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8.08 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes J

8.08 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes J

8.08 Anaerobic : Linkages to Trace Gases and Aerobic Processes J. P. Megonigal Smithsonian Environmental Research Center, Edgewater, MD, USA M. E. Mines University of Massachusetts Lowell, MA, USA and P. T. Visscher University of Connecticut Avery Point, Groton, CT, USA

8.08.1 OVERVIEW OF IN THE ABSENCE OF 02 319 8.08.1.1 Introduction 319 8.08.1.2 Overview of Anaerobic Metabolism 319 8.08.1.3 Anaerobic-Aerobic Interface 321 &08J.4 gy»!az of MekzWism 321 8.08.2 AUTOTROPHIC METABOLISM 322 8.08.2.1 (Photolithoautotrophy) Diversity and Metabolism 322 8.08.2.2 (Chemolithoautotrophy) Diversity and Metabolism 324 &0&ZJ Podwayaq/'CC^Fwan'on 325 8.08.3 AND 325 8.08.3.1 Degradation 326 8.08.3.1.1 Polysaccharides 327 &08JJ.2 iignin 327 &0&3.2 FenMfnfoA'on 328 &08J.27 AcfiogenaMf 329 8.08.3.2.2 Syntrophy and interspecies transfer 329 8.08.4 332 8.08.4.1 Methane in the Environment 333 8.08.4.2 Diversity and Metabolism 334 8.08.4.3 Regulation of 335 8.08.4.3.1 O2 and oxidant inhibition 335 &0&4J.2 Mdrwmfa,W;,# 335 8.08.4.3.3 Temperature 336 8.08.4.3.4 quantity and quality 336 8.08.4.3.5 as carbon sources 337 8.08.4.4 Contributions of Acetotrophy versus Hydrogenotrophy 339 8.08.4.4.1 Organic carbon availability 339 &08.^.^.2 rgmpemfure 341 8.08.4.5 Anaerobic Methane Oxidation 342 8.08.4.6 Aerobic Methane Oxidation 344

317 318 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

8.08.4.6.1 diversity 344 8.08.4.6.2 Regulation of methanotrophy 346 8.08.4.63 Methane oxidation efficiency 347 8.08.4.7 Methane Emissions and Global Change 348 8.08.5 350 8.08.5.1 Nitrogen in the Environment 350 8.08.5.2 351 8.08.5.3 Respiratory 351 8.08.5.3.1 Denitrifier diversity and metabolism 352 8.08.5.3.2 Regulation of denitrification 353 8.08.5.3.3 -denitrification Coupling 355 8.08.5.3.4 Animals and plants 356 &0&5J.5 NzOfWNO/bzej 356 8.08.5.4 Dissimilatory Reduction to (DNRA) 357 8.08.5.4.1 Physiology and diversity of DNRA 357 8.08.5.4.2 DNRA versus denitrification 358 8.08.5.5 Alternative Pathways to N2 Production 359 8.08.5.5.1 359 8.08.5.5.2 Nitrifier denitrification 361 8.08.5.5.3 Abiotic and autotrophic denitrification 361 8.08.6 AND 362 8.08.6.1 Iron and Manganese in the Environment 362 8.08.6.2 Iron and Manganese Geochemistry 363 8.08.6.3 Microbial Reduction of Iron and Manganese 364 8.08.6.3.1 Metabolic diversity in Fe(III)- and Mn(IV)-reducing 364 8.08.6.3.2 Physical mechanisms for accessing 365 8.08.6.4 Factors that Regulate Fe(III) Reduction 366 8.08.6.4.1 shuttling compounds 366 &0&.6.4.2 Fef/f/JcWafora 367 8.08.6.4.3 reactivity 367 8.08.6.4.4 Abiotic versus biotic reduction 369 8.08.6.4.5 Separating enzymatic and nonenzymatic Fe(III) reduction 369 8.08.6.4.6 Fe(III) reduction in 370 8.08.6.5 Microbial Oxidation of Iron and Manganese 371 8.08.6.5.1 Energetics of Fe(II) and Mn(II) oxidation 371 8.08.6.5.2 Anaerobic Fe(II) oxidation 371 &0&6.5J Aemh'c Fg(7f) ojcidafion 372 &0&&6 /ronCyc/ing 373 8.08.7 374 &0&7.7 W/urGeoc/wmwfry 374 8.08.7.2 Microbial Reduction of 375 8.08.7.2.1 Overview of sulfate reduction 375 8.08.7.2.2 Metabolic diversity 376 8.08.7.3 Taxonomic Considerations 377 8.08.7.4 Sulfate-reducing Populations 378 8.08.7.5 Factors Regulating Sulfate Reduction Activity 379 8.08.7.5.1 Sulfate-reducing activity 379 &0& 7.5.2 Tempenzfurg 380 &0&7.5J Carbon 380 8.08.7.5.4 Sulfate and molecular concentrations 381 8.08.7.6 Microbial Reduction of Sulfur 381 8.08.7.7 Disproportionation 382 &0&7.& W/ur Gases 383 &0&. 7.8.7 gyifwgg»si#(k 383 &0& 7.8.2 MefAyW/Mea 383 8.08.7.8.3 Carbonyl and carbon disulfide 385 8.08.7.9 Microbial Oxidation of Sulfur 386 8.08.7.9.1 Colorless sulfur bacteria 386 8.08.7.9.2 Single- colorless sulfur bacteria 387 8.08.7.9.3 Filamentous colorless sulfur bacteria 388 8.08.7.9.4 Anoxyphototrophic bacteria 388 8.08.7.9.5 Ecological aspects of sulfide oxidation 389 8.08.8 COUPLED ANAEROBIC ELEMENT CYCLES 389 8.08.8.1 Evidence of Competitive Interactions 389 8.08.8.2 Mechanisms of 390 &0&&3 ExcgfAofu 391 8.08.8.4 Noncompetitive Interactions 391 8.08.8.5 Contributions to Carbon Metabolism 391 8.08.8.6 Concluding Remarks 392 ACKNOWLEDGMENTS 392 REFERENCES 392 Overview of Life in the Absence of O2 319 8.08.1 OVERVIEW OF LIFE IN (Schlesinger, 1997). Anaerobic THE ABSENCE OF 02 are common symbionts of , cattle, and many other animals, where they aid digestion. 8.08.1.1 Introduction Nutrient and pollutant chemistry are strongly Life evolved and flourished in the absence of modified by the reduced conditions that prevail in wetland and aquatic ecosystems. molecular oxygen (02). As the 02 content of the rose to the present level of 21% This review of anaerobic metabolism empha- beginning about two billion years ago, anaerobic sizes aerobic oxidation, because the two pro- metabolism was gradually supplanted by aerobic cesses cannot be separated in a complete metabolism. Anaerobic environments have per- treatment of the topic. It is process oriented and sisted on despite the transformation to an highlights the fascinating microorganisms that oxidized state because of the combined influence mediate anaerobic . We begin of and organic matter. Molecular oxygen this review with a brief discussion of C02 diffuses about 104 times more slowly through assimilation by , the source of most water than air, and organic matter supports a large of the reducing power on Earth, and then consider the biological processes that harness biotic 02 demand that consumes the supply faster than it is replaced by diffusion. Such conditions this potential . Energy liberation begins exist in , , estuaries, coastal marine with the decomposition of organic macromol- sediments, , anoxic water columns, sew- ecules to relatively simple compounds, which are age digesters, landfills, the intestinal tracts of simplified further by fermentation. Methanogen- animals, and the rumen of . Anaerobic esis is considered next because CH4 is a product microsites are also embedded in oxic environ- of fermentation, and thus completes the ments such as upland and marine water of organic matter, particularly in the columns. Appreciable rates of aerobic respiration absence of inorganic electron acceptors. Finally, are restricted to areas that are in direct contact the organisms that use nitrogen, manganese, iron, with air or those inhabited by organisms that and sulfur for terminal electron acceptors are considered in order of decreasing free-energy produce 02. yield of the reactions. Rising atmospheric 02 reduced the global area of anaerobic , but enhanced the overall rate of anaerobic metabolism (at least on an area basis) by increasing the supply of electron donors and 8.08.1.2 Overview of Anaerobic Metabolism acceptors. Organic carbon production increased dramatically, as did oxidized forms of nitrogen, Microorganisms derive energy by transferring manganese, iron, sulfur, and many other elements. from an external electron source or In contemporary anaerobic ecosystems, nearly all donor to an external electron sink or terminal of the reducing power is derived from photosyn- . Organic electron donors vary thesis, and most of it eventually returns to 02, the from monomers that support fermentation to most electronegative electron acceptor that is simple compounds such as acetate and CH4. The abundant. This photosynthetically driven common inorganic electron donors are molecular gradient has been thoroughly exploited by aerobic hydrogen (H2), ammonium (NK|~), manganous and anaerobic microorganisms for metabolism. manganese (Mn(II)), ferrous iron (Fe(II)), and The same is true of hydrothermal vents (H2S). Energy is harnessed by (Tunnicliffe, 1992) and some deep subsurface shuttling electrons through chains within environments (Chapelle et ah, 2002), where a cell until a final transfer is made to a terminal thermal energy is the ultimate source of the electron acceptor. The common terminal electron reducing power. acceptors are nitrate (NOJ), manganic manganese Although anaerobic habitats are currently a Mn(IV), ferric iron (Fe(III)), sulfate (SO4"), and small fraction of Earth's surface area, they have a (C02) (Table 1). profound influence on the biogeochemistry of the Anaerobic organisms often have the capacity to planet. This is evident from the observation that reduce two or more terminal electron acceptors. In the 02 and CH4 content of Earth's atmosphere are many cases, these alternative reactions do not in extreme disequilibrium (Sagan et ah, 1993). support growth, as with the fermenting bacteria The combination of high aerobic primary pro- that reduce Fe(III) (Lovley, 2000b). In other cases, duction and anoxic sediments provided the large the ability to use multiple electron acceptors is deposits of fuels that have become vital and presumably an adaptation for remaining active in contentious sources of energy for modern indus- an environment where the supply of specific trialized societies. Anaerobic metabolism is electron acceptors is variable. For example, responsible for the of N2 in the denitrification permits normally aerobic bacteria atmosphere; otherwise N2-fixing bacteria would to respire in the absence of 02, albeit at a slower have consumed most of the N2 pool long ago rate. 320 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

