pathway mediates the formation and structural evolution of biogenic Mn oxides. Geochim. Cosmochim. Acta 2011, 75 (20), 60486063,  DOI: 10.1016/j.gca.2011.07.026
  45. 45
    Madden, A. S.; Hochella, M. F. A test of geochemical reactivity as a function of mineral size: Manganese oxidation promoted by hematite nanoparticles. Geochim. Cosmochim. Acta 2005, 69 (2), 389398,  DOI: 10.1016/j.gca.2004.06.035
  46. 46
    Luo, Y. T. W.; Suib, S. L.; Qiu, G.; Liu, F. Dissolution and phase transformation processes of hausmannite in acidic aqueous systems under anoxic conditions. Chem. Geol. 2018, 487, 5462,  DOI: 10.1016/j.chemgeo.2018.04.016
  47. 47
    Tu, S. R. G. J.; Goh, T. B. Transformation of synthetic birnessite as affected by pH and manganese concentration. Clays Clay Miner. 1994, 42, 321330,  DOI: 10.1346/CCMN.1994.0420310
  48. 48
    Lefkowitz, J. P.; Rouff, A. A.; Elzinga, E. J. Influence of pH on the reductive transformation of birnessite by aqueous Mn(II). Environ. Sci. Technol. 2013, 47, 1036410371,  DOI: 10.1021/es402108d
  49. 49
    Yin, H.; Li, H.; Wang, Y.; Ginder-Vogel, M.; Qiu, G.; Feng, X.; Zheng, L.; Liu, F. Effects of Co and Ni co-doping on the structure and reactivity of hexagonal birnessite. Chem. Geol. 2014, 381, 1020,  DOI: 10.1016/j.chemgeo.2014.05.017
  50. 50
    Webb, S. M.; Fuller, C. C.; Tebo, B. M.; Bargar, J. R. Determination of uranyl incorporation into biogenic manganese oxides using X-ray absorption spectroscopy and scattering. Environ. Sci. Technol. 2006, 40, 771777,  DOI: 10.1021/es051679f
  51. 51
    Ruiz-Garcia, M. V. M.; Voegelin, A.; Pi-Puig, T.; Martínez-Villegas, N.; Göttlicher, J. Transformation of hexagonal birnessite upon reaction with thallium(I): Effects of birnessite crystallinity, pH, and thallium concentration. Environ. Sci. Technol. 2021, 55, 48624870,  DOI: 10.1021/acs.est.0c07886
  52. 52
    Lin, H.; Szeinbaum, N. H.; DiChristina, T. J.; Taillefert, M. Geochim. Cosmochim. Acta 2012, 99, 179192,  DOI: 10.1016/j.gca.2012.09.020
    Lin, H.; Szeinbaum, N. H.; DiChristina, T. J.; Taillefert, M. Microbial Mn(IV) reduction requires an initial one-electron reductive solubilization step. Geochim. Cosmochim. Acta 2012, 99, 179192,  DOI: 10.1016/j.gca.2012.09.020
  53. 53
    Stone, A. T.; Morgan, J. J. Reduction and Dissolution of Manganese(III) and Manganese(IV) Oxides by Organics: 2. Survey of the Reactivity of Organics. Environ. Sci. Technol. 1984, 18, 617624,  DOI: 10.1021/es00126a010
  54. 54
    Wang, Y.; Stone, A. T. Reaction of MnIII,IV (hydr)oxides with oxalic acid, glyoxylic acid, phosphonoformic acid, and structurally-related organic compounds. Geochim. Cosmochim. Acta 2006, 70 (17), 44774490,  DOI: 10.1016/j.gca.2006.06.1548
  55. 55
    Sparrow, L. A.; Uren, N. C. Manganese oxidation and reduction in soils: Effects of temperature, water potential, pH and their interactions. Soil Res. 2014, 52, 483494,  DOI: 10.1071/SR13159
  56. 56
    Jobbágy, E. J.; Jackson, R. B. The uplift of soil nutrients by plants: biogeochemical consequences across scales. Ecology 2004, 85 (9), 23802389,  DOI: 10.1890/03-0245
  57. 57
    Oh, N.-H.; Richter, D. D. Elemental translocation and loss from three highly weathered soil-bedrock profiles in the southeastern United States. Geoderma 2005, 126 (1–2), 525,  DOI: 10.1016/j.geoderma.2004.11.005
  58. 58
    Broadley, M.; Brown, P.; Cakmak, I.; Rengel, Z.; Zhao, F., Function of nutrients: micronutrients. In Marschner’S Mineral Nutrition of Higher Plants; Elsevier, 2012; pp 191248.
  59. 59
    Herndon, E. M.; Martínez, C. E.; Brantley, S. L. Spectroscopic (XANES/XRF) characterization of contaminant manganese cycling in a temperate watershed. Biogeochemistry 2014, 121 (3), 505517,  DOI: 10.1007/s10533-014-0018-7
  60. 60
    Fernando, D. R.; Mizuno, T.; Woodrow, I. E.; Baker, A. J.; Collins, R. N. Characterization of foliar manganese (Mn) in Mn (hyper)accumulators using X-ray absorption spectroscopy. New Phytol. 2010, 188 (4), 101427,  DOI: 10.1111/j.1469-8137.2010.03431.x
  61. 61
    McCain, D. C.; Markley, J. L. More manganese accumulates in maple sun leaves than in shade leaves. Plant Physiol. 1989, 90 (4), 14171421,  DOI: 10.1104/pp.90.4.1417
  62. 62
    Keiluweit, M.; Nico, P.; Harmon, M. E.; Mao, J.; Pett-Ridge, J.; Kleber, M. Long-term litter decomposition controlled by manganese redox cycling. Proc. Natl. Acad. Sci. U. S. A. 2015, 112 (38), E525360,  DOI: 10.1073/pnas.1508945112
  63. 63
    Herndon, E. M.; Jin, L.; Andrews, D. M.; Eissenstat, D. M.; Brantley, S. L. Importance of vegetation for manganese cycling in temperate forested watersheds. Global Biogeochem. Cy. 2015, 29 (2), 160174,  DOI: 10.1002/2014GB004858
  64. 64
    Marschner, H., Mechanisms of manganese acquisition by roots from soils. In Manganese in Soils and Plants; Springer, 1988; pp 191204.
