Clément J-C, Shrestha J, Ehrenfeld JG, Jaffe PR. 2005. Ammonium oxidation coupled to dissimilatory reduction of iron under anaerobic conditions in wetland soils. Soil Biology and Biochemistry 37: 2323-2328.
These authors observed an unexpected chemical reaction involving the accumulation of both nitrite (NO2-) and ferrous iron (Fe(II)) under anaerobic conditions. They investigated this phenomenon further, and propose a chemical reaction in wet soils in which ammonium is oxidized under reducing conditions by transferring electrons to Fe(III), generating NO2- and Fe(II).
Nitrite does not usually accumulate in soils. These authors suggest that under normal conditions, it is consumed at least as fast as it is produced, but their experimental conditions included inhibition of denitrification, allowing nitrite to build up to detectable levels. Other oxidizers besides Fe(III), such as Mn(IV), were not detected in soil samples and were not included in the experiment.
The proposed chemical pathway is thermodynamically feasible at pH 7, though it appears to rely on goethite as the ferric iron source; from my understanding of dissimilatory iron reduction (e.g. Lovley 1991), I would expect strongly crystalline forms of iron oxide such as goethite to be highly resistant to such destructive forces, and the iron source in the systems (natural and experimental) described here to be amorphous ferric oxides instead. But the underlying chemistry appears plausible to me.
One unexpected aspect of this short communication was the authors’ use of a Dionex ion chromatography system, apparently very similar to the device I will be using to analyze root exudates. Additionally, this paper discusses the “ferrous wheel”, a memorable name for chemical cycling of iron between valencies, as an established hypothesis; I need to track down the origins of this term and learn its importance regarding my own attempts to relate measured Fe(III) contents to redox conditions.
Wednesday, January 13, 2010
Tuesday, January 12, 2010
Lovley 1991
Lovley DR. 1991. Dissimilatory Fe(III) and Mn(IV) reduction. Microbiological Reviews 55: 259-287.
This author comprehensively reviews microbe-mediated reduction of iron and manganese in soils and sediments, in a long and detailed paper. Dissimilatory reduction is distinct from assimilatory reduction, in which metal ions are reduced when they are incorporated into cellular macromolecules such as enzymes and cofactors. Dissimilatory reduction is a process that ends with the accumulation of reduced metal outside the cell, and is responsible for the majority of iron and manganese reduction in sediments. While it has been observed in aerobic environments, such reduction occurs mainly in anaerobic conditions. Fe(III) in particular is most often reduced when it is the final electron acceptor in the anaerobic oxidation of organic molecules.
Microbes capable of dissimilatory metal reduction can be categorized in a number of ways. This author presents 5 categories, though I think there is considerable overlap between them, as in cases where one cell is able to reduce Fe(III) and metabolize a range of carbon sources. In the first category, reducing fermenting bacteria, the amount of Fe(III) reduced during metabolism is far less than the stoichiometry of the redox couple would suggest. This implies that Fe(III) is a minor electron acceptor during fermentation reactions, rather than the electron acceptor of choice or necessity for these organisms. Sulfur-oxidizing species in contrast do appear to reduce quantities of iron in line with stoichiometric predictions, but do not seem to gain significant energy from these reactions which occur under aerobic conditions on elemental sulfur. Hydrogen-oxidizing reducers appear to be abundant in anaerobic sediments, and seem to have high affinity for hydrogen gas; where Fe(III) is being reduced, hydrogen concentrations are low, and when hydrogen is added, Fe(III) reduction increases. These organisms need other material, such as simple organic molecules, to grow, as this reaction provides energy but little else. Organic-acid oxidizing reducers and aromatic-oxidizers are probably in many cases the same cells. As these molecules are sometimes the result of fermentation metabolisms, one possible food chain in anaerobic environments is fermentation, with some Fe(III) reduction, followed by greater levels of Fe(III) reduction linked to the decomposition of smaller organic molecules such as acetate. Such an ecological pairing is probably widespread, given the known abundance and diversity of fermenting species and the probable abundance of more aggressively iron-reducing species, many of which are probably Archaea rather than Bacteria.
There are three competing models for how Fe(III) is reduced in natural environments. The first is the enzymatic model, in which microbial cells employ membrane-bound or intracellular enzymes to transfer electrons to Fe(III) ions during the process of anaerobic oxidation of organic matter. The second is termed the redox model, and posits the majority of Fe(III) reduction is driven by equilibrium thermodynamics, with the relative levels of Fe(III) and Fe(II) in sediments controlled by abiotic factors such as temperature and pH. The third model is termed the direct-reduction model, in which some organic molecules react directly with Fe(III), without the intervention of cells or enzymes.