Table 1 Thermodynamic sequence for reduction of fermentation, S04~ reduction, methanogenesis inorganic substances by organic matter2. and most other anaerobic pathways make them particularly sensitive to the concentrations of Reaction Eh AGh substrates and products. Because fermentation is (V) inhibited by its end products, the process is Reduction of O? dependent on end product consumption by non- 02 + 4H+ + 4e~ ^ 2H20 0.812 -294 fermentative organisms. This thermodynamic Reduction of NO^ dependence among anaerobic microorganisms, in NO^ + 6H+ + 6e~ ^ N2 0.747 -28.4 which several must function together in + 3H20 4+ 2+ order to consume a single substrate, is known as Reduction of Mn to Mn 2+ syntrophy, which is a type of . Mutual- Mn02 + 4H+ + 2e" i=! Mn 0.526 -233 + 2H20 ism is a particularly important ecological inter- Reduction of Fe3+ to Fe2+ action that determines the structure and function of 2+ Fe(OH)3 + 3H+ +e'^ Fe - 0.047 -10.1 anaerobic communities. + 3H20 Competition is another important ecological Reduction of SOj~ to H2S 2- mechanism that affects the structure and function SO + 10H+ + 8e~ ^ H2S -0.221 -5.9 of anaerobic microbial communities. Competition + 4H20 for fermentation products, such as H and Reduction of C0 to CH 2 2 4 acetate, is particularly keen, because they C02 + 8H+ + 8e" ^ CH4 - 0.244 -5.6 + 2H20 usually limit microbial activity in anaerobic ecosystems. The outcome of competition for Source: Schlesinger (1997). a Units are kcal mol e assuming coupling to the oxidation reaction these substrates favors those pathways with the + b |CH20 + iH20 — iC02 + H + e". AC = -RT In (K); pH 7.0, highest thermodynamic yield: NO^ reduction > 25 °C. Mn(IV) reduction > Fe(III) reduction > SO4" reduction > HCOJ reduction (i.e., methanogen- esis) (Table 1). This hierarchy also tends to apply The most fundamental difference between to the expression of metabolic pathways within a aerobic and anaerobic metabolism is energy yield. single that can reduce multiple -1 Oxidation of yields —2,900 kJ mol under electron acceptors (Nealson and Myers, 1992). aerobic conditions (Equation (1)) compared to The presence of 02 suppresses all the anaerobic -1 —400 kJ mol under typical methanogenic metabolic pathways primarily because it is a conditions (Equation (2)): toxin, but also because it has the highest energy C H 0 + 60 — 6C0 + 6H 0 (1) yield of all the common terminal electron 6 12 6 2 2 2 acceptors. To a first approximation, a single C<;H,206-»3C02 + 3CH4 (2) metabolic pathway dominates anaerobic carbon cycling until it is limited by the availability of The high-energy yield from aerobic metabolism electron acceptors. However, it is not uncommon permits a single to completely oxidize for the pathways to coexist because of spatial complex organic compounds to C0 . In compari- 2 variation in the abundance of terminal electron son, no single anaerobic microorganism can acceptors. Coexistence may also occur because completely degrade organic to C0 and 2 the supply of a competitive (e.g., H20 (Fenchel and Finlay, 1995), and most have highly specialized substrate demands. Under H2) is large and nonlimiting, or because there is a anaerobic conditions, the mineralization of organic supply of a noncompetitive electron donor that can be used by some organisms, but not others carbon to C02 is a multistep process that involves a consortium of organisms, each of which conserves (e.g., Oremland et al., 1982). The outcome of a fraction of the potential-free energy that was competition for electron donors establishes the available in the original organic carbon substrate. redox potential (Yao and Conrad, 1999), and not As a result, anaerobic organisms are adapted to vice versa as is often suggested. conserve quantities of energy that often approach The thermodynamic yield of the various anaero- the theoretical minimum (20kJmol_1) required bic pathways is related to their affinity for for metabolism (Valentine, 2001). substrates. Sulfate-reducing bacteria outcompete The first step in anaerobic decomposition is the by maintaining H2 concentrations breakdown of complex organic molecules to below the threshold for the process to be thermo- simple molecules such as sugars (Section dynamically feasible (Lovley et al., 1982), and 8.08.3.1). Next, these are fermented to even Fe(III) reducers do the same to SO4" reducers. In simpler molecules such as acetate and H2 (Section fact, the equilibrium concentration of H2 can be 8.08.3.2). In the final step, fermentation products used to predict the dominant metabolic pathways in serve as electron donors for the reduction of a given environment (Lovley and Goodwin, 1988; inorganic compounds. The low-energy yield of Lovley et al., 1994a). Overview of Life in the Absence of O2 321

8.08.1.3 Anaerobic-Aerobic Interface Habitats atmosphere or water column (e.g., 02 and S04~ in marine ecosystems); others are regenerated at the Spatial variation in the abundance of electron aerobic-anaerobic interface due to oxidation of donors and acceptors explains large-scale and NH|, Fe(II), Mn(II), or H2S. Regeneration at the small-scale patterns of anaerobic metabolism. aerobic-anaerobic interface can supply a large Sulfate reduction dominates anaerobic carbon fraction of the terminal electron acceptors con- metabolism on about two-thirds of the planet sumed in anaerobic metabolism. because of the high abundance of SO4 in Oxidant regeneration is stimulated dramatically seawater (Capone and Kiene, 1988). Fe(III) by the presence of animals and plants. The bur- reduction is important in all anaerobic ecosystems rowing activity of animals (i.e., bioturbatiori) is a with mineral-dominated soils or sediments, particularly important agent of Fe(III) and Mn(IV) regardless of whether they are marine or fresh- regeneration in marine sediments and salt marshes water (Thamdrup, 2000). Methanogenesis is (Thamdrup, 2000; Kostka et al, 2000c; Gribsholt important in freshwater environments generally, et al, 2003; Gribsholt and Kristensen, 2002). and it dominates the anaerobic carbon metabolism Wetland plants promote regeneration by serving as of bogs, fens, and other wetlands that exist on conduits of 02 infusion deep into the profile. organic (i.e., peat) soils. Such plants tolerate flooding partly by supplying The contribution of the various anaerobic 02 to their root system, where some of it leaks into metabolic pathways to carbon metabolism varies the soil due to radial 02 loss. In the absence of temporally and spatially due to changes in the physical mixing, burrowing, or radial 02 loss, the abundance of electron donors and acceptors depth of 02 penetration into saturated soils and (Figure 1). Organic carbon is most abundant at sediments is a few millimeters. The influence of the surface of soils and sediments where is plants and animals on anaerobic metabolism is deposited, most of which is derived from aerobic similar in the sense that both effectively increase . The aerobic zone is also the the aerobic-anaerobic surface area (Mayer et al., source of most terminal electron acceptors, some 1995; Armstrong, 1964). In addition, plants are a of which diffuse into the anaerobic zone from the source of labile organic carbon compounds that fuel anaerobic metabolism. Competition for fermentation products produces a succession of dominant metabolic pathways as Aerobic distance from the source of electron donors and acceptors increases. Aerobic metabolism domi- nates the surface of sediments, and methanogenesis Nitrate reduction the deeper depths, as suggested by their free energy yield (Table 1). There is often very little overlap between each zone, suggesting nearly complete exclusion of one group by another. The same Manganese reduction pattern is observed with distance from the surface of a root or burrow, or with distance downstream from an organic pollutant source in rivers and Iron aquifers. reduction 8.08.1.4 Syntax of Metabolism Sulfate reduction In ecological and biogeochemical studies, any organism can be described by three basic attributes that define a "feeding" (Greek trophos) niche: Methane energy source, electron donor source, and carbon reduction source (Table 2). Organisms that use energy are phototrophic, while those that use chemical energy are chemotrophic. Organisms that use Figure 1 Vertical biogeochemical zones in sediments. inorganic electron donors are lithotrophic, while The top is the sediment-water interface. Processes on those that use organic electron donors are organo- the left represent the use of various electron acceptors trophic. Finally, autotrophic organisms assimilate (respirations) during the degradation of organic matter. Plots on the right represent the chemical profiles most C02 for , while heterotrophic organ- widely used to delineate the vertical extent of each zone. isms assimilate organic carbon (e.g., Lwoff et al., Rotating the figure 90° to the left shows the sequence of 1946; Barton et al., 1991). The modifiers are linked electron acceptors used over time (x-axis) if a sample of in the order energy-electrons-carbon. For oxic sediment were enclosed and allowed to become example, photosynthesis is technically anaerobic over time. photolithoautotrophy. It is not surprising that 322 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes Table 2 Classification terms for . Each term is composed of modifiers for the source of energy (chemical versus light), the source of electrons (inorganic versus organic), and the source of carbon (inorganic versus organic). The modifiers are linked in the order energy-electrons-carbon. for which organic carbon is both the carbon and energy source are abbreviated as "heterotrophy." Energy and carbon source Electron source Inorganic (lithotrophy) Organic (organotrophy)

I. Chemical energy (c/ie/rcotrophy) Carbon source organic (heterotrophy) Chemolithoheterotrophy Chemoheterotrophy Carbon source inorganic (autotrophy) Chemolithoautotrophy Chemoorganoautotrophy II. Light energy (p/iototrophy) Carbon source organic (heterotrophy) Photolithoheterotrophy Photoheterotrophy Carbon source inorganic (autotrophy) Photolithoautotrophy Photoorganoautotrophy