  65. 65
    Horsley, S. B.; Long, R. P.; Bailey, S. W.; Hallett, R. A.; Hall, T. J. Factors associated with the decline disease of sugar maple on the Allegheny Plateau. Can. J. For. Res. 2000, 30 (9), 13651378,  DOI: 10.1139/x00-057
  66. 66
    Gonzalez, A.; Lynch, J. P. Effects of manganese toxicity on leaf CO2 assimilation of contrasting common bean genotypes. Physiol. Plant. 1997, 101 (4), 872880,  DOI: 10.1034/j.1399-3054.1997.1010427.x
  67. 67
    Fernando, D. R.; Lynch, J. P. Manganese phytotoxicity: new light on an old problem. Ann. Bot. 2015, 116 (3), 313319,  DOI: 10.1093/aob/mcv111
  68. 68
    Tian, Q.; Liu, N.; Bai, W.; Li, L.; Chen, J.; Reich, P. B.; Yu, Q.; Guo, D.; Smith, M. D.; Knapp, A. K.; Cheng, W.; Lu, P.; Gao, Y.; Yang, A.; Wang, T.; Li, X.; Wang, Z.; Ma, Y.; Han, X.; Zhang, W.-H. A novel soil manganese mechanism drives plant species loss with increased nitrogen deposition in a temperate steppe. Ecology 2016, 97 (1), 6574,  DOI: 10.1890/15-0917.1
  69. 69
    Lovley, D. R. Dissimilatory Fe (III) and Mn (IV) reduction. Microbiol. Rev. 1991, 55 (2), 259287,  DOI: 10.1128/mr.55.2.259-287.1991
  70. 70
    Lovley, D. R.; Phillips, E. J. Novel mode of microbial energy metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron or manganese. Appl. Environ. Microbiol. 1988, 54 (6), 14721480,  DOI: 10.1128/aem.54.6.1472-1480.1988
  71. 71
    Lovley, D. R.; Chapelle, F. H. Deep subsurface microbial processes. Rev. Geophys. 1995, 33 (3), 365381,  DOI: 10.1029/95RG01305
  72. 72
    Warrinnier, R.; Bossuyt, S.; Resseguier, C.; Cambier, P.; Houot, S.; Gustafsson, J. P.; Diels, J.; Smolders, E. Anaerobic Respiration in the Unsaturated Zone of Agricultural Soil Mobilizes Phosphorus and Manganese. Environ. Sci. Technol. 2020, 54 (8), 49224931,  DOI: 10.1021/acs.est.9b06978
  73. 73
    Grybos, M.; Davranche, M.; Gruau, G.; Petitjean, P.; Pédrot, M. Increasing pH drives organic matter solubilization from wetland soils under reducing conditions. Geoderma 2009, 154 (1–2), 1319,  DOI: 10.1016/j.geoderma.2009.09.001
  74. 74
    Li, H.; Yelle, D. J.; Li, C.; Yang, M.; Ke, J.; Zhang, R.; Liu, Y.; Zhu, N.; Liang, S.; Mo, X.; Ralph, J.; Currie, C. R.; Mo, J. Lignocellulose pretreatment in a fungus-cultivating termite. Proc. Natl. Acad. Sci. U. S. A. 2017, 114 (18), 47094714,  DOI: 10.1073/pnas.1618360114
  75. 75
    Poulsen, M.; Hu, H.; Li, C.; Chen, Z.; Xu, L.; Otani, S.; Nygaard, S.; Nobre, T.; Klaubauf, S.; Schindler, P. M.; Hauser, F.; Pan, H.; Yang, Z.; Sonnenberg, A. S.; de Beer, Z. W.; Zhang, Y.; Wingfield, M. J.; Grimmelikhuijzen, C. J.; de Vries, R. P.; Korb, J.; Aanen, D. K.; Wang, J.; Boomsma, J. J.; Zhang, G. Complementary symbiont contributions to plant decomposition in a fungus-farming termite. Proc. Natl. Acad. Sci. U. S. A. 2014, 111 (40), 145005,  DOI: 10.1073/pnas.1319718111
  76. 76
    Hofrichter, M. Review: lignin conversion by manganese peroxidase (MnP). Enzyme Microb. Technol. 2002, 30 (4), 454466,  DOI: 10.1016/S0141-0229(01)00528-2
  77. 77
    Verspoor, R. L.; Soglo, M.; Adeoti, R.; Djouaka, R.; Edwards, S.; Fristedt, R.; Langton, M.; Moriana, R.; Osborne, M.; Parr, C. L.; Powell, K.; Hurst, G. D. D.; Landberg, R. Mineral analysis reveals extreme manganese concentrations in wild harvested and commercially available edible termites. Sci. Rep. 2020, 10 (1), 6146,  DOI: 10.1038/s41598-020-63157-7
  78. 78
    Chao, T. T. Selective Dissolution of Manganese Oxides from Soils and Sediments with Acidified Hydroxylamine Hydrochloride. Soil Sci. Soc. Am. J. 1972, 36, 764768,  DOI: 10.2136/sssaj1972.03615995003600050024x
  79. 79
    Herndon, E.; Havig, J. R.; Singer, D.; McCormick, M.; Kump, L. Manganese and iron geochemistry in sediments underlying the redox-stratified Fayetteville Green Lake. Geochim. Cosmochim. Acta 2018, 231, 5063,  DOI: 10.1016/j.gca.2018.04.013