At first glance, the observation that Fe(III) reduction rates fall when microbes are removed from sediments supports the enzymatic model, but in fact all three models rely on microbial metabolisms. In the redox model, competing electron acceptors such as oxygen and nitrate are consumed by microbes, lowering their concentrations to levels too low to influence the transitions between Fe(III) and Fe(II). And the direct model relies on microbes releasing key organic molecules known to reduce Fe(III) in vitro. However, several other lines of evidence, such as the lack of spontaneous shifts in iron valency ratios in stored sediments, the large changes in pH associated with widely-used extraction methods, and the general rarity of rapidly-acting direct-reducing molecules all lead to the conclusion that while the other mechanisms may contribute some iron reducing activity in some situations, the overwhelming majority of Fe(III) reduction to Fe(II) occurring is driven by microbes and their enzymes.
This author spends nearly as much time discussing Mn(IV) reduction as Fe(III) reduction, but I have little interest in Mn chemistry at this time. However, in the discussion of the competing models of metal reduction, mention is made that Fe(II) in solution may reduce Mn(IV), removing accumulated Fe(II) from iron reduction and abiotically returning the iron to Fe(III) while generating Mn(II). This abiotic back-reaction closes the loop on cycling Fe(III).
Iron reduction is postulated as one of the first, if not the first, globally important metabolic pathway, with early Archaea using Fe(III) as their final electron acceptor in a generally reducing environment lacking free oxygen and nitrate. If Fe(III), a non-soluble, precipitating substance is the primary oxidizing agent, the redox environment of the Earth 2 billion years ago was upside-down compared to today: the surface and atmosphere was reducing, while buried and water-saturated sediments were oxidizing.
Freshwater swamps differ from aquatic and marine sediments in a number of ways, including a generally high organic content and plenty of sulfates. Under these conditions, both iron and manganese may cycle rapidly between reduced and oxidized forms, further reinforcing the idea that iron may cycle in a closed or nearly-closed loop between Fe(III) and Fe(II) in wet soils. Soils that periodically dry and become oxic, such as rice paddies, may also employ molecular oxygen in this cycling, with the formation of amorphous Fe(III) oxides during the dry season, and reduction of this iron during the wet season. These reactions would restrict methane production, at least during the early part of the wet season, because Fe(III) reduction diverts electrons away from methane production.
The physical structure of iron oxides in the environment has a large effect on rates of Fe(III) reduction and populations of microbes. More strongly crystalline forms, such as goethite and hematite, are not readily reduced, while amorphous forms are consumed rapidly. Presumably it is amorphous forms that accumulate when Fe(II) is oxidized to Fe(III) and precipitates from solution as an oxide, leading to a labile pool of iron distinct from the highly resistant pool of crystalline mineral iron.
When it is not back-oxidized rapidly, Fe(II) and Mn(II) accumulate in solution, and are subject to water movements, potentially removing them from the site of production. Dissolved, reduced metal can be problematic, as these ions are readily oxidized by atmospheric oxygen if water flows contact the atmosphere, and will precipitate as rusty powder in drinking water and irrigation infrastructure. This also suggests that Fe(II) will not accumulate in natural systems over long periods, and that measuring the amount of Fe(II) in a sample is not a good indicator of iron-reducing microbial populations outside of microcosm experiments.
Iron oxides tend to provide strong adsorption surfaces for many other substances, including phosphate and heavy metals. The release to solution and movement of these substances can be a major concern when iron oxides are reduced.
Banded iron formations in some sediments and rocks, in which magnetite is deposited, appear to be the result of dissimilatory iron reduction. Magnetite is a mixed-valence iron oxide that behaves magnetically; small, characteristically-shaped crystals of it are produced intracellularly by magnetotactic bacteria, but much larger amounts are produced by a range of iron-reducing bacteria. The associated organic matter, masses and crystal structures of banded iron formations strongly suggest ancient dissimilatory iron reduction coupled to the decomposition of organic matter.
There are a range of factors controlling iron and manganese reduction in natural environments. Metal reduction is decreased in the presence of alternate electron acceptors, especially oxygen and nitrate. Oxygen is a thermodynamically favourable electron acceptor compared to either Fe(III) or Mn(IV), and nitrate appears to inhibit Fe(III) reduction by lowering electron availability below necessary levels. From a biological perspective, aerobic bacteria can usually outcompete iron reducers, many of whom are obligate anaerobes and are killed or inhibited by the presence of oxygen.