Energy H;0 Energy H,S

Atmosphere t : Inorganic C 02 Organic C Inorganic C S /S04 Organic C t tr Litosphere v. / 1 Energy n20 Energy H2S

Figure 2 A diagram of the biological . The conversion from inorganic to organic carbon requires light or chemical energy and an electron donor (e.g., H20, H2S, Fe(II)), and is the process of autotrophy. The reverse reaction, in which organic carbon is oxidized to C02, releases energy while reducing an electron acceptor (e.g., Q2, SO|~/S°, Fe(III)). This part of the cycle is referred to as heterotrophy. some organisms do not fit neatly into this 8.08.2 AUTOTROPHIC METABOLISM classification scheme. Mixotrophic organisms From a microbial point of view, the carbon alternate between two or more distinct types of cycle is merely an energy cycle (Figure 2). metabolism, such as autotrophy and heterotrophy. Reduction of C0 through a variety of biochemi- In practice, microbiologists use abbreviated 2 cal pathways produces organic carbon, thereby terms because certain combinations of physiologi- changing the of carbon from +IV cal traits commonly occur together (Syliva et al., to between +111 and —IV. The main source of 1998). Because most organisms that metabolize energy that drives this process on Earth is organic compounds can extract electrons, energy, quantum energy or light, but there are ecosystems and carbon from them; they are referred to simply on Earth that are entirely dependent on chemo- as . Chemolithoautotrophs are often lithoautotrophy. Both forms of autotrophic metab- abbreviated as chemolithotrophs, , or olism are possible under anaerobic conditions. autotrophs because lithotrophic organisms are Here, we briefly consider the energy sources that usually also autotrophic are the ultimate source of reducing power for Microorganisms are also described by more . specific substrate requirements or environmental preferences. For example, organisms that require H are hydrogenotrophic (i.e., H -feeding) and 2 2 8.08.2.1 Phototroph (Photolithoautotrophy) those that thrive in very cold environments are Diversity and Metabolism psychrophilic (i.e., "cold-loving"). The most important aerobic organisms in oxic-anoxic Perhaps the most familiar photosynthetic pro- interfaces are , which thrive at karyotes are the (formerly blue- very low 02 concentrations. green ). These organisms use the Autotrophic Metabolism 323 to assimilate C02, releasing 02 as a waste product sulfur bacteria. In contrast, purple nonsulfur in the process (i.e., oxygenic photolithoautotrophy; bacteria typically thrive in less sulfidic, organic Table 3). Cyanobacteria include single-cell and rich habitats and are metabolically more diverse. filamentous pelagic and benthic species, and they For example, some are , which are found across a range of environmental use light as an energy source, but also require conditions from frigid Antarctic dry valleys to organic precursors to synthesize a portion of their mildly hot springs. The upper temperature limit of organic compounds. use the Calvin photosynthesis is 70-74 °C. The physiological cycle for autotrophic C02 fixation (Table 3). In versatility of this group is evident in the capacity of contrast, the are strict anae- some members to fix N2, to derive energy from robes that store S° only outside the cell and fermentation and to use HS~ as an electron donor predominantly use the reversed (through PS I) for photosynthesis. Cyanobacteria for . Some use ferrous iron as their are dominant species in microbial mats and photosynthetic electron donor (Heising et al., 1999; stromatolites and are believed to be responsible Widdel et al., 1993). Green nonsulfur bacteria for the development of an oxidized (02-containing) typically grow photoheterotrophically, but also atmosphere 2.2-2.0 Gyr ago (Des Marais, 2000). use H2 or H2S during autotrophic growth. The In addition, N2-fixing cyanobacteria are able to upper temperature of anoxygenic photosynthesis produce copious amounts of H2 during the dark is 70-73 °C. period, which may have had important impli- During photosynthesis, the energy from light cations for the development of an 02-rich atmos- quanta is converted into chemical energy that can phere (Hoehler et al., 2001). be used to drive biochemical reactions (Warburg Many do not produce 02 as a waste and Negelein, 1923). Photosynthetic organisms product. Such anoxygenic phototrophs are com- use a variety of light-harvesting systems, which prised of purple sulfur, purple nonsulfur, green collectively cover most of the visible spectrum sulfur, and green nonsulfur bacteria. Although (Stolz, 1991; Samsonoff and MacColl, 2001). are typically found in anoxic Eukaryotic phototrophic organisms contain either zones of and sediments, many are capable of a (Chla) or Chl£>, which have their photosynthesis under oxic conditions (Van Gemer- maximum absorption at 680 nm and 660 nm, den, 1993). Most fix N2 and store S° intra- or respectively. Among the , aerobic extracellularly, and some are capable of chemo- organisms such as cyanobacteria and prochlor- lithoautotrophic growth. Extreme halophilic, sulfi- ophytes typically contain Chla, while anoxygenic dic, and mildly thermophilic environments harbor phototrophs have a wide variety of chlorophyll

Table 3 Autotrophic reduction-oxidation reactions coupled to C02 assimilation. The most common carbon assimilation pathways for each type of metabolism are listed, with alternative modes of C02 fixation provided in parentheses. The free energy yield of the reaction is provided for pathways that do not require light. Metabolism Reaction^ AG01 C assimilation (kJ) pathway

H2 oxidation H2 + 0.5O2 — H20 -237 Calvin (RTCA) CO oxidation CO + 0.5O2 — C02 -257 Calvin S oxidation HS~ + 2Q2 -» SOT +H+ -798 Calvin (AcetylCoA) 2 - S oxidation 5HS~ + 8NOJT + 3H+ — 5S0 . + 4N2 + 4HzO - 3,723 Calvin NHj oxidation NH+ + 1.502 — N02 + 2HzO -287 Calvin (PEP) N02 oxidation NQ2 + 0.5O2 -» N03" -74 Calvin 2+ 3+ Fe(II) oxidation Fe + H++ 0.25O2 — Fe + 0.5H2O -33 Calvin 2+ Mn(II) oxidation Mn 0.5O2 + HzO — Mn02 + 2H+ -68 Calvin Anaerobic phototrophic Fe(II) 4FeCO, + l0H2O + hv Calvin oxidation (and denitrification) -» 4Fe(OH)3 + (CHzO) + 3HCO^ 4H2 + 2HC03" + H+ — CH3COO~ + 4H2Q -105 AcetylCoA Hydrogenotrophic 4H2 + HCO3" -» CH4 + 3H20 -136 AcetylCoA methanogenesis Oxygenic phototrophy COz + HzO + Az,-, (CH20) + 02 Calvin Anoxygenic phototrophy 2CQ2 + H2S + 2H20 + hv Calvin 2- + -» 2(CH2Q) + SO + H (RTCA, HPC) a hv — photon energy that is required to drive the reaction (hence no AG reported). b Calvin — Calvin cycle, RTCA — reductive or reversed tri-carboxylic acid cycle, PEP — Phosphoenolpyruvate carboxylation, and AcetylCoA — acetyl-CoA pathway, HPC — hydroxy-priopionate cycle. 324 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes pigments. The purple bacteria contain either whereas in anoxygenic photosynthesis, electron a Bchla (805 nm and 830- transfer is cyclic and only energy generation is 890 nm) or BChlb (835-850; 1,020-1,040 nm); driven by light. In hypersaline environments, green bacteria contain BChlc (745-755 nm), -based phototrophy is found in BChW (705-740 nm) or BChle (719-726 nm); some halobacteria (Hartman et ah, 1980) that can and contain BChlg (670 nm and grow in brine (32%, or 5.5 M NaCl). Phototrophic 788 nm). BChla is also found in some aerobic, growth in these red-pigmented (bacterioruberin, heterotrophic phototrophs (e.g., erythrobacters; an antenna pigment) microbes does not involve Yurkov and Van Gemerden, 1993) and anaerobic, chlorophyll and takes place under microaerophilic heterotrophic phototrophs (e.g., some purple conditions. The phototrophic growth potential is nonsulfur bacteria). A variety of antenna pig- extremely limited. Maximum bacteriorhodopsin ments, such as and phycobiliproteins, absorption is at 570 nm. cover the remaining windows in the visible spectrum. Oxygenic photosynthesis requires two photo- systems with different standard potentials and 8.08.2.2 Chemotroph (Chemolithoautotrophy) Diversity and Metabolism reaction center (Figure 3): PS II (E° = +1.0 V; P680), which uses H20 as elec- Inorganic redox reactions provide an alternative tron donor, and PS I (E° = +0.3 V; P700). In to using light as a source of energy and reducing contrast, anoxygenic photosynthesis utilizes only equivalents to assimilate C02 (Table 3). Electron a single photosystem: P870 (£° = +0.5V) in donors for chemolithotrophy include H2 (Bowien purple bacteria (Figure 3), P840 (E° = +0.3 V) in and Schlegel, 1981), (C02) green bacteria and P798 (E° = +0.2 V) in helio- (Shiba et ah, 1985), H2S and other reduced sulfur bacteria. Electron donors for anoxygenic photo- compounds, NH^, and other reduced nitrogen trophs include a range of reduced S-compounds, compounds, Fe(II), and Mn(II). It is likely that H2 and Fe(II). Oxygenic photosynthesis involves other reduced elements (e.g., As(IV), Cr(III), noncyclic electron transfer, light-driven energy Sb(III), Se(-II)/Se(0), U(-IV)) also function generation and light-driven reducing power, as reductants (Battaglia-Brunet et ah, 2002;