  80. 80
    Ling, F. T.; Post, J. E.; Heaney, P. J.; Santelli, C. M.; Ilton, E. S.; Burgos, W.; Rose, A. W. A multi-method characterization of natural terrestrial birnessites. Am. Mineral. 2020, 105, 833847,  DOI: 10.2138/am-2020-7303
  81. 81
    Newville, M. Fundamentals of XAFS. Rev. Mineral. Geochem. 2004, 78 (1), 3374,  DOI: 10.2138/rmg.2014.78.2
  82. 82
    Ling, F. T.; Post, J. E.; Heaney, P. J.; Kubicki, J. D.; Santelli, C. M. Fourier-transform infrared spectroscopy (FTIR) analysis of triclinic and hexagonal birnessites. Spectrochim. Acta, Part A 2017, 178, 3246,  DOI: 10.1016/j.saa.2017.01.032
  83. 83
    Post, J. E.; McKeown, D. A.; Heaney, P. J. Raman spectroscopy study of manganese oxides: Tunnel structures. Am. Mineral. 2020, 105 (8), 11751190,  DOI: 10.2138/am-2020-7390
  84. 84
    Bernardini, S.; Bellatreccia, F.; Casanova Municchia, A.; Della Ventura, G.; Sodo, A. Raman spectra of natural manganese oxides. J. Raman Spectrosc. 2019, 50 (6), 873888,  DOI: 10.1002/jrs.5583
  85. 85
    Ilton, E. S.; Post, J. E.; Heaney, P. J.; Ling, F. T.; Kerisit, S. N. XPS determination of Mn oxidation states in Mn (hydr)oxides. Appl. Surf. Sci. 2016, 366, 475485,  DOI: 10.1016/j.apsusc.2015.12.159
  86. 86
    Sutherland, K. M.; Wankel, S. D.; Hansel, C. M. Oxygen isotope analysis of bacterial and fungal manganese oxidation. Geobiology 2018, 16 (4), 399411,  DOI: 10.1111/gbi.12288
  87. 87
    Sutherland, K. M.; Wostbrock, J. A. G.; Hansel, C. M.; Sharp, Z. D.; Hein, J. R.; Wankel, S. D., Ferromanganese crusts as recorders of marine dissolved oxygen. Earth Planet. Sci. Lett. 2020, 533. 116057 DOI: 10.1016/j.epsl.2019.116057
  88. 88
    Lützow, M. V.; Kögel-Knabner, I.; Ekschmitt, K.; Matzner, E.; Guggenberger, G.; Marschner, B.; Flessa, H. Stabilization of organic matter in temperate soils: Mechanisms and their relevance under different soil conditions—a review. Eur. J. Soil Sci. 2006, 57, 426445,  DOI: 10.1111/j.1365-2389.2006.00809.x
  89. 89
    Porras, R. C.; Hicks Pries, C. E.; Tom, M. S.; Nico, P. S. Synthetic iron (hydr)oxide-glucose associations in subsurface soil: Effects on decomposability of mineral associated carbon. Sci. Total Environ. 2018, 613–614, 342351,  DOI: 10.1016/j.scitotenv.2017.08.290
  90. 90
    Gu, B.; Schmitt, J.; Chen, Z.; Liang, L.; McCarthy, J. F. Adsorption and desorption of natural organic matter on iron oxide: Mechanisms and models. Environ. Sci. Technol. 1994, 28, 3846,  DOI: 10.1021/es00050a007
  91. 91
    Estes, E. R.; Andeer, P. F.; Nordlund, D.; Wankel, S. D.; Hansel, C. M. Biogenic manganese oxides as reservoirs of organic carbon and proteins in terrestrial and marine environments. Geobiology 2017, 15 (1), 158172,  DOI: 10.1111/gbi.12195
  92. 92
    Stuckey, J. W.; Goodwin, C.; Wang, J.; Kaplan, L. A.; Vidal-Esquivel, P.; Beebe, T. P., Jr.; Sparks, D. L. Impacts of hydrous manganese oxide on the retention and lability of dissolved organic matter. Geochem. Trans. 2018, 19, 6,  DOI: 10.1186/s12932-018-0051-x
  93. 93
    Rennert, T.; Handel, M.; Hoschen, C.; Lugmeier, J.; Steffens, M.; Totsche, K. U. A NanoSIMS study on the distribution of soil organic matter, iron and manganese in a nodule from a Stagnosol. Eur. J. Soil Sci. 2014, 65, 684692,  DOI: 10.1111/ejss.12157
  94. 94
    Johnson, K.; Purvis, G.; Lopez-Capel, E.; Peacock, C.; Gray, N.; Wagner, T.; Marz, C.; Bowen, L.; Ojeda, J.; Finlay, N.; Robertson, S.; Worrall, F.; Greenwell, C. Towards a mechanistic understanding of carbon stabilization in manganese oxides. Nat. Commun. 2015, 6, 7628,  DOI: 10.1038/ncomms8628
  95. 95
    Sparks, D. L. Environmental Soil Chemistry, 2nd ed.; Academic Press: San Diego, CA, 2002.