As previously described, the form of metal available in the environment also has a major effect on metal reduction rates, with more strongly crystalline forms most resistant to chemical alteration by microorganisms. “Poorly-crystalline” forms, presumably including amorphous metal oxides, are the major source of oxidized metals for reducing cells. Thus, extraction and measurement procedures that involve only the least crystalline forms of Fe(III) oxides are a good measure of the iron available for reduction by microbes. The widely-used oxalate extraction method normally does not extract significant quantities of Fe(III) oxides bound in strongly crystalline forms, except when “catalytic quantities” of Fe(II) are also available, as when weakly crystalline mixed-valence minerals such as magnetite are present. In such systems, oxalate extraction will gather a larger fraction of the total dithionate-citrate extractable iron, even though much of the iron so extracted will be in a form not actually available to soil microbes.
This paper is an enormous review of iron metabolisms in soils and sediments. It did not actually answer my current question, about the cycling of Fe(III) in soils and how comparisons of amorphous (oxalate-extractable) and total (dithionate-extractable) can be used to support inferences about long-term redox status in soils. However, I do think I now have a better grasp of general soil iron and manganese conditions and chemical transformations.
This author comprehensively reviews microbe-mediated reduction of iron and manganese in soils and sediments, in a long and detailed paper. Dissimilatory reduction is distinct from assimilatory reduction, in which metal ions are reduced when they are incorporated into cellular macromolecules such as enzymes and cofactors. Dissimilatory reduction is a process that ends with the accumulation of reduced metal outside the cell, and is responsible for the majority of iron and manganese reduction in sediments. While it has been observed in aerobic environments, such reduction occurs mainly in anaerobic conditions. Fe(III) in particular is most often reduced when it is the final electron acceptor in the anaerobic oxidation of organic molecules.
Microbes capable of dissimilatory metal reduction can be categorized in a number of ways. This author presents 5 categories, though I think there is considerable overlap between them, as in cases where one cell is able to reduce Fe(III) and metabolize a range of carbon sources. In the first category, reducing fermenting bacteria, the amount of Fe(III) reduced during metabolism is far less than the stoichiometry of the redox couple would suggest. This implies that Fe(III) is a minor electron acceptor during fermentation reactions, rather than the electron acceptor of choice or necessity for these organisms. Sulfur-oxidizing species in contrast do appear to reduce quantities of iron in line with stoichiometric predictions, but do not seem to gain significant energy from these reactions which occur under aerobic conditions on elemental sulfur. Hydrogen-oxidizing reducers appear to be abundant in anaerobic sediments, and seem to have high affinity for hydrogen gas; where Fe(III) is being reduced, hydrogen concentrations are low, and when hydrogen is added, Fe(III) reduction increases. These organisms need other material, such as simple organic molecules, to grow, as this reaction provides energy but little else. Organic-acid oxidizing reducers and aromatic-oxidizers are probably in many cases the same cells. As these molecules are sometimes the result of fermentation metabolisms, one possible food chain in anaerobic environments is fermentation, with some Fe(III) reduction, followed by greater levels of Fe(III) reduction linked to the decomposition of smaller organic molecules such as acetate. Such an ecological pairing is probably widespread, given the known abundance and diversity of fermenting species and the probable abundance of more aggressively iron-reducing species, many of which are probably Archaea rather than Bacteria.
There are three competing models for how Fe(III) is reduced in natural environments. The first is the enzymatic model, in which microbial cells employ membrane-bound or intracellular enzymes to transfer electrons to Fe(III) ions during the process of anaerobic oxidation of organic matter. The second is termed the redox model, and posits the majority of Fe(III) reduction is driven by equilibrium thermodynamics, with the relative levels of Fe(III) and Fe(II) in sediments controlled by abiotic factors such as temperature and pH. The third model is termed the direct-reduction model, in which some organic molecules react directly with Fe(III), without the intervention of cells or enzymes.
At first glance, the observation that Fe(III) reduction rates fall when microbes are removed from sediments supports the enzymatic model, but in fact all three models rely on microbial metabolisms. In the redox model, competing electron acceptors such as oxygen and nitrate are consumed by microbes, lowering their concentrations to levels too low to influence the transitions between Fe(III) and Fe(II). And the direct model relies on microbes releasing key organic molecules known to reduce Fe(III) in vitro. However, several other lines of evidence, such as the lack of spontaneous shifts in iron valency ratios in stored sediments, the large changes in pH associated with widely-used extraction methods, and the general rarity of rapidly-acting direct-reducing molecules all lead to the conclusion that while the other mechanisms may contribute some iron reducing activity in some situations, the overwhelming majority of Fe(III) reduction to Fe(II) occurring is driven by microbes and their enzymes.