NADH

o -

+1 PS II

Figure 3 A schematic representation of photosystems and associated standard potentials involved in oxygenic photosynthesis in cyanobacteria (left panel) and anoxygenic photosynthesis (right panel). The flow of electrons in the former is linear, while in the latter it is cyclic. Decomposition and Fermentation 325 Dowdle and Oremland, 1998; Ehrlich, 1999, Many other carboxylation reactions exist 2002). Several of these pathways are discussed (Barton et ah, 1991). For example, in methylo- here in the context of the element cycles they trophic bacteria, and C02 are influence, including acetogenesis (Section combined to produce acetyl-CoA in the serine 8.08.3.2.1), hydrogenotrophic methanogenesis or hydroxypyruvate pathway. In contrast, the (Section 8.08.4.2), NHj oxidation (Section ribulose monophosphate cycle, which is another 8.08.5.3.3), Fe(II) oxidation (Section 8.08.6.5), methylotrophic pathway of formaldehyde fix- and sulfur oxidation (Section 8.08.7.9). ation, does not involve carboxylation steps. In addition to those described above, commonly found carboxylation reactions include those of pyruvate or phosphoenol pyruvate. In view of 8.08.2.3 Pathways of C02 Fixation several relatively recent discoveries of novel All microbes, including chemoorganohetero- C02 assimilation pathways (e.g., the hydoxypro- trophs, possess some ability to engage in revers- pionate cycle and anaerobic ammonium oxi- dation) and growing interest in deep-subsurface ible carboxylation (i.e., C02-C assimilation into an ) and decarboxylation , novel pathways of C02 incorpor- reactions, some of which lead to the incorporation ation may be discovered in the near future. of a significant amount of C02 (Wood, 1985). Here, we briefly consider the biochemical path- ways that photo- and chemolithotrophic bacteria 8.08.3 DECOMPOSITION deploy in order to produce the majority of their AND FERMENTATION . The four major C02-fixing pathways are the Calvin cycle, the acetyl-CoA pathway, the Most knowledge regarding the anaerobic reductive tricarboxylic acid (TCA) cycle, and the decomposition of organic materials in deposi- 3-hydroxypriopionate cycle. tional environments has been derived from studies The Calvin cycle or reductive pentose of terminal processes in the microbial , cycle occurs in all green plants and many particularly NO^, Fe(III), Mn(IV), and SO4" microorganisms. Carboxylation is catalyzed by reduction, and methanogenesis. These processes ribulose-bis-phosphate carboxylase-oxygenase are far easier to study than those responsible for (Rubisco). Rubisco also functions as an oxygenase polymer degradation. Because the degradation during photorespiration, but its affinity for 02 products of labile compounds eventually pass is 20-80 times lower than for C02. As with through a terminal step, it is generally considered most , Rubisco has a preference that terminal decomposition exerts a major control for lighter stable isotopes and C02 fixation results on the decomposition pathway by regulating 13 13 in the depletion of C (5 C) ranging from -10%0 intermediary metabolism and the rate of degra- to -20%„. dation (Conrad, 1999). However, it is clear that The acetyl-CoA or Ljungdahl-Wood pathway the overall rate of organic matter decomposition is is found in anaerobes, including methanogenic limited by the rate at which polymeric materials bacteria, acetogenic bacteria, and autotrophic are depolymerized. Much information about sulfate reducers. Two parallel pathways fix C02, decomposition can be provided by studies limited one of which results in an -bound carbonyl solely to intermediary or terminal metabolic group, the other one in an enzyme-bound methyl events. group. Combining the two ultimately yields the Organic matter deposited in sedimentary or acetyl-CoA pathway. The key enzyme in this wetland habitats is composed of a complex pathway is carbon monoxide dehydrogenase. In mixture of biopolymers. Some of these com- methanogens, the biosynthesis proceeds via the pounds, such as , , and acetyl-CoA pathway as well. Isotope fractionation are easily degraded by microorganisms (i.e., in the acetyl-CoA pathway yields S13C values labile), while other compounds, such as ranging from —20 to —40%c. and hemicellulose, are resistant to decomposition The reductive tricarboxylic acid cycle is (i.e., recalcitrant). Biopolymers are degraded in a basically the reverse of the oxidative tricarboxylic multistep process. First, microorganisms simplify acid cycle that heterotrophs use to generate polymers to monomers such as amino acids, fatty reducing equivalents (NADH, FADH2) that func- acids, and monosaccharides (Figure 4). The tion as electron donors for energy generation. The monomers are further mineralized to C02, or to 3-hydroxyproprionate cycle was found in the a combination of C02 and CH4. green nonsulfur bacterium Chloroflexus (Strauss Under aerobic conditions, the conversion of and Fuchs, 1993) and more recently also in some monomers to fully mineralized products is rather autotrophic Archea (Mendez et al., 1999). In this simple because 02-respiring bacteria can degrade pathway, 3-hydroxyproprionate is a key monomers completely to C02. However, under intermediate. anaerobic conditions this process requires 326 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