  96. 96
    Allard, S.; Gutierrez, L.; Fontaine, C.; Croué, J.; Gallard, H. Organic matter interactions with natural manganese oxide and synthetic birnessite. Sci. Total Environ. 2017, 583, 487495,  DOI: 10.1016/j.scitotenv.2017.01.120
  97. 97
    Zhao, W.; Cheng, H.; Tao, S. Structure-reactivity relationships in the adsorption and degradation of substituted phenylarsonic acids on birnessite. Environ. Sci. Technol. 2020, 54, 14751483,  DOI: 10.1021/acs.est.9b04203
  98. 98
    Wang, Q.; Yang, P.; Zhu, M. Structural transformation of birnessite by fulvic acid under anoxic conditions. Environ. Sci. Technol. 2018, 52, 18441853,  DOI: 10.1021/acs.est.7b04379
  99. 99
    Joshi, T. P.; Zhang, G.; H, C.; Liu, R.; Liu, H.; Qu, J. Transformation of para arsanilic acid by manganese oxide: Adsorption, oxidation, and influencing factors. Water Res. 2017, 116, 126134,  DOI: 10.1016/j.watres.2017.03.028
  100. 100
    Rabe, M.; Verdes, D.; Seeger, S. Understanding protein adsorption phenomena at solid surfaces. Adv. Colloid Interface Sci. 2011, 162, 87106,  DOI: 10.1016/j.cis.2010.12.007
  101. 101
    Keil, R. G.; Mayer, L. M. Mineral matrices and organic matter. Treatise Geochem. 2014, 12, 337359,  DOI: 10.1016/B978-0-08-095975-7.01024-X
  102. 102
    Sheng, X.; Qin, C.; Yang, B.; Hu, X.; Liu, C.; Waigi, M. G.; Li, X.; Ling, W. Metal cation saturation on montmorillonites facilitates the adsorption of DNA via cation bridging. Chemosphere 2019, 235, 670678,  DOI: 10.1016/j.chemosphere.2019.06.159
  103. 103
    Cheng, H.; Yang, T.; Jiang, J.; Li, X.; Wang, P.; Ma, J. Mn2+ effect on manganese oxides (MnOx) nanoparticles aggregation in solution: Chemical adsorption and cation bridging. Environ. Pollut. 2020, 267, 115561,  DOI: 10.1016/j.envpol.2020.115561
  104. 104
    Franchi, M.; Ferris, J. P.; Callori, E. Cations as mediators of the adsorption of nucleic acids on clay surfaces in prebiotic environments. Origins Life Evol. Biospheres 2003, 33, 116,  DOI: 10.1023/A:1023982008714
  105. 105
    Klewicki, J. K.; Morgan, J. J. Dissolution of β-MnOOH particles by ligands: Pyrophosphate, ethylenediaminetetraacetate, and citrate. Geochim. Cosmochim. Acta 1999, 63, 30173024,  DOI: 10.1016/S0016-7037(99)00229-X
  106. 106
    Chorover, J.; Amistadi, M. K. Reaction of forest floor organic matter at goethite, birnessite and smectite surfaces. Geochim. Cosmochim. Acta 2001, 65 (1), 95109,  DOI: 10.1016/S0016-7037(00)00511-1
  107. 107
    Tamrat, W. Z.; Rose, J.; Grauby, O.; Doelsch, E.; Levard, C.; Chaurand, P.; Basile-Doelsch, I. Soil organo-mineral associations formed by co-precipitation of Fe, Si and Al in presence of organic ligands. Geochim. Cosmochim. Acta 2019, 260, 1528,  DOI: 10.1016/j.gca.2019.05.043
  108. 108
    Chen, C.; Dynes, J. J.; Wang, J.; Sparks, D. L. Properties of Fe-organic matter associations via coprecipitation versus adsorption. Environ. Sci. Technol. 2014, 48, 1375113759,  DOI: 10.1021/es503669u
  109. 109
    Mikutta, R.; Zang, U.; Chorover, J.; Haumaier, L.; Kalbitz, K. Stabilization of extracellular polymeric substances (bacillus subtilis) by adsorption to and coprecipitation with Al forms. Geochim. Cosmochim. Acta 2011, 75, 31353154,  DOI: 10.1016/j.gca.2011.03.006
  110. 110
    Eusterhues, K.; Rennert, T.; Knicker, H.; Kogel-Knabner, I.; Totsche, K. U.; Schwertmann, U. Fractionation of organic matter due to reaction with ferrihydrite: Coprecipitation versus adsorption. Environ. Sci. Technol. 2011, 45, 527533,  DOI: 10.1021/es1023898
  111. 111
    Totsche, K. U.; Amelung, W.; Gerzabek, M. H.; Guggenberger, G.; Klumpp, E.; Knief, C.; lehndorff, E.; Mikutta, R.; Peth, S.; Prechtel, A.; Ray, N.; Kögel-Knabner, I. Microaggregates in soils. J. Plant Nutr. Soil Sci. 2018, 181, 104136,  DOI: 10.1002/jpln.201600451
  112. 112
    Toner, B.; Fakra, S.; Villalobos, M.; Warwick, T.; Sposito, G. Spatially resolved characterization of biogenic manganese oxide production within a bacterial biofilm. Appl. Environ. Microbiol. 2005, 71 (3), 13001310,  DOI: 10.1128/AEM.71.3.1300-1310.2005
  113. 113
    Yang, W.; Zhang, Z.; Zhang, Z.; Chen, H.; Liu, J.; Ali, M.; Liu, F.; Li, L. Population structure of manganese-oxidizing macteria in stratified soils and properties of manganese oxide aggregates under manganese-complex medium enrichment. PLoS One 2013, 8, e73778  DOI: 10.1371/journal.pone.0073778
  114. 114
    Piccolo, A.; Spaccini, R.; Nebbioso, A.; Mazzei, P. Carbon sequestration in soil by in situ catalyzed photo-oxidative polymerization of soil organic matter. Environ. Sci. Technol. 2011, 45, 66976702,  DOI: 10.1021/es201572f