This author spends nearly as much time discussing Mn(IV) reduction as Fe(III) reduction, but I have little interest in Mn chemistry at this time. However, in the discussion of the competing models of metal reduction, mention is made that Fe(II) in solution may reduce Mn(IV), removing accumulated Fe(II) from iron reduction and abiotically returning the iron to Fe(III) while generating Mn(II). This abiotic back-reaction closes the loop on cycling Fe(III).
Iron reduction is postulated as one of the first, if not the first, globally important metabolic pathway, with early Archaea using Fe(III) as their final electron acceptor in a generally reducing environment lacking free oxygen and nitrate. If Fe(III), a non-soluble, precipitating substance is the primary oxidizing agent, the redox environment of the Earth 2 billion years ago was upside-down compared to today: the surface and atmosphere was reducing, while buried and water-saturated sediments were oxidizing.
Freshwater swamps differ from aquatic and marine sediments in a number of ways, including a generally high organic content and plenty of sulfates. Under these conditions, both iron and manganese may cycle rapidly between reduced and oxidized forms, further reinforcing the idea that iron may cycle in a closed or nearly-closed loop between Fe(III) and Fe(II) in wet soils. Soils that periodically dry and become oxic, such as rice paddies, may also employ molecular oxygen in this cycling, with the formation of amorphous Fe(III) oxides during the dry season, and reduction of this iron during the wet season. These reactions would restrict methane production, at least during the early part of the wet season, because Fe(III) reduction diverts electrons away from methane production.
The physical structure of iron oxides in the environment has a large effect on rates of Fe(III) reduction and populations of microbes. More strongly crystalline forms, such as goethite and hematite, are not readily reduced, while amorphous forms are consumed rapidly. Presumably it is amorphous forms that accumulate when Fe(II) is oxidized to Fe(III) and precipitates from solution as an oxide, leading to a labile pool of iron distinct from the highly resistant pool of crystalline mineral iron.
When it is not back-oxidized rapidly, Fe(II) and Mn(II) accumulate in solution, and are subject to water movements, potentially removing them from the site of production. Dissolved, reduced metal can be problematic, as these ions are readily oxidized by atmospheric oxygen if water flows contact the atmosphere, and will precipitate as rusty powder in drinking water and irrigation infrastructure. This also suggests that Fe(II) will not accumulate in natural systems over long periods, and that measuring the amount of Fe(II) in a sample is not a good indicator of iron-reducing microbial populations outside of microcosm experiments.
Iron oxides tend to provide strong adsorption surfaces for many other substances, including phosphate and heavy metals. The release to solution and movement of these substances can be a major concern when iron oxides are reduced.
Banded iron formations in some sediments and rocks, in which magnetite is deposited, appear to be the result of dissimilatory iron reduction. Magnetite is a mixed-valence iron oxide that behaves magnetically; small, characteristically-shaped crystals of it are produced intracellularly by magnetotactic bacteria, but much larger amounts are produced by a range of iron-reducing bacteria. The associated organic matter, masses and crystal structures of banded iron formations strongly suggest ancient dissimilatory iron reduction coupled to the decomposition of organic matter.
There are a range of factors controlling iron and manganese reduction in natural environments. Metal reduction is decreased in the presence of alternate electron acceptors, especially oxygen and nitrate. Oxygen is a thermodynamically favourable electron acceptor compared to either Fe(III) or Mn(IV), and nitrate appears to inhibit Fe(III) reduction by lowering electron availability below necessary levels. From a biological perspective, aerobic bacteria can usually outcompete iron reducers, many of whom are obligate anaerobes and are killed or inhibited by the presence of oxygen.
As previously described, the form of metal available in the environment also has a major effect on metal reduction rates, with more strongly crystalline forms most resistant to chemical alteration by microorganisms. “Poorly-crystalline” forms, presumably including amorphous metal oxides, are the major source of oxidized metals for reducing cells. Thus, extraction and measurement procedures that involve only the least crystalline forms of Fe(III) oxides are a good measure of the iron available for reduction by microbes. The widely-used oxalate extraction method normally does not extract significant quantities of Fe(III) oxides bound in strongly crystalline forms, except when “catalytic quantities” of Fe(II) are also available, as when weakly crystalline mixed-valence minerals such as magnetite are present. In such systems, oxalate extraction will gather a larger fraction of the total dithionate-citrate extractable iron, even though much of the iron so extracted will be in a form not actually available to soil microbes.
This paper is an enormous review of iron metabolisms in soils and sediments. It did not actually answer my current question, about the cycling of Fe(III) in soils and how comparisons of amorphous (oxalate-extractable) and total (dithionate-extractable) can be used to support inferences about long-term redox status in soils. However, I do think I now have a better grasp of general soil iron and manganese conditions and chemical transformations.