Polymers of organic matter in sediments was generally (e.g., polysaccharides, proteins) amino acids > neutral sugars > > lignin. This pattern was independent Exo-enzymes of the 02 content and of sediments. Seasonal variations in decomposition are Monomers affected strongly by temperature, but temporal (e.g., monosaccharides, changes in organic matter deposition also affect amino acids) the degradation rates in sediments. Evidence of / i \ this is the observation that carbon mineralization Primary fermentation varies strongly across seasons in sediments that are uniformly cold year-around because of seasonal variation in plankton production (Schulz and Conrad, 1995). It is clear that polymer breakdown is often the rate limiting step in organic matter degradation (Glissmann and Conrad, 2002; Reineke, 2001; Wu et ah, 2001). Mineralization of organic macromolecules is initiated by extracellular enzymes because bacteria are unable to hydrolyze substrates that are much larger than about 600 Da (Weiss et ah, Figure 4 Metabolic scheme for the degradation of 1991). Not all bacteria are capable of synthesizing complex organic matter, culminating in methanogen- these enzymes, as is often the case with those esis. Polymers are cleaved via extracellular or cell- responsible for terminal decomposition and some surface associated enzymes to monomers that are intermediary metabolisms. As a result, these fermented to organic products, H2 and C02. Methane terminal organisms depend heavily on the activi- is formed primarily from the oxidation of H2 coupled to ties of other bacteria for substrates. It is clear that C02 reduction or by the fermentation of acetate. Acetate polymer hydrolysis occurs since these compounds is formed by primary fermentation, acetogenesis from are required to support microbial activities in H2/C02, and from secondary fermentation of primary sediments, but some studies have failed to detect fermentation products. polymer hydrolysis potentials sufficient to support in situ rates of metabolism (Arnosti, 1998). Such studies underscore the difficulties of examining a consortium of bacteria that degrade monomers hydrolytic processes. in a series of steps. The first step is primary The approaches used in studies of polymer fermentation to low molecular weight products hydrolysis include additions of intact plant such as and volatile fatty acids. Next, material to incubating sediments (Battersby and primary fermentation products are either miner- Brown, 1982; Wainwright, 1981), inhibition of alized to C02 and CH4, or they undergo secondary consumption with and fermentation to smaller volatile fatty acids. measurement of carbohydrate increases (Boschker Finally, the secondary fermentation products are et ah, 1995), and the use of labeled materials. mineralized by respiratory organisms using inor- The latter approach usually involves esti- ganic terminal electron acceptors, a process that mating the hydrolytic potential of microbial yields C02, or C02 and CH4. Secondary fermen- communities by amending sediments with ffuor- tation is prevalent under methanogenic conditions. escently labeled precursors that serve as proxy organic macromolecules. A particularly common Huorophore is methylumbelliferyl (MUF) 8.08.3.1 Polymer Degradation attached to a monomer such as a monosaccharide or an (Boetius and Lochte, 1994; The quality of organic matter degraded in Boschker and Cappenberg, 1994; Hoppe, 1983; anaerobic environments varies depending on its King, 1986; Meyer-Reil, 1986). When these labels origin with a gradient of structural complexity are attached to a monomer, they may fail to occurring from (labile) to vascular discern the activities of true extracellular enzymes plants (recalcitrant) (Wetzel, 1992). Vascular because they are small enough to enter the plant detritus is resistant to decomposition bacterial periplasmic space. Thus, labeled com- because of an abundance of high-molecular- pounds may be poor proxies of large-molecular- weight structural compounds such as lignocellu- weight materials (Arnosti, 1998; Martinez and loses and complex polysaccharides (Benner et ah, Azam, 1993). In some cases, hydrolytic poten- 1991), while phytoplankton cells are composed tial has been assayed using whole polymers largely of carbohydrates (Benner et ah, 1992). that were ffuorescently labeled (Arnosti, 1996) Cowie and Hedges (1993) noted that the reactivity or specific polymers have been introduced Decomposition and Fermentation 327 into growing cultures of isolated bacteria 8.08.3.1.2 Lignin (Reichardt, 1988). Organic matter deposited in near-shore marine and most freshwater habitats is composed primar- ily (—75% by weight) of lignocellulose, a complex 8.08.3.1.1 Polysaccharides mixture of lignin and the polysaccharides and hemicellulose (Benner et al., 1985). Lignin is a Studies of polymer use have often focused on the unique phenolic polymer of nonrepeating units hydrolysis of polysaccharides because carbo- that makes up 25-30% of the biopolymers in hydrates make up a large portion of phytoplankton vascular plants and is second only to cellulose biomass (Parsons et al., 1961) and cellulose is the as the most abundant organic carbon source in main polysaccharide in terrestrial ecosystems the . It is highly resistant to microbial (Watanabe et al, 1993; Glissman and Conrad, degradation (Kawakami, 1989) and its association 2002). In general, hydrolytic activity in sediments with cellulose and hemicellulose polysaccharides decreases rapidly with depth in parallel to declin- imparts degradation resistance to these polymers ing organic matter reactivity and general bacterial as well (Crawford, 1981). Hence, lignin is widely metabolic activity (King, 1986). However, Arnosti distributed in depositional environments such as (1998) found that the potential for hydrolysis of soils and peats (Hedges and Oades, 1997; Miyajima algal-derived polysaccharides, such as pullulan et al, 1997; Tsutsuki et al, 1994; Yavitt et al, and laminarin, was uniformly rapid throughout the 1997), riverine sediments (Hedges et al, 1986, upper —11 cm of Arctic sediments. Additions of 2000; Meyers et al., 1995), and coastal marine metabolic inhibitors failed to completely arrest sediments (Dittmar and Lara, 2001; Gough et al, hydrolytic activity, indicating that a significant 1993; Hedges et al, 1997; Miltner and Emeis, portion of hydrolytic enzymes in sediments are 2001). Lignin's aromatic character makes it the either free in pore or attached to particles major source of naturally occurring aromatic and active in lieu of bacterial metabolism (Arnosti, compounds. 1998; Boschker et al, 1995). Although hydrolytic rate optima have been shown to correlate with The degradation of lignocelluloses in detritus- environmental factors such as pH or salinity (King, based ecosystems like wetlands is crucial to 1986), they often exhibit temperature optima that maintain carbon balance since macrophytes are greatly exceed ambient temperatures (King, 1986; composed of 50-80% lignocellulose (Maccubbin Mayer, 1989; Reichardt, 1988). For example, and Hodson, 1980). Early studies suggested that extracellular enzyme activity in -ice bacterial the lignin polymer was essentially inert in the communities exhibited temperature optima near absence of oxygen (Hackett etal., 1977). However, 15 °C and a psychrophilic isolate yielded a protease it has since been shown that the complex lignin extract with an optimum activity at 20 °C (Huston polymer can undergo anaerobic degradation ef a/., 2000). (Benner et al, 1985, 1986), but the process can Relatively slow decomposition in anaerobic be rather slow and tends to yield unmetabolized environments may partly be due to depressed products (Colberg and Young, 1982; Young and extracellular enzyme activity caused by low pH, Frazer, 1987). Anaerobic loss of lignin polymers is —3-30% as complete as aerobic degradation of low 02 concentrations, or other factors (Kang and Freeman, 1999). Freeman et al. (2001) determined the same polymers (Benner et al, 1984); the that depressed extracellular enzyme activity in a cellulose component is more labile than the lignin peatland soil was due to the high content of component. phenolic compounds, and concluded that phenols Both fungi and bacteria degrade lignocellu- were not degraded because of an 02 limitation on loses. Fungi tend to dominate decomposition in phenol oxidase activity. This led them to speculate upland soils (Orth et al, 1993; Witkamp and that the vast pool of organic carbon currently Ausmus, 1976), whereas bacteria dominate in sequestered in peatland soils (one-third of all soil most aquatic environments (Benner et al, 1986). carbon) is under the control of a single enzymatic The latter authors noted that bacteria were "latch," the activity of which could increase responsible for the degradation of lignin and dramatically if drought or drainage were to associated polysaccharides in marine wetlands. increase 02 availability. In other cases, extra- also contributed significantly to lignin cellular hydrolase activities may not be affected degradation in an acidic freshwater wetland, and by environmental factors such as 02, sulfide, or eukaryotes degraded both lignin and polysacchar- iron chemistry (King, 1986), and do not necess- ides in a slightly alkaline freshwater wetland. The arily follow substrate concentrations or microbial polysaccharide component is usually decomposed biomass on an annual basis (Mayer, 1989). several fold more rapidly than lignin, and the Because microbial activity is consistently affected lignocellulose of herbaceous plants decomposes by such factors, these findings reflect the extra- more readily than that of woody plants (Benner cellular nature of the hydrolytic enzymes. et al, 1985). The aerobic catabolism of lignin by 328 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes fungi and filamentous bacteria utilizes ligninolytic xenobiotic aromatics such as chlorinated peroxidases (Black and Reddy, 1991; Eriksson pollutants. et ah, 1990), that often require manganese to be active (Brown et al, 1990; Gettemy et al., 1998). Less is known of the enzymes involved in the anaerobic decomposition process. 