  115. 115
    Shindo, H.; Huang, P. M. Role of Mn(IV) oxide in abiotic formation of humic substances in the environment. Nature 1982, 298, 363365,  DOI: 10.1038/298363a0
  116. 116
    Liu, M. M.; Cao, X. H.; Tan, W. F.; Feng, X. H.; Qiu, G. H.; Chen, X. H.; Liu, F. Structural controls on the catalytic polymerization of hydroquinone by birnessite. Clays Clay Miner. 2011, 59, 525537,  DOI: 10.1346/CCMN.2011.0590510
  117. 117
    Xia, X.; Stone, A. T. Mandelic acid and phenyllactic acid “Reaction Sets” for exploring the kinetics and mechanism of oxidations by hydrous manganese oxide (HMO). Environ. Sci. Process. Impacts 2019, 21, 10381051,  DOI: 10.1039/C9EM00128J
  118. 118
    Li, H.; Liu, F.; Zhu, M.; Feng, X.; Zhang, J.; Yin, H. Structure and properties of Co-doped cryptomelane and its enhanced removal of Pb2+ and Cr3+ from wastewater. J. Environ. Sci. 2015, 34, 7785,  DOI: 10.1016/j.jes.2015.02.006
  119. 119
    Sunda, W. G.; Kieber, D. J. Oxidation of humic substances by manganese oxides yields low-molecular-weight organic substrates. Nature 1994, 367, 6264,  DOI: 10.1038/367062a0
  120. 120
    Wang, Y.; Stone, A. T. The citric acid-MnIII,IVO2 (birnessite) reaction. Electron transfer, complex formation, and autocatalytic feedback. Geochim. Cosmochim. Acta 2006, 70 (17), 44634476,  DOI: 10.1016/j.gca.2006.06.1551
  121. 121
    Ramstedt, M.; Sjöberg, S. Phase transformation and proton promoted dissolution of hydrous Manganite (γ-MnOOH). Aquat. Geochem. 2005, 11, 413431,  DOI: 10.1007/s10498-005-7441-2
  122. 122
    Majcher, E. H.; Chorover, J.; Bollag, J.; Huang, P. M. Evolution of CO2 during birnessite-induced oxidation of 14C-labeled catechol. Soil Sci. Soc. Am. J. 2000, 64, 157163,  DOI: 10.2136/sssaj2000.641157x
  123. 123
    Harrington, J. M.; Parker, D. L.; Bargar, J. R.; Jarzecki, A. A.; Tebo, B. M.; Sposito, G.; Duckworth, O. W. Structural dependence of Mn complexation by siderophores: Donor group dependence on complex stability and reactivity. Geochim. Cosmochim. Acta 2012, 88, 106119,  DOI: 10.1016/j.gca.2012.04.006
  124. 124
    Akafia, M. M.; Harrington, J. M.; Bargar, J. R.; Duckworth, O. W. Metal oxyhydroxide dissolution as promoted by structurally diverse siderophores and oxalate. Geochim. Cosmochim. Acta 2014, 141, 258269,  DOI: 10.1016/j.gca.2014.06.024
  125. 125
    Duckworth, O. W.; Sposito, G. Siderophore-Manganese(III) Interactions II. Manganite Dissolution Promoted by Desferrioxamine B. Environ. Sci. Technol. 2005, 39, 60456051,  DOI: 10.1021/es050276c
  126. 126
    Duckworth, O. W.; Sposito, G. Siderophore-promoted dissolution of synthetic and biogenic layer-type Mn oxides. Chem. Geol. 2007, 242 (3–4), 497508,  DOI: 10.1016/j.chemgeo.2007.05.007
  127. 127
    Peña, J.; Duckworth, O. W.; Bargar, J. R.; Sposito, G. Dissolution of hausmannite (Mn3O4) in the presence of the trihydroxamate siderophore desferrioxamine B. Geochim. Cosmochim. Acta 2007, 71 (23), 56615671,  DOI: 10.1016/j.gca.2007.03.043
  128. 128
    Saal, L. B.; Duckworth, O. W. Synergistic Dissolution of Manganese Oxides as Promoted by Siderophores and Small Organic Acids. Soil Sci. Soc. Am. J. 2010, 74 (6), 20212031,  DOI: 10.2136/sssaj2009.0465
  129. 129
    Berg, B.; Steffen, K. T.; McClaugherty, C. Litter decomposition rate is dependent on litter Mn concentrations. Biogeochemistry 2007, 82 (1), 2939,  DOI: 10.1007/s10533-006-9050-6
  130. 130
    Whalen, E. D.; Smith, R. G.; Grandy, A. S.; Frey, S. D. Manganese limitation as a mechanism for reduced decomposition in soils under atmospheric nitrogen deposition. Soil Biol. Biochem. 2018, 127, 252263,  DOI: 10.1016/j.soilbio.2018.09.025
  131. 131
    Stendahl, J.; Berg, B.; Lindahl, B. D. Manganese availability is negatively associated with carbon storage in northern coniferous forest humus layers. Sci. Rep. 2017, 7 (1), 15487,  DOI: 10.1038/s41598-017-15801-y
  132. 132
    Hou, S. L.; Hattenschwiler, S.; Yang, J. J.; Sistla, S.; Wei, H. W.; Zhang, Z. W.; Hu, Y. Y.; Wang, R. Z.; Cui, S. Y.; Lu, X. T.; Han, X. G. Increasing rates of long-term nitrogen deposition consistently increased litter decomposition in a semi-arid grassland. New Phytol. 2021, 229 (1), 296307,  DOI: 10.1111/nph.16854
  133. 133
    Kranabetter, J. M. Increasing soil carbon content with declining soil manganese in temperate rainforests: is there a link to fungal Mn?. Soil Biol. Biochem. 2019, 128, 179181,  DOI: 10.1016/j.soilbio.2018.11.001
  134. 134