Saturday, January 9, 2010
Liang and Balser 2008
Liang C, Balser TC. 2008. Preferential sequestration of microbial carbon in subsoils of a glacial-landscape toposequence, Dane County, WI, USA. Geoderma 148: 113-119.
These authors examined the microbial communities at a range of depths in soils near the University of Wisconsin, Madison campus. Soil organic carbon includes markers of microbial groups such as amino sugars, molecules that are absent from plants and specific to some groups of soil microorganisms. As markers, these molecules have several advantages; besides their utility in identifying organisms, they are stable in soils, persist after cell depth, and can apparently be extracted and examined using fairly simple laboratory techniques plus access to a gas chromatograph.
This study represents a general survey of soil carbon in this system, an examination of the pools and fates of different carbon molecule classes as well as the contributions of broad groups such as bacteria and fungi to soil physiology in different soil horizons. There were three main conclusions:
1. Upper soil horizons are relatively enriched both in total SOC and amino sugars. The source of this material almost certainly is some combination of surface plant litter and root exudates, not surprisingly supporting a large community of microorganisms in the near-surface soil.
2. Amino sugars accumulate in subsoils, despite the redox environment (presumably somewhat negative) associated with the water table.
3. Amino sugars, while useful, are not sufficient on their own to elucidate mechanisms of SOC turnover and sequestration by soil microbes. Variability between sites and between horizons suggests a major role of both history and site-specific factors in structuring communities at a level distinguishable by ratios of various amino sugars.
This paper is one of a handful I have that explicitly examine microbial communities and variation by depth. However, as this paper describes what I think is a first-look at a soil microbiological system, it lacks some detail and strong conclusions. Amino sugars may be useful in my own research, though I think our lab has more familiarity with other techniques useful for examining interactions between soil bacteria and fungi.
Beyond my current research, this paper is clearly written, not too long, and presents a set of well-described investigations built on a solid foundation of general theory. This suggests it may be useful as a teaching tool, perhaps as a paper a 2nd-year undergraduate would have the skills and knowledge to understand.
These authors examined the microbial communities at a range of depths in soils near the University of Wisconsin, Madison campus. Soil organic carbon includes markers of microbial groups such as amino sugars, molecules that are absent from plants and specific to some groups of soil microorganisms. As markers, these molecules have several advantages; besides their utility in identifying organisms, they are stable in soils, persist after cell depth, and can apparently be extracted and examined using fairly simple laboratory techniques plus access to a gas chromatograph.
This study represents a general survey of soil carbon in this system, an examination of the pools and fates of different carbon molecule classes as well as the contributions of broad groups such as bacteria and fungi to soil physiology in different soil horizons. There were three main conclusions:
1. Upper soil horizons are relatively enriched both in total SOC and amino sugars. The source of this material almost certainly is some combination of surface plant litter and root exudates, not surprisingly supporting a large community of microorganisms in the near-surface soil.
2. Amino sugars accumulate in subsoils, despite the redox environment (presumably somewhat negative) associated with the water table.
3. Amino sugars, while useful, are not sufficient on their own to elucidate mechanisms of SOC turnover and sequestration by soil microbes. Variability between sites and between horizons suggests a major role of both history and site-specific factors in structuring communities at a level distinguishable by ratios of various amino sugars.
This paper is one of a handful I have that explicitly examine microbial communities and variation by depth. However, as this paper describes what I think is a first-look at a soil microbiological system, it lacks some detail and strong conclusions. Amino sugars may be useful in my own research, though I think our lab has more familiarity with other techniques useful for examining interactions between soil bacteria and fungi.
Beyond my current research, this paper is clearly written, not too long, and presents a set of well-described investigations built on a solid foundation of general theory. This suggests it may be useful as a teaching tool, perhaps as a paper a 2nd-year undergraduate would have the skills and knowledge to understand.
Wagner et al. 2009
Wagner D, Kobabe S, Liebner S. 2009. Bacterial community structure and carbon turnover in permafrost-affected soils of the Lena Delta, northeastern Siberia. Canadian Journal of Microbiology 55: 73-83.
These authors examined the microbial communities at two depth bands (near-surface and near-permafrost) in low-centred tundra polygons at the vast permafrost wetland of the delta of the Lena River. The delta covers more than 60 000 km^2, and much of it appears to be a reserve or national park of Russia. The CAVM (Walker et al. 2002) describes most of the delta as vegetation type W2, sedge, moss, dwarf-shrub wetland, and satellite images from Google maps shows very extensive lake and pond coverage of the landscape. In short, it’s pretty wet, and generally cold.