8.08.3.2 Fermentation Under anaerobic conditions, lignin oligomers Fermentation is a metabolic process in which can be depolymerized to monomers, and lignin organic compounds serve as both electron donors monomers can be mineralized to C02 (Colberg, and acceptors. Primary fermentation is the exer- 1988; Colberg and Young, 1982; Young and gonic breakdown of glucose and other monomers Frazer, 1987). The degradation of the lignin to products such as alcohols, fatty acids, H2, and polymers releases aromatic subunits, and many C02. Secondary fermentation further converts studies have examined the anaerobic pathways by these primary products to acetate and other low- which these monomers are used. In fact, lignin molecular weight organic acids (Figure 4). Fer- monomers have been used often as lignin model mentation differs dramatically from anaerobic compounds (Pareek et al., 2001; Phelps and respiration because it occurs inside the cell, and Young, 1997). Although it is clear that aromatic energy is generated by organic matter dismutation rings are readily cleaved aerobically by dioxy- and substrate-level . By compari- genase and peroxidase enzymes, these reactions son, aerobic and anaerobic respiration requires an do not occur anaerobically; aromatic rings are external electron acceptor and they proceed by cleaved in the absence of 02 via a reductive oxidative phosphorylation via electron transport. mechanism whereby hydrogenation of the aro- Fermentation generates very little energy relative matic ring nucleus results in a cyclohexane to aerobic or anaerobic respiration. However, it is derivative, the ring of which is then opened a key component of the anaerobic mineralization (Reineke, 2001; Young and Frazer, 1987). Lignin process because most nonfermentative organisms monomers generally contain hydroxyl, methoxyl, cannot use the typical monomers released during and/or carboxyl groups that are removed from polymer hydrolysis. As a result, fermenters can aromatic rings prior to reduction and cleavage of greatly outnumber bacteria dependent on terminal the ring. The O-demethylation (demethyoxyla- respiration processes in some environments. For example, in SO^-reducing sediments, the tion) of phenylmethylethers has received con- - siderable attention since the methyl product can SO4 reducers typically account for —5% of the serve as a C\ substrate (i.e., not C-C bonds) for all bacteria present, while most bacteria are bacterial growth (Evans and Fuchs, 1988; Young involved in polymer hydrolysis and fermentation and Frazer, 1987). Many of the O-demethylating (Devereux et al, 1996). The situation is different strains of bacteria are acetogenic (Section in anaerobic wetland soils, where plant roots release acids and alcohols that can be used directly 8.08.3.2.1) (Frazer, 1995; Kreft and Schink, by SO4 reducers. In this case, the SO4 reducer 1993; Kreft and Schink, 1997; Kiisel et al, 2000; populations on roots can account for over 30% of Wu et al., 1988), but several other types of the total microbial because fermenta- anaerobic and facultatively anaerobic bacteria tion is not required (Hines et al, 1999; Rooney- have the ability to demethoxylate aromatic rings, Varga gf oA, 1997). including sulfate reducers, nitrate reducers, fer- In pure culture, fermenting bacteria consume a mentative bacteria, and other strains involved in large variety of organic compounds. The list anaerobic syntrophy (Section 8.08.3.2.2) (Cocaign includes sugars, amino acids, purines, pyrimidines, et al., 1991; Krumholz and Bryant, 1985; Liu and some aromatics, acetylene, and a broad range of Suflita, 1993; Mountfort et al, 1988; Phelps and organic acids (Ci-Ci8). They produce a wide Young, 1997; Young and Frazer, 1987). Some spectrum of fermentation products, but Ci-C18 pure cultures are capable of removing a variety of acids and alcohols, H2, and C02 are most prevalent functional groups from aromatic rings (Kiisel (Schink and Starns, 2002). In methanogenic et al., 2000). In many instances, the aromatic ring habitats, all fatty acids longer than two , is not cleaved after these groups are removed branched-chain and aromatic fatty acids, and all (Young and Frazer, 1987). During O-demethyox- alcohols longer than one carbon require secondary ylation, the released methyl group can be used to fermentation prior to use by methanogenic bacteria methylate sulfide, which leads to the production of (Schmitz etal, 2001). Hence, secondary fermenta- the gases methane thiol and dimethylsulfide; the tion is essential for complete mineralization to remaining aromatic compound is not further occur during methanogenesis. However, SO^-- degraded (Bak et al., 1992; Finster et al, 1990, and metal-reducing bacteria are more metaboli- 1992). Bacteria capable of degrading plant- cally versatile than methanogens and are capable of derived aromatic compounds such as ferulic and mineralizing most primary fermentation products syringic acids are often capable of degrading directly. Organic mineralization can proceed via Decomposition and Fermentation 329 _ a two-step process in metal and S04 -rich The autotrophic formation of acetate from environments like marine sediments where pri- H2/C02 can directly compete with autotrophic mary fermenters hydrolyze polymers and ferment methanogenesis. Methane formation from H2/C02 monomers while metal and S>0\~ reducers utilize is thermodynamically more favorable than the fermentation products (Widdel, 1988). How- acetogenesis, but the latter pathway is dominant ever, if fatty acid oxidizing SO4" reducers are in some habitats. Kusel and Drake (1994) noted absent, then mineralization of fatty acids is possible that flooded soils accumulated acetate for several via secondary fermentation and the use of H2 by months without becoming methanogenic. Acetate SC>4_ -reducing bacteria (Monetti and Scranton, also accumulates transiently in sediments and 1992). The SO|~-reducing community does exhi- wetland peat soils prior to its use by methanogens, bit some other cooperative activities (Section suggesting that acetoclastic methanogens grow 8.08.7.2.2). more slowly than H2/C02 utilizers or are more sensitive to 02 (Avery et al., 1999; Crill and Martens, 1986; Sansone and Martens, 1982; 8.08.3.2.1 Acetogenesis Shannon and White, 1996). In addition, it appears Besides the "classical" fermentation described that acetate accumulates all season in colder high above, organic monomers can also be degraded to latitude wetlands (Duddleston et al., 2002; Hines acetate via acetogenesis. This term is somewhat et al., 2001), but it is unknown what portion of this nebulous and has been used to describe any acetate is due to fermentation or acetogenesis. In reaction leading to acetate, including fermentation. rice paddy soils, up to 40% of the fatty acids can However, a widely accepted definition holds that be derived from C02 reduction, including vir- are obligately anaerobic bacteria that use tually all of the acetate formed (Conrad and Klose, the acetyl-CoA pathway both for the reductive 2000). The bacteria physically associated with rice roots exhibited diverse anaerobic metab- synthesis of acetyl-CoA from C02, and as a terminal electron-accepting and energy conserving olisms, including the coexistence of acetogenesis process (Drake, 1994). If acetate is usually the and methanogenesis from C02 reduction (Conrad sole reduced end product, the is con- and Klose, 1999; Liesack et al, 2000). This sidered to be a homoacetogen. The acetyl-CoA apparent relief of competition may have been due pathway is responsible for the anaerobic dark to the abundance of substrates released by roots (Conrad and Klose, 1999; Liesack et al, 2000). fixation of C02 into organic compounds, and most acetogens are able to use this pathway to reduce CO2 during the oxidation of H2 to form acetate autotrophically. In a typical acetate fermentation 8.08.3.2.2 Syntrophy and interspecies reaction, glucose is converted to two moles of hydrogen transfer acetate and two moles of C02. However, acetogens can reduce the two C02 molecules to acetate Many anaerobic microorganisms require sec- yielding a total of three moles of acetate. It is ondary fermentation products such as acetate for theoretically possible for acetogenic catabolism of electron donors. However, the ability of bacteria to glucose to acetate, followed by its degradation to perform secondary fermentation can be limited if CH4 and C02, to be a major degradative pathway in H2 (a metabolic waste product of the process) anaerobic environments because it is highly accumulates, and causes the reaction to become favorable thermodynamically. However, this joint endergonic. For example, fermentation of organic pathway does not seem to be important in situ, and acids and alcohols such as butyrate, propionate and the role of acetogenesis in methanogenic habitats does not yield sufficient energy to support remains unclear (Conrad, 1999). growth if the H2 concentration rises above Acetogenic bacteria are a very diverse group 10-4 atm. Yet, these fermentation reactions are and display a wide range of catabolic capabilities generally exergonic in situ, because H2 is con- including the utilization of sugars, acids, alcohols, tinually removed by H2-consuming bacteria lignin-derived aromatic methoxyl groups, Ci and (Table 4). As a result, H2 has a short turnover C2 compounds, CO, and methyl halides (Drake, time and usually occurs at very low concentrations 1994). In addition, acetogens have been isolated (Conrad et al., 1986,1989) even though it is an that are capable of reverse acetogenesis where common intermediate in metabolism (Conrad, acetate is converted to C02 and H2 (Zinder and 1999). The exchange of H2 makes these organisms Koch, 1984). The acetyl-CoA pathway is also metabolic partners, and allows primary fermenta- present in some SO4 -reducing and methanogenic tion products to be completely mineralized in bacteria and is involved in acetate oxidation, anaerobic habitats. Such syntrophic relationships disproportionation (i.e., CH* production from (Biebl and Pfennig, 1978) are a symbiotic acetate), autotrophic fixation of C02, CO oxi- cooperation between two metabolically different dation, and the assimilation of C: compounds bacteria that depend on each other for the (Fuchs, 1994). degradation of a substrate, typically for energetic 330 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes Table 4 Examples of reactions occurring in methanogenic environments illustrating the effect on energy yield of the consumption of fermentation products. Maintenance of low reactant concentrations allows secondary fermentation reactions that are endergonic under standard conditions to be exergonic (negative AG).