    Kellner, H.; Luis, P.; Pecyna, M. J.; Barbi, F.; Kapturska, D.; Krüger, D.; Zak, D. R.; Marmeisse, R.; Vandenbol, M.; Hofrichter, M. Widespread occurrence of expressed fungal secretory peroxidases in forest soils. PLoS One 2014, 9, e95557  DOI: 10.1371/journal.pone.0095557
  135. 135
    Hatakka, A.; Lundell, T.; Hofrichter, M.; Maijala, P. Manganese peroxidase and its role in the degradation of wood lignin. ACS Symp. Ser. 2003, 855, 230243,  DOI: 10.1021/bk-2003-0855.ch014
  136. 136
    Sun, T.; Cui, Y.; Berg, B.; Zhang, Q.; Dong, L.; Wu, Z.; Zhang, L. A test of manganese effects on decomposition in forest and cropland sites. Soil Biol. Biochem. 2019, 129, 178183,  DOI: 10.1016/j.soilbio.2018.11.018
  137. 137
    Collins, P. J.; Dobson, A. D. Oxidation of fluorene and phenanthrene by Mn (II) dependent peroxidase activity in whole cultures of Trametes (Coriolus) versicolor. Biotechnol. Lett. 1996, 18 (7), 801804,  DOI: 10.1007/BF00127892
  138. 138
    Steffen, K. T.; Hatakka, A.; Hofrichter, M. Degradation of humic acids by the litter-decomposing basidiomycete Collybia dryophila. Appl. Environ. Microbiol. 2002, 68 (7), 34423448,  DOI: 10.1128/AEM.68.7.3442-3448.2002
  139. 139
    Kadri, T.; Rouissi, T.; Brar, S. K.; Cledon, M.; Sarma, S.; Verma, M. Biodegradation of polycyclic aromatic hydrocarbons (PAHs) by fungal enzymes: A review. J. Environ. Sci. 2017, 51, 5274,  DOI: 10.1016/j.jes.2016.08.023
  140. 140
    Bao, W.; Fukushima, Y.; Jensen Jr, K. A.; Moen, M. A.; Hammel, K. E. Oxidative degradation of non-phenolic lignin during lipid peroxidation by fungal manganese peroxidase. FEBS Lett. 1994, 354 (3), 297300,  DOI: 10.1016/0014-5793(94)01146-X
  141. 141
    Kapich, A. N.; Korneichik, T. V.; Hatakka, A.; Hammel, K. E. Oxidizability of unsaturated fatty acids and of a non-phenolic lignin structure in the manganese peroxidase-dependent lipid peroxidation system. Enzyme Microb. Technol. 2010, 46 (2), 136140,  DOI: 10.1016/j.enzmictec.2009.09.014
  142. 142
    Kapich, A.; Hofrichter, M.; Vares, T.; Hatakka, A. Coupling of manganese peroxidase-mediated lipid peroxidation with destruction of nonphenolic lignin model compounds and 14C-labeled lignins. Biochem. Biophys. Res. Commun. 1999, 259 (1), 212219,  DOI: 10.1006/bbrc.1999.0742
  143. 143
    Trum, F.; Titeux, H.; Cornelis, J.-T.; Delvaux, B. Effects of manganese addition on carbon release from forest floor horizons. Can. J. For. Res. 2011, 41 (3), 643648,  DOI: 10.1139/X10-224
  144. 144
    Subedi, P.; Jokela, E. J.; Vogel, J. G.; Bracho, R.; Inglett, K. S., The effects of nutrient limitations on microbial respiration and organic matter decomposition in a Florida Spodosol as influenced by historical forest management practices. For. Ecol. Manage. , 2021, 479, 118592. DOI: 10.1016/j.foreco.2020.118592
  145. 145
    Trum, F.; Titeux, H.; Ponette, Q.; Berg, B. Influence of manganese on decomposition of common beech (Fagus sylvatica L.) leaf litter during field incubation. Biogeochemistry 2015, 125 (3), 349358,  DOI: 10.1007/s10533-015-0129-9
  146. 146
    Nowack, B.; Stone, A. T. Manganese-catalyzed degradation of phosphonic acids. Environ. Chem. Lett. 2003, 1 (1), 2431,  DOI: 10.1007/s10311-002-0014-3
  147. 147
    Duckworth, O. W.; Sposito, G. Siderophore-Manganese(III) Interactions. I. Air-Oxidation of Manganese(II) Promoted by Desferrioxamine B. Environ. Sci. Technol. 2005, 39, 60376044,  DOI: 10.1021/es050275k
  148. 148
    Harrington, J. M.; Bargar, J. R.; Jarzecki, A. A.; Roberts, J. G.; Sombers, L. A.; Duckworth, O. W. Trace metal complexation by the triscatecholate siderophore protochelin: structure and stability. BioMetals 2012, 25 (2), 393412,  DOI: 10.1007/s10534-011-9513-7
  149. 149
    Wang, Y.; Stone, A. T. Phosphonate- and Carboxylate-Based Chelating Agents that Solubilize (Hydr)oxide-Bound MnIII. Environ. Sci. Technol. 2008, 42, 43974403,  DOI: 10.1021/es7032668
  150. 150
    Madison, A. S.; Tebo, B. M.; Luther, G. W., 3rd Simultaneous determination of soluble manganese(III), manganese(II) and total manganese in natural (pore)waters. Talanta 2011, 84 (2), 37481,  DOI: 10.1016/j.talanta.2011.01.025
  151. 151
    Madison, A. S.; Tebo, B. M.; Mucci, A.; Sundby, B.; Luther, G. W. Abundant porewater Mn (III) is a major component of the sedimentary redox system. Science 2013, 341 (6148), 875878,  DOI: 10.1126/science.1241396
  152. 152
    Li, Q.; Xie, L.; Jiang, Y.; Fortner, J. D.; Yu, K.; Liao, P.; Liu, C. Formation and stability of NOM-Mn(III) colloids in aquatic environments. Water Res. 2019, 149, 190201,  DOI: 10.1016/j.watres.2018.10.094
  153. 153