The general finding of this paper is that while near-surface communities include a wide diversity of aerobic and facultatively-anaerobic bacteria, the deeper, colder, anaerobic portions of the soil contain almost no aerobes, and are instead dominated by “fermenting” species capable of decomposing recalcitrant organic carbon molecules under negative-redox conditions. There is a sharp temperature gradient, which combined with the poorer quality of carbon, the lack of oxygen and negative redox conditions, and the general water saturation at depth creates conditions near the permafrost suitable only for the slow microbial metabolisms. None of this is particularly surprising, but the observation of decreased biodiversity with water saturation does suggest the worrying possibility that increased water in this system, driven by melting permafrost and climate change (particularly upstream in the long and North-flowing Lena) could drive these microbial communities to lose some “physiological skills” such as the ability to oxidize methane, a metabolic pathway possessed only by some aerobic prokaryotes.
This paper is quite important to my own work, I think. Besides emphasizing the role of water content in structuring soil chemical and especially biological conditions, the description of the methods used to measure microbial biodiversity should be useful. However, while the BIOLOG plates seem interesting, the results of this technique are not at all well explained in this paper. I do not know what is indicated by the relationship shown in Figure 3, for example, of changes in colour development associated with carbon turnover of various categories of organic substrates. Several of the figures are simple plots of principal component analysis (PCA), literally just PC1 vs. PC2 with some outlines drawn around some clusters. I’m sure there is more of interest in this paper besides the coarse outline of biodiversity differences in communities, but without a more thorough explanation of the nearly-raw data I cannot see it.
These authors examined the microbial communities at two depth bands (near-surface and near-permafrost) in low-centred tundra polygons at the vast permafrost wetland of the delta of the Lena River. The delta covers more than 60 000 km^2, and much of it appears to be a reserve or national park of Russia. The CAVM (Walker et al. 2002) describes most of the delta as vegetation type W2, sedge, moss, dwarf-shrub wetland, and satellite images from Google maps shows very extensive lake and pond coverage of the landscape. In short, it’s pretty wet, and generally cold.
The general finding of this paper is that while near-surface communities include a wide diversity of aerobic and facultatively-anaerobic bacteria, the deeper, colder, anaerobic portions of the soil contain almost no aerobes, and are instead dominated by “fermenting” species capable of decomposing recalcitrant organic carbon molecules under negative-redox conditions. There is a sharp temperature gradient, which combined with the poorer quality of carbon, the lack of oxygen and negative redox conditions, and the general water saturation at depth creates conditions near the permafrost suitable only for the slow microbial metabolisms. None of this is particularly surprising, but the observation of decreased biodiversity with water saturation does suggest the worrying possibility that increased water in this system, driven by melting permafrost and climate change (particularly upstream in the long and North-flowing Lena) could drive these microbial communities to lose some “physiological skills” such as the ability to oxidize methane, a metabolic pathway possessed only by some aerobic prokaryotes.
This paper is quite important to my own work, I think. Besides emphasizing the role of water content in structuring soil chemical and especially biological conditions, the description of the methods used to measure microbial biodiversity should be useful. However, while the BIOLOG plates seem interesting, the results of this technique are not at all well explained in this paper. I do not know what is indicated by the relationship shown in Figure 3, for example, of changes in colour development associated with carbon turnover of various categories of organic substrates. Several of the figures are simple plots of principal component analysis (PCA), literally just PC1 vs. PC2 with some outlines drawn around some clusters. I’m sure there is more of interest in this paper besides the coarse outline of biodiversity differences in communities, but without a more thorough explanation of the nearly-raw data I cannot see it.
Friday, January 8, 2010
Lovley and Phillips 1986
Lovley DR, Phillips EJP. 1986. Organic matter mineralization with reduction of ferric iron in anaerobic sediments. Applied and Environmental Microbiology 51: 683-689.
These authors examined the reduction of iron, both supplemented and already present, in sediments collected from the bottom of the Potomac river across a salinity gradient from freshwater to brackish estuarine. A previous hypothesis in the literature suggested that observed decreases in methane production in the presence of trivalent iron, Fe(III) were caused by Fe(III) being toxic to methanogenic prokaryotes. This hypothesis was disproved in this study, and the results of this study suggest instead that methanogens are outcompeted by iron-reducing species because of the greater thermodynamic benefits of Fe(III) reduction compared to methane production. Van Bodegom et al. (2004) reported instead that rather than competition, Fe(III) directly inhibits methanogenesis, though I think this is due to metabolic switching within individual cells, not interactions between distinct methanogenic and iron-reducing populations.
The other major finding of this paper is that the form of the Fe(III) in the environment has a major effect on Fe(III) reduction. Amorphous ferric oxyhydroxides are reduced much more readily than are crystalline forms. These authors do not speculate on the mechanism underlying this difference, though I suspect surface area exerts a major controlling influence.