Reaction Free- •energy char ige (kJ) 4G"» AGh

Glucose + 4H20 -• 2 acetate" + 2HCO^ + 4H+ + 4H2 -207 -319 + Glucose + 2H20 -• butyrate" + 2HCO^ + 3H + 2H2 -135 -284 + Butyrate" + 2H20 — 2 acetate" + H + 2H2 + 48.2 -17.6 Propionate" + 3H20 — acetate" + HCO^ + H+ + H2 + 76.2 -5.5 2Ethanol + 2H2Q -» 2 acetate" + 2H+ + 4H2 + 19.4 -37 Benzoate + 6H20 -» 3 acetate" + 2H+ + C02 + 3H2 + 47 -18 Source: Zinder (1984). a Standard conditions: solutes, 1 M; gases, 1 atm. Concentrations of reactants typical of anaerobic habitats: fatty acids, 1 mM; glucose, 4 10 uJM; CH4, 0.6 atm; H2, 10" atm; HCO^; 20 mM.

reasons. In many cases, H2 is the compound H2-scavenging methanogen because the reaction transferred to the terminal organism and this is endergonic at the H2 concentrations encoun- process is known as interspecies H2 transfer. tered. However, the reaction is exergonic if the H2 3 Although syntrophy and interspecies H2 transfer partial pressure is maintained at <10 atm by are synonymous in many instances, it should be the methanogen with a net co-culture. noted that H2 transfer is a mechanism of electron Primary fermenting bacteria can also profit from transfer and other electron carriers can be involved. the activities of H2-consuming partners at the end of Because formate is transferred between organisms the degradation chain. The maintenance of low H2 similarly to H2, the term interspecies H2 transfer partial pressures allows primary fermentation to often refers to both formate and H2 transfer (Boone favor the production of H2 and more oxidized end etal., 1989; Thiele and Zeikus, 1988). However, in products like acetate and C02. This permits some instances formate transfer is insignificant fermentation to be more efficient in general, and compared to H2 transfer and vice versa (Schmidt leads to additional ATP synthesis (Thauer et ai., and Ahring, 1995). In addition, electron transfer 1977; Schink and Stams, 2002; Schmitz et al, can occur via other electron carriers such as acetate 2001). The H2-utilizing population is the primary (Dong et al, 1994), or amino acids such as regulator in the degradation process because its cysteine (Cord-Ruwisch et al., 1998). presence dictates which pathways are energetically The classical example of interspecies H2 capable of proceeding (Bryant, 1979; Zeikus, transfer was recognized when a presumably pure 1977). In an active anaerobic community, the ethanol-consuming methanogenic culture, Metha- presence of a microbial H sink allows electron and nobacillus omelianskii (Barker, 1940), was found 2 carbon flow to bypass much of the secondary to actually be a co-culture of an ethanol utilizing fermentation process, except for the use of long- syntroph (S Strain) and an H -consuming metha- 2 chained or branched intermediates produced nogen (Strain M.o.H.) (Bryant et al, 1967): during and amino acid (Schink, Strain S 1997). Therefore, it is conceivable that secondary fermentation could be a rather small portion of 2CH CH OH + 2H 0 3 2 2 anaerobic metabolism if primary fermentation — 2CH3COCT + 2H+ + 4H2 yielded only acetate as the organic product. However, many natural habitats are not as micro- AG '=+19 kJ mol" 0 bially active as those encountered in culture, and Strain M.o.H. : secondary fermentation tends to be an important, 4H2 + C02 — CH4 + 2H20 but rather poorly understood, component of natural AGo' 131 kJ mol anaerobic environments (Conrad, 1999). Since its discovery around the mid-1960s, a variety of interspecies H2 transfer reactions have Co-culture : been elucidated. Early studies focused primarily on processes involving methanogenic bacteria CH CH OH + C0 — 2CH COO + 2H+ + CH 3 2 2 3 4 because this association was the first described. AGr/ 112kJmol The small amount of energy available in metha- nogenic metabolism effectively forces the bacteria The bacterium responsible for ethanol fermen- into symbiotic relationships that neither partner tation will not grow using ethanol without the can function without (Schink, 1997). Although Decomposition and Fermentation 331 the original ethanol-utilizing partner (S Strain) of bacterium. Microscopic studies have shown that the co-culture comprising "M. omelianskii" was syntrophic associations between fatty acid-oxidi- lost, other syntrophic ethanol-oxidizing bacteria zers and methanogens can produce structured or have been isolated (Ben-Bassat et ah, 1981; layered microcolonies composed of propionate Schink, 1984). The suite of H2-releasing com- utilizers and H2-consuming methanogens, and pounds used by syntrophs includes alcohols these structures can be surrounded by acetate- (Bryant et al, 1967; Eichler and Schink, 1986; utilizing methanogens and other bacteria Schink et al., 1985), fatty acids (Mclnerney et al., (Harmsen et al., 1996b). Finding an organized 1981, 1979; Schink, 1985b; Schink and Friedrich, juxtaposition of bacteria is not surprising con- 1994; Zinder and Koch, 1984), aromatic com- sidering the physically close association required pounds (Doffing and Tiedje, 1991; Elshahed et al., for the transfer of products between syntrophic 2001; Knoll and Winter, 1989; Mountfort et al, partners. Further work has shown that acetate 1984), glycolic acid (Friedrich et al., 1991), and and H2-consuming methanogens tend to segregate amino acids (Nagase and Matsuo, 1982; Nanninga into micro-clusters as well (Gonzalez- Gil et al., and Gottschal, 1985; Stams and Hansen, 1984; 2001; Rocheleau et al, 1999). Winter et al, 1987). Anaerobic bacteria can ferment a wide range of Many syntrophic microorganisms can grow in amino acids and many do so using interspecies the absence of a H2-utilizing partner by using H2 transfer in association with H2 utilizers. The slightly more oxidized substrates in a dismuta- classic mode of amino acid fermentation is the tion fermentation reaction (Eichler and Schink, Strickland reaction in which a pair of amino acids 1986; Elshahed and Mclnerney, 2001a; Schink, is consumed, one acting as an oxidant and the 1997). For example, a bacterium that can only other as a reductant. For example, a single degrade propionate when in co-culture with an bacterium can oxidize alanine to acetate and H2-consuming partner can be grown in pure NKt while reducing glycine to acetate and NH4": culture on pyruvate (Wallrabenstein et al., 1994), and an ethanol-utilizing syntroph can be grown CH3CH(NIT})COCr (alanine) f alone on acetaldehyde analogs like acetylene + 2CH2(NH3 )COO~ (glycine) (Schink, 1985a). The need for physical or — 3NH| + 3CH3COO + C02 chemical removal of H2 allows for only limited growth of pure culture syntrophs (Mountfort and However, this reaction can be uncoupled in which Kaspar, 1986). one bacterium oxidizes alanine to acetate, C02, Because the degradation of fatty acids to NH4, and H2, and a second bacterium uses the H2 acetate and H2 is more endergonic than ethanol to reduce glycine to acetate and NH^: oxidation, the H2 partial pressure must be ~1 atm lower for fatty acids to be degraded CH3CH(NK|-)COCr + 2H20 — CH3COCT under methanogenic conditions than for ethanol + NH j + C0 + 2H degradation (Schink, 1997; Wallrabenstein et al., 2 2 1 1994). The energy yield of fatty acid oxidation is AG0' = +2.7kJmor improved when both acetate and H2 are being removed, thereby "pulling" the reactions to f CH (NH )COCT + H — CH3COO" + NH^ completion. This has been demonstrated in tri- 2 3 2 cultures that degraded either butyrate (Ahring AG0'= -78kJmor' and Westermann, 1988) or propionate (Dong et al., 1994). Fatty acid degradation by a Adding these equations, we obtain fermenter was enhanced when one methanogenic f f strain consumed acetate while the other con- CH3CH(NH3 )COO + 2CH2(NH3 )COO sumed H2. Thus, the combination of low acetate + 2H20 — 3CH3COO + 3NH| + C02 and low H2 concentrations improved degradation efficiency via syntrophy. It should be mentioned AG/ 153 kJ mol that acetate can also be consumed indirectly by methanogens. This occurs when acetate is first A single bacterium has been isolated that can converted to H2 and C02, both of which are then conduct all three reactions (Zindel et al, 1988). consumed by a hydrogenotrophic methanogen However, H2 and acetate generated from alanine (Zinder and Koch, 1984). oxidation can be consumed by other partners such All known propionate oxidizers are also as methanogens (Nagase and Matsuo, 1982), capable of reducing SO4", and the biochemical SO4- reducers (Nanninga and Gottschal, 1985), components may include part of the SO|~- or acetogens (Zindel et al, 1988). The Strickland reducing apparatus (Schink, 1997). Syntrophic reaction is energetically favorable over the propionate use occurs when SO4" is limiting and syntrophic fermentation of amino acids (Schink, involves the transfer of H2 to a methanogenic 1997) and the former may dominate in habitats 332 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes rich in amino acids. However, syntrophic path- with other fatty acid fermenting species (Monetti ways may be important where amino acid and Scranton, 1992). Metal-reducing bacteria such concentrations are low and methanogenesis is as sp. are also capable of degrading fatty active. A wide variety of bacteria are capable of acids like acetate by interspecies H2 transfer to fermenting various amino acids syntrophically nitrate reducing bacteria (Cord-Ruwisch et al., using methanogens as H2 with acetate, 1998). It appears that the transfer of electrons to the propionate, and butyrate as common end products nitrate-reducing partner by the Geobacter sp. (Baena et al., 2000, 1999; Meijer et al, 1999; occurs using cysteine as the electron-transferring Starns and Hansen, 1984; Wildenauer and Winter, agent (Kaden et al., 2002). 1986; Winter et al., 1987). The ratio of organic Fermenting bacteria can also transfer electrons products formed is a function of the H2 partial to oxidized humic acids, allowing for the for- pressure, which underscores the importance of the mation of more oxidized fermentation products H2 in controlling electron and carbon than those produced in the absence of humics flow (Stams and Hansen, 1984). (Benz et al., 1998). It was shown previously that The role of the H2 scavenger in syntrophic Fe(III)-respiring bacteria can transfer reducing relationships can be accomplished by a variety of equivalents from acetate to various humic acid H2-utilizing anaerobes. Methanogens have preparations including the quinoid model com- received the most attention due to the importance pound 2,4-anthraquinone disulfonate (AQDS) of syntrophy in methanogenic habitats since CH4- (Lovley et al, 1996b, 1998). Humic acids act producing bacteria are extremely limited in the catalytically as electron shuttles by transferring scope of electron donors used. Anaerobic, S04- electrons to oxidized iron. It has also been dependent CH4 oxidation is the most recent shown that fermenting bacteria can transfer example of microbial syntrophy between a metha- electrons to Fe(III) in this manner, but it nogen-like organism and a S04~ reducer (Section appears that these bacteria do not conserve 8.08.4.5). However, this role can be replaced by energy via electron transport like iron-respiring several other physiologic groups including S04"-, bacteria do (Benz et al., 1998). However, the S0-, metal-, nitrate-, glycine-, and fumarate-redu- electron sink serves a similar function as a cing, and acetogenic bacteria (Schink, 1997). The syntrophic partner. For example, propionate ability of an H2-scavenging reaction to compete for fermentation is endergonic in the absence of an transferred electrons depends on the redox poten- H2-consuming partner, but propionate can be tial of the terminal electron acceptor with nitrate fermented exergonically to acetate without a and fumarate > S04~ > C02/CH4 > C02/acet- bacterial partner when electrons are transferred ate (Cord-Ruwisch et al, 1988). to humic acids (Benz et al., 1998). Respiring bacteria resort to syntrophy to decom- pose low molecular weight compounds when exogenous oxidant levels are low. The most widely 8.08.4 METHANE studied are S04" reducers that utilize fermentation processes coupled to H2 transfer to methanogens Methanogenesis is the final step in the anaero- when S04 is depleted (Bryant et al., 1977; bic degradation of organic carbon. The principal Cord-Ruwisch et al., 1986; Harmsen et al., steps performed by methanogens are fermentation 1996a; van Kuijk and Stams, 1995; Wallrabenstein of acetate to C02 and CHj, and oxidation of H2 to etal., 1995). Interestingly, S04 -reducing bacteria H20. In both cases, the waste product, CH4, still can fill an opposite role in which H2 produced holds potential energy in the form of reducing during the degradation of and acetate by a equivalents that can support additional anaerobic methanogenic bacterium is consumed by a S04- metabolism. Thus, the true end point of anaerobic reducing bacterium during S04 reduction (Phelps organic matter degradation in some ecosystems etal., 1985).ThisS04 -dependentinterspeciesH2 is the anaerobic oxidation of CH4 to C02 transfer results in a more complete oxidation of (Section 8.08.4.5). In the presence of 02, methane organic carbon substrates yielding more C02 and is a source of energy for CH4-oxidizing bacteria less CH* than the methanogen would produce (i.e., ), some of which are sym- alone (Achtnich et ah, 1995b). Hence, although bionts of deep-sea mussels (Childress et al., S04 reducers and CH4 producers often compete 