    Hu, E.; Zhang, Y.; Wu, S.; Wu, J.; Liang, L.; He, F. Role of dissolved Mn(III) in transformation of organic contaminants: Non-oxidative versus oxidative mechanisms. Water Res. 2017, 111, 234243,  DOI: 10.1016/j.watres.2017.01.013
  154. 154
    Jones, M. E.; Nico, P. S.; Ying, S.; Regier, T.; Thieme, J.; Keiluweit, M. Manganese-Driven Carbon Oxidation at Oxic-Anoxic Interfaces. Environ. Sci. Technol. 2018, 52 (21), 1234912357,  DOI: 10.1021/acs.est.8b03791
  155. 155
    Jones, M. E.; LaCroix, R. E.; Zeigler, J.; Ying, S. C.; Nico, P. S.; Keiluweit, M. Enzymes, Manganese, or Iron? Drivers of Oxidative Organic Matter Decomposition in Soils. Environ. Sci. Technol. 2020, 54 (21), 1411414123,  DOI: 10.1021/acs.est.0c04212
  156. 156
    Karolewski, J. S.; Sutherland, K. M.; Hansel, C. M.; Wankel, S. D. An isotopic study of abiotic nitrite oxidation by ligand-bound manganese (III). Geochim. Cosmochim. Acta 2021, 293, 365378,  DOI: 10.1016/j.gca.2020.11.004
  157. 157
    Birkner, N.; Navrotsky, A. Thermodynamics of manganese oxides: Sodium, potassium, and calcium birnessite and cryptomelane. Proc. Natl. Acad. Sci. U. S. A. 2017, 114, E1046E1053,  DOI: 10.1073/pnas.1620427114
  158. 158
    Hem, J. D.; Lind, C. J. Nonequilibrium models for predicting forms of precipitated manganese oxides. Geochim. Cosmochim. Acta 1983, 47, 20372046,  DOI: 10.1016/0016-7037(83)90219-3
  159. 159
    Shaughnessy, D. A.; Nitsche, H.; Booth, C. H.; Shuh, D. K.; Waychunas, G. A.; Wilson, R. E.; Gill, H.; Cantrell, K. J.; Serne, R. J. Molecular interfacial reactions between Pu(VI) and manganese oxide minerals Manganite and hausmannite. Environ. Sci. Technol. 2003, 37, 33673374,  DOI: 10.1021/es025989z
  160. 160
    Tebo, B. M.; Bargar, J. R.; Clement, B. G.; Dick, G. J.; Murray, K. J.; Parker, D.; Verity, R.; Webb, S. M. Biogenic manganese oxides: Properties and mechanisms of formation. Annu. Rev. Earth Planet. Sci. 2004, 32, 287328,  DOI: 10.1146/
  161. 161
    Yin, H.; Tan, W.; Zhang, L.; Cui, H.; Qiu, G.; Liu, F.; Feng, X. Characterization of Ni-rich hexagonal birnessite and its geochemical effects on aqueous Pb2+/Zn2+ and As(III). Geochim. Cosmochim. Acta 2012, 93, 4762,  DOI: 10.1016/j.gca.2012.05.039
  162. 162
    Li, H.; Bolscher, T.; Winnick, M.; Tfaily, M. M.; Cardon, Z. G.; Keiluweit, M. Simple Plant and Microbial Exudates Destabilize Mineral-Associated Organic Matter via Multiple Pathways. Environ. Sci. Technol. 2021, 55 (5), 33893398,  DOI: 10.1021/acs.est.0c04592
  163. 163
    Chen, C.; Hall, S. J.; Coward, E.; Thompson, A. Iron-mediated organic matter decomposition in humid soils can counteract protection. Nat. Commun. 2020, 11 (1), 2255,  DOI: 10.1038/s41467-020-16071-5
  164. 164
    Wang, Q.; Yang, P.; Zhu, M. Effects of metal cations on coupled birnessite structural transformation and natural organic matter adsorption and oxidation. Geochim. Cosmochim. Acta 2019, 250, 292310,  DOI: 10.1016/j.gca.2019.01.035
  165. 165
    Villalobos, M.; Bargar, J.; Sposito, G. Mechanisms of Pb(II) sorption on a biogenic manganese oxide. Environ. Sci. Technol. 2005, 39, 569576,  DOI: 10.1021/es049434a
  166. 166
    Villalobos, M.; Escobar-Quiroz, I. N.; Salazar-Camacho, C. The influence of particle size and structure on the sorption and oxidation behavior of birnessite: I. Adsorption of As(V) and oxidation of As(III). Geochim. Cosmochim. Acta 2014, 125, 564581,  DOI: 10.1016/j.gca.2013.10.029
  167. 167
    Zhu, M.; Paul, K. W.; Kubicki, J. D.; Sparks, D. L. Quantum chemical study of arsenic(III, V) adsorption on Mn-oxides: Implications for arsenic(III) oxidation. Environ. Sci. Technol. 2009, 43, 66556661,  DOI: 10.1021/es900537e
  168. 168
    Zhang, S.; Chen, S.; Liu, F.; Li, J.; Liang, X.; Chu, S.; Xiang, Q.; Huang, C.; Yin, H. Effects of Mn average oxidation state on the oxidation behaviors of As(III) and Cr(III) by vernadite. Appl. Geochem. 2018, 94, 3545,  DOI: 10.1016/j.apgeochem.2018.05.002
  169. 169
    Liu, J.; Zhang, Y.; Gu, Q.; Sheng, A.; Zhang, B., Tunable Mn Oxidation State and Redox Potential of Birnessite Coexisting with Aqueous Mn(II) in Mildly Acidic Environments. Minerals 2020, 10, 690 DOI: 10.3390/min10080690
  170. 170
    Tang, Y.; Webb, S. M.; Estes, E. R.; Hansel, C. M. Chromium(III) oxidation by biogenic manganese oxides with varying structural ripening. Environ. Sci. Process Impacts 2014, 16, 21272136,  DOI: 10.1039/C4EM00077C
  171. 171
    Huang, J. Z. S.; Dai, Y.; Liu, C.-C.; Zhang, H. Effect of MnO2 phase structure on the oxidative reactivity toward bisphenol A degradation. Environ. Sci. Technol. 2018, 52, 1130911318,  DOI: 10.1021/acs.est.8b03383