These authors did not measure Fe(III) forms in native sediments, rather they added amorphous Fe(III) and measured Fe(II) after incubation. Thus, while my own studies of the ratios of amorphous to crystalline Fe(III) are not assisted by these techniques, these authors do provide a clear and apparently fairly simple protocol for the measurement of Fe(II), involving extraction by HCl and reaction with a molecule that turns purple when complexed with Fe(II) allowing measurement of Fe(II) amounts from the absorbance spectrum of the resulting solution. I need to learn more about Fe(III) biogeochemistry before deciding to pursue such an analysis; I suspect one fate of Fe(III) is to cycle through an organism and be returned to the oxidized state, rather than shuttling directly from Fe(III) to Fe(II).
These authors examined the reduction of iron, both supplemented and already present, in sediments collected from the bottom of the Potomac river across a salinity gradient from freshwater to brackish estuarine. A previous hypothesis in the literature suggested that observed decreases in methane production in the presence of trivalent iron, Fe(III) were caused by Fe(III) being toxic to methanogenic prokaryotes. This hypothesis was disproved in this study, and the results of this study suggest instead that methanogens are outcompeted by iron-reducing species because of the greater thermodynamic benefits of Fe(III) reduction compared to methane production. Van Bodegom et al. (2004) reported instead that rather than competition, Fe(III) directly inhibits methanogenesis, though I think this is due to metabolic switching within individual cells, not interactions between distinct methanogenic and iron-reducing populations.
The other major finding of this paper is that the form of the Fe(III) in the environment has a major effect on Fe(III) reduction. Amorphous ferric oxyhydroxides are reduced much more readily than are crystalline forms. These authors do not speculate on the mechanism underlying this difference, though I suspect surface area exerts a major controlling influence.
These authors did not measure Fe(III) forms in native sediments, rather they added amorphous Fe(III) and measured Fe(II) after incubation. Thus, while my own studies of the ratios of amorphous to crystalline Fe(III) are not assisted by these techniques, these authors do provide a clear and apparently fairly simple protocol for the measurement of Fe(II), involving extraction by HCl and reaction with a molecule that turns purple when complexed with Fe(II) allowing measurement of Fe(II) amounts from the absorbance spectrum of the resulting solution. I need to learn more about Fe(III) biogeochemistry before deciding to pursue such an analysis; I suspect one fate of Fe(III) is to cycle through an organism and be returned to the oxidized state, rather than shuttling directly from Fe(III) to Fe(II).
Wednesday, January 6, 2010
Siciliano et al. 2009
Siciliano SD, Ma WK, Ferguson S, Farrell RE. 2009. Nitrifier dominance of Arctic soil nitrous oxide emissions arises to due fungal competition with denitrifiers for nitrate. Soil Biology and Biochemistry 41: 1104-1110.
These authors examined the nitrous oxide emissions, microbial communities, and some components of nitrogen cycling in soils from three landforms at Truelove Lowland, on Devon Island. Previous results (Ma et al. 2007) had indicated that Arctic nitrous oxide emissions are not sensitive to soil moisture, at least in the range of 50% to saturated water filled pore space. This study includes a series of incubations of soil samples at a range of temperatures similar to ambient conditions, and treatments to disrupt fungi or particular types of prokaryotes.
Large differences in community composition were found between the three landforms, with the highest biomass and fungi:bacteria ratio in the wet sedge meadow and lowest in the raised beach crest (the lower foreslope was intermediate by these measures). Competition between fungi and denitrifiers for soil nitrate pools was inferred as the mechanism allowing dominance of emitted N2O by nitrifiers; fungi and denitrifiers are busy scavenging every available electron acceptor starting with nitrate and running all the way down to N2 gas, so almost any N2O that escapes was generated by nitrifiers in conditions not favoured by either of the other major groups.
This paper serves to demonstrate the very complex nature of soil biology, especially regarding the multiple and interacting pathways that may produce or consume materials of interest such as N2O. The references in this paper should be useful for digging into this complexity.
These authors examined the nitrous oxide emissions, microbial communities, and some components of nitrogen cycling in soils from three landforms at Truelove Lowland, on Devon Island. Previous results (Ma et al. 2007) had indicated that Arctic nitrous oxide emissions are not sensitive to soil moisture, at least in the range of 50% to saturated water filled pore space. This study includes a series of incubations of soil samples at a range of temperatures similar to ambient conditions, and treatments to disrupt fungi or particular types of prokaryotes.
Large differences in community composition were found between the three landforms, with the highest biomass and fungi:bacteria ratio in the wet sedge meadow and lowest in the raised beach crest (the lower foreslope was intermediate by these measures). Competition between fungi and denitrifiers for soil nitrate pools was inferred as the mechanism allowing dominance of emitted N2O by nitrifiers; fungi and denitrifiers are busy scavenging every available electron acceptor starting with nitrate and running all the way down to N2 gas, so almost any N2O that escapes was generated by nitrifiers in conditions not favoured by either of the other major groups.
This paper serves to demonstrate the very complex nature of soil biology, especially regarding the multiple and interacting pathways that may produce or consume materials of interest such as N2O. The references in this paper should be useful for digging into this complexity.
Tuesday, January 5, 2010
Elberling 2007
Elberling B. 2007. Annual soil CO2 effluxes in the High Arctic: the role of snow thickness and vegetation type. Soil Biology and Biochemistry 39: 646-654.
This author studied the total annual efflux of CO2 at three vegetation communities in Endalen valley on Svalbard. The three communities are each dominated by one characteristic species of plant, and are named accordingly: Dryas, Cassiope, and Salix, and from the description of the sites and their environmental parameters, there appears to be high agreement between these communities and those found at Alexandra Fjord, Ellesmere Island.
The depth and duration of snow cover was a major factor controlling (directly and indirectly) soil conditions and thus respiration. Snow depth varied with vegetation type, though the causal relationship is probably snow to plants, via soil temperature (more snow = higher winter temperatures) and soil moisture content (snow accumulates at and melts into depressions and certain slope positions). Higher temperatures and wetter conditions correlated with higher soil respiration, both in winter and summer. All sites experienced a brief period of water saturation in the upper 5cm of the soil during spring thaw, though sites varied in when thaw happened, with Dryas first and Salix last, corresponding with winter snow cover depth.
Soil conditions among the sites seem to have been broadly similar; not surprising considering the close proximity of sites and the consistent soil type across the valley, though soil under Cassiope tetragonal patches was more acidic. This acidity seems related to a reduced concentration of base cations (especially Ca2+ and K+) under Cassiope plants.
Summer water content did not correlate with annual CO2 flux, which this author attributes to the generally well-drained soils, a lack of large precipitation events, and long periods without rain leading to typically dry soils everywhere, though soil respiration at the Dryas site may have been water-limited, as this was the driest site.
Winter temperatures in the soil averaged warmer than -10ºC at all sites, warm enough for microbial activity. A burst of CO2 during spring thaw was not predicted from soil parameters, but was attributed to increasing microbial activity associated with warming temperatures and the release of high-quality organic material from winter-killed microbial cells. Winter CO2 efflux averages were 0.11 to 0.28 µmol / m^2 / s, not far from values we found (for example) at the Cassiope site at Alexandra Fjord.
This paper contains much that is valuable to my current research, including both the data and patterns found and the discussion with other relevant references.
This author studied the total annual efflux of CO2 at three vegetation communities in Endalen valley on Svalbard. The three communities are each dominated by one characteristic species of plant, and are named accordingly: Dryas, Cassiope, and Salix, and from the description of the sites and their environmental parameters, there appears to be high agreement between these communities and those found at Alexandra Fjord, Ellesmere Island.
The depth and duration of snow cover was a major factor controlling (directly and indirectly) soil conditions and thus respiration. Snow depth varied with vegetation type, though the causal relationship is probably snow to plants, via soil temperature (more snow = higher winter temperatures) and soil moisture content (snow accumulates at and melts into depressions and certain slope positions). Higher temperatures and wetter conditions correlated with higher soil respiration, both in winter and summer. All sites experienced a brief period of water saturation in the upper 5cm of the soil during spring thaw, though sites varied in when thaw happened, with Dryas first and Salix last, corresponding with winter snow cover depth.
Soil conditions among the sites seem to have been broadly similar; not surprising considering the close proximity of sites and the consistent soil type across the valley, though soil under Cassiope tetragonal patches was more acidic. This acidity seems related to a reduced concentration of base cations (especially Ca2+ and K+) under Cassiope plants.
Summer water content did not correlate with annual CO2 flux, which this author attributes to the generally well-drained soils, a lack of large precipitation events, and long periods without rain leading to typically dry soils everywhere, though soil respiration at the Dryas site may have been water-limited, as this was the driest site.
Winter temperatures in the soil averaged warmer than -10ºC at all sites, warm enough for microbial activity. A burst of CO2 during spring thaw was not predicted from soil parameters, but was attributed to increasing microbial activity associated with warming temperatures and the release of high-quality organic material from winter-killed microbial cells. Winter CO2 efflux averages were 0.11 to 0.28 µmol / m^2 / s, not far from values we found (for example) at the Cassiope site at Alexandra Fjord.
This paper contains much that is valuable to my current research, including both the data and patterns found and the discussion with other relevant references.
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