  172. 172
    Zhang, Y. Y. Y.; Zhang, Y.; Zhang, T.; Ye, M. Heterogeneous oxidation of naproxen in the presence of α-MnO2 nanostructures with different morphologies. Appl. Catal., B 2012, 127, 182189,  DOI: 10.1016/j.apcatb.2012.08.014
  173. 173
    Balgooyen, S.; Alaimo, P. J.; Remucal, C. K.; Ginder-Vogel, M. Structural Transformation of MnO2 during the Oxidation of Bisphenol A. Environ. Sci. Technol. 2017, 51 (11), 60536062,  DOI: 10.1021/acs.est.6b05904
  174. 174
    Gulley-Stahl, H.; Hogan II, P. A.; Schmidt, W. L.; Wall, S. J.; Buhrlage, A.; Bullen, H. A. Surface complexation of catechol to metal oxides: An ATR-FTIR, adsorption. and dissolution study. Environ. Sci. Technol. 2010, 44, 41164121,  DOI: 10.1021/es902040u
  175. 175
    Johnson, S. B.; Yoon, T. H.; Kocar, B. D.; Brown, G. E., Jr. Adsorption of organic matter at mineral/water interfaces. 2. Outersphere adsorption of maleate and implications for dissolution processes. Langmuir 2004, 20, 49965006,  DOI: 10.1021/la036288y
  176. 176
    Gao, J.; Hedman, C.; Liu, C.; Guo, T.; Pedersen, J. A. Transformation of sulfamethazine by manganese oxide in aqueous solution. Environ. Sci. Technol. 2012, 46, 26422651,  DOI: 10.1021/es202492h
  177. 177
    McBride, M. B. Adsorption and Oxidation of Phenolic Compounds by Iron and Manganese Oxides. Soil Sci. Soc. Am. J. 1987, 51 (6), 14661472,  DOI: 10.2136/sssaj1987.03615995005100060012x
  178. 178
    Nowack, B.; Stone, A. T. Homogeneous and heterogeneous oxidation of nitrilotrismethylenephosphonic acid (NTMP) in the presence of manganese(II, III) and molecular oxygen. J. Phys. Chem. B 2002, 106, 62276233,  DOI: 10.1021/jp014293+
  179. 179
    Hinkle, M. A. G.; Flynn, E. D.; Catalano, J. G. Structural response of phyllomanganates to wet aging and aqueous Mn(II). Geochim. Cosmochim. Acta 2016, 192, 220234,  DOI: 10.1016/j.gca.2016.07.035
  180. 180
    Yang, P.; Post, J. E.; Wang, Q.; Xu, W.; Geiss, R.; McCurdy, P. R.; Zhu, M. Metal adsorption controls stability of layered manganese oxides. Environ. Sci. Technol. 2019, 53, 74537462,  DOI: 10.1021/acs.est.9b01242
  181. 181
    Eusterhues, K.; Rumpel, C.; Kleber, M.; Kögel-Knabner, I. Stabilisation of soil organic matter by interactions with minerals as revealed by mineral dissolution and oxidative degradation. Org. Geochem. 2003, 34 (12), 15911600,  DOI: 10.1016/j.orggeochem.2003.08.007
  182. 182
    Mikutta, R.; Kleber, M.; Torn, M. S.; Jahn, R. Stabilization of Soil Organic Matter: Association with Minerals or Chemical Recalcitrance?. Biogeochemistry 2006, 77 (1), 2556,  DOI: 10.1007/s10533-005-0712-6
  183. 183
    Kögel-Knabner, I.; Guggenberger, G.; Kleber, M.; Kandeler, E.; Kalbitz, K.; Scheu, S.; Eusterhues, K.; Leinweber, P. Organo-mineral associations in temperate soils: Integrating biology, mineralogy, and organic matter chemistry. J. Plant Nutr. Soil Sci. 2008, 171 (1), 6182,  DOI: 10.1002/jpln.200700048
  184. 184
    Rasmussen, C.; Heckman, K.; Wieder, W. R.; Keiluweit, M.; Lawrence, C. R.; Berhe, A. A.; Blankinship, J. C.; Crow, S. E.; Druhan, J. L.; Hicks Pries, C. E.; Marin-Spiotta, E.; Plante, A. F.; Schädel, C.; Schimel, J. P.; Sierra, C. A.; Thompson, A.; Wagai, R. Beyond clay: towards an improved set of variables for predicting soil organic matter content. Biogeochemistry 2018, 137 (3), 297306,  DOI: 10.1007/s10533-018-0424-3
  185. 185
    Stone, A. T. Microbial metabolites and the reductive dissolution of manganese oxides: Oxalate and pyruvate. Geochim. Cosmochim. Acta 1987, 51, 919925,  DOI: 10.1016/0016-7037(87)90105-0
  186. 186
    Stone, A. T.; Ulrich, H. Kinetics and Reaction Stoichiometry in the Reductive Dissolution of Manganese(IV) Dioxide and Co(Ill) Oxide by Hydroquinone. J. Colloid Interface Sci. 1989, 132 (2), 509522,  DOI: 10.1016/0021-9797(89)90265-8
  187. 187
    Stone, A. T. Oxidation and hydrolysis of ionizable organic pollutants at hydrous metal oxide surfaces. Rates of Soil Chemical Processes 2015, 27, 231254,  DOI: 10.2136/sssaspecpub27.c9
  188. 188
    Heckman, K.; Lawrence, C. R.; Harden, J. W. A sequential selective dissolution method to quantify storage and stability of organic carbon associated with Al and Fe hydroxide phases. Geoderma 2018, 312, 2435,  DOI: 10.1016/j.geoderma.2017.09.043
  189. 189
    Zhao, Q.; Poulson, S. R.; Obrist, D.; Sumaila, S.; Dynes, J. J.; McBeth, J. M.; Yang, Y. Iron-bound organic carbon in forest soils: quantification and characterization. Biogeosciences 2016, 13 (16), 47774788,  DOI: 10.5194/bg-13-4777-2016