1986; Fiala-Medioni et al, 2002; Kochevar for substrates, this competition is circumvented in ef a/., 1992). the absence of S04~ when fermenting S04~ The many environmental and economic issues reducers rely on methanogens as H2 sinks, or that involve CH* have stimulated research in all when methanogens use S04~ reducers as H2 sinks. aspects of methane cycling. Here, we consider the Studies using sediment slurries revealed a dual role processes that influence CH4 production for S04~ reducers in which they act as direct and oxidation. The literature on CH4 emissions is consumers of fatty acids as well as affecting fatty considered only briefly in this article acid degradation by removing H2 during syntrophy (Section 8.08.4.7). Methane 333 8.08.4.1 Methane in the Environment Table 5 Estimated sources and sinks of methane in the l2 -1 atmosphere in units of 10 g CH4 yr . Human activity has increased the atmospheric _1 CH4 concentration from ~0.7|JLLL to Range Likely 1.8 |xL L_1 (ppmv) since about the 1850s, primarily by stimulating methanogenesis in soils Sources Natural and sediments (Lelieveld et al, 1998; see below). Wetlands Methane concentrations are currently double the Tropics 30-80 65 highest level recorded in a 420,000-year ice core Northern latitude 20-60 40 (Petit et al, 1999), and it accounts for 20% of Others 5-15 10 human-induced radiative forcing (Prather et al, Termites 10-50 20 2001). Rising CH4 concentrations during periods Ocean 5-50 10 of increased solar insolation and interglacial Freshwater 1-25 5 warming have substantially amplified global Geological 5-15 10 Total 160 warming in the past (Petit et al., 1999), and there is concern that this will occur in the future. Anthropogenic In the past, changes in the atmospheric CH4 Fossil fuel related content were accompanied by changes in the area Coal mines 15-45 30 of northern and tropical wetlands (Blunier et al., Natural gas 25-50 40 Petroleum industry 5-30 15 1995; Chappellaz et al, 1993). Because CH4 Coal combustion 5-30 15 hydrate deposits may have been a large and Waste management system sudden source of atmospheric CH4 that contrib- Landfills 20-70 40 uted to past climate change (Nisbet, 1992; Thorpe Animal waste 20-30 25 etal., 1996), the current stability of these vast CH4 Domestic sewage treatment 15-80 25 reservoirs is a topic of considerable importance Enteric fermentation 65-100 85 (Kvenvolden, 1999; Wood et al, 2002). Biomass burning 20-80 40 Rice paddies 20-100 60 About 70% of the current CH4 sources are anthropogenic, with roughly equal contribu- Total 375 tions from fossil fuel-related industries, waste Total Sources 535 management systems, and enteric fermentation Sinks associated with raising livestock (Table 5). Of the Reaction with OH 330-560 445 natural sources, wetlands are 70% of the total and Removal in stratosphere 25-55 40 therefore a major research focus. About 60% of Removal by soils 15-45 30 natural wetland sources and most rice paddies occur in the tropical latitudes. Another 35% are in Total Sinks 515 northern latitudes, and 10% are in mid-latitudes. Atmospheric increase 30-35 30 The combined contribution of natural and mana- Source : Schlesinger (1997). ged wetlands to global CH4 emissions is —32% at present, and perhaps 70% of all sources before the industrial revolution (Lelieveld et al., 1998). a clean-burning energy source and the major Estuaries are <9% of ocean CH4 sources carbon substrate for alkane formation in economic (Middelburg et al, 2002). gas reservoirs (Sherwood Lollar et al, 2002). The dynamics of CH4 in the atmosphere are Methane represents an unwelcome loss of meta- quite different than C02. It is a reactive gas and bolic energy from flatulent livestock (Johnson participates in atmospheric chemical reactions that et al., 1993). Although methanogens do not influence the concentrations of NO, NO%, CO, and directly degrade organic pollutants, they partici- 03 (Crutzen, 1995). On a molar basis, CH4 is 3-22 pate in the process by consuming fermentation times stronger as a than C02, products, thereby maintaining an environment that depending on the period of time over which its is thermodynamically conducive to the anaerobic impact is considered (Rodhe, 1990; Whiting and degradation of multi-carbon compounds (Lovley, Chanton, 2001). Methane concentrations are more 2000a; Weiner and Lovley, 1998; Section responsive than C02 to changes in sources or sinks 8.08.3.2.2). It has been demonstrated that metha- because of a far shorter atmospheric residence time nogens play this role in the degradation of alkanes, for CH4 (—10 yr versus >100yr). These charac- a group of particularly stable that are teristics have inspired the recommendation that abundant in fossil fuel deposits (Anderson and efforts to slow the pace of global warming Lovley, 2000; Zengler et al, 1999). Aerobic should focus initially on abating CH4 emissions bacteria that depend entirely on CH4 for their (Hansen et al, 2000; Fuglestvedt et al, 2000). carbon and energy (i.e., methanotrophs) directly Methane is an important compound econo- degrade organic pollutants such as trichloroethy- mically. In the form of natural gas, it is lene (TCE) (R. S. Hanson and T. E. Hanson, 1996). 334 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes 8.08.4.2 Methanogen Diversity and Metabolism Whereas hydrogenotrophy is wide spread among the methanogens, acetotrophy (also Methanogens are strict anaerobes that produce known as acetate fermentation or acetoclastic CH4 as a waste product of energy metabolism. methanogenesis) is restricted to just two genera, Several characteristics set the methanogens apart the and Methanosaeta (formerly from most other microorganisms. They are the Methanothrix), which comprise —10% of metha- largest and most diverse group in the Archea, nogenic species: which is phylogenetically distinct from the other two domains of life, the Bacteria and CH COOH^C0 +CH Eukaryota. Many members of the Archea grow 3 2 4 (4) at extremes of temperature, pH, or salinity, although the distribution of methanogens is fairly The Methanosarcina use a wide variety of cosmopolitan. Methanogens have a number of substrates and have high potential growth rates, unique coenzymes (Wolfe, 1996) and differ from but their affinity for acetate is low (Jetten et al., Bacteria in the construction of their cell walls, a 1992). In contrast, the Methanosaeta specialize in characteristic that makes them insensitive to using acetate and have a high affinity for the penicillin and other antibiotics. The methanogens substrate, but their potential growth rate is low. and other Archea can be identified by an Despite its limited taxonomic distribution, aceto- trophy is the dominant methanogenic pathway in electron carrier, F42o, that autofluoresces at 420 nm under UV light (Edwards and McBride, many ecosystems. Acetotrophic methanogenesis 1975). Many methanogens lack can be considered a special case of methylotrophy and other features of electron transport chains whereby a portion of the substrate molecule is such as (Fenchel and Finlay, 1995). It oxidized to C02, while a methyl group on the same has been suggested that a -like role molecule is reduced to CH4 (Fenchel and Finlay, in electron transfer may be filled by phenazine 1995; Hornibrook et al, 2000; Pine and Barker, compounds in the Methanosarcinales (Abken 1956). In this case, the methyl group is the electron et al., 1998; Deppenmeier et ai, 1999). In a donor and the carboxyl group is the electron detailed taxonomic treatment, Boone et al. acceptor. (1993) defined 26 genera and 74 species of About 26% of methanogenic species can use methanogens. methylated substrates other than acetate, such as Despite a great deal of taxonomic diversity, the methanol, methylated amines, and methylated methanogens use a limited variety of simple sulfur compounds (Hippe et al., 1979; Kiene, energy sources compared to the other major 1991b). Because the process does not require an forms of anaerobic metabolism (Zinder, 1993). external electron acceptor, methylotrophy is a The compounds that support energy conservation type of fermentation. The contribution of these organisms to CH production in natural ecosys- for growth are H2, acetate, formate, some 4 alcohols, and methylated compounds. The most tems appears to be minor, but they may contribute substantially to the metabolism of methylated important of these are generally H2 and acetate. sulfur compounds. Scholten et al. (2003) provided About 73% of methanogenic species consume H2 (Garcia et al, 2000): thermodynamic constraints on the feasibility of such reactions. Some methanogens are dependent on a single 4H2 + C02 — CH4 + 2H20 (3) substrate, such as acetate or methanol, while others are able to grow on two or more alternative Hydrogenotrophic methanogenesis (also known as substrates. All known formate-oxidizing metha- CO2/H2 reduction or H2-dependent methanogen- nogens are also (Garcia et al., esis) is a chemoautotrophic process in which H2 is 2000), and at least one methanogen also grows by the source of both energy and electrons, and C02 fermenting pyruvate (Bock and Schonheit, 1995). is often both an electron sink and the source of Members of the Methanosaetaceae use acetate but cellular carbon. Some hydrogenotrophic methano- not H2, whereas the Methanosarcinaceae can use gens require an additional organic carbon source both substrates. Both genera fall in the order for growth (Vogels et al, 1988). C02/H2 Methanosarcinales. reduction requires a 4:1 molar ratio of H2 to Methanogens metabolize several substrates C02, yet H2 is typically at nM concentrations that do not support growth. Methanosarcina while C02 is at mM concentrations in natural barkeri was able to lower the redox potential systems. Thus, substrate limitation of hydrogeno- by dissimilatory reduction of Fe(III) until the trophic methanogenesis must always be caused by redox potential reached 50 mV, at which point a lack of the electron donor, H2. Roughly 45% of methanogenesis began (Fetzer and Conrad, all hydrogenotrophic methanogens can substitute 1993). In a survey of five methanogenic species formate for H2 in reaction (3) (Garcia et al., 2000; drawn from a wide range of phylogenetic and Thiele and Zeikus, 1988). physiological types, all the hydrogenotrophs Methane 335 could reduce Fe(III) (Bond and Lovley, 2002). are fairly tolerant of 02 (Kiener and Leisinger, The ability to grow on Fe(III) was not investi- 1983) and have adaptations such as superoxide gated. Carbon monoxide is converted to CH4, but dismutase (Kirby et al., 1981). Methane production the physiological role this substrate plays is by Methanosarcina barkeri starts when the redox uncertain (Vogels et ah, 1988). Rich and King potential drops below 50 mV (Fetzer and Conrad, (1999) measured maximum potential uptake 1993), and CH4 production has been observed in velocities of 1-2 nmol CO cm-3 sediment h_1 rice paddy soils at redox potentials >200 mV in anaerobic soils, but the response to amend- (Peters and Conrad, 1996; Yao and Conrad, 1999). ments of S04", Fe(III), and the methanogen Despite the fact that they do not form spores or inhibitor bromoethanesulfonic acid (BES), other resting stages, methanogens can survive for suggested that no more than 30% of the oxidation long periods of time in largely dry and oxic soils. activity was due to methanogens. Methanogens Mayer and Conrad (1990) observed a rapid degrade chlorinated pollutants and have been increase in CH4 production within 25 d of flooding used for (Fathepure and Boyd, an upland agricultural soil and a forest soil. Viable 1988; Mikesell and Boyd, 1990). methanogen populations survived in an oxic, air- Methanogens grow at temperatures ranging dried paddy soil for at least 2 yr (Ueki et al., 1997). from 4 °C to 100 °C, salinities from freshwater von Fischer and Hedin (2002) recently developed a to brine, and pH from 3 to 9. Most grow optimally 13C pool dilution method that gives a sensitive at temperatures > 30 °C, but thermophilic metha- estimate of gross CH4 production, and found that nogens have temperature optima near 100 °C, and CH4 is in fact produced in many dry, oxic soils -2 _1 a few isolates are adapted to frigid conditions (mean = 0.15 mg CH4-C m d ). However, (Franzmann et al., 1997). In a survey of 68 they stopped short of attributing this activity to methanogenic species, most species grew best in a methanogens because of reports that CH4 is a pH range from 6 to 8, and none could grow at pH minor metabolic product in some eubacteria such <5.6 (Garcia et al., 2000). Observations of CH4 as (Rimbaultef a/., 1988). Methanogens production in acidic environments suggests that may be able to survive in small anaerobic there are uncultured methanogens that can grow at microsites imbedded in dry soils, or they may be pH< 5.6 (Walker, 1998). protected from 02 by reactive soil . For example, the presence of FeS2 () in paddy soils was shown to enhance the survival of methanogens exposed to 0 (Fetzer et al., 1993). 8.08.4.3 Regulation of Methanogenesis 2 The lack of CH4 production in the presence of 02 Methane production is regulated mainly by in situ may be due to a combination of factors, of 02 concentration, pH, temperature, salinity, which 02 toxicity is just one. For example, meth- organic substrates, and nutrient availability. The anogens were more sensitive to desiccation than amount of CH4 emitted from wetlands is further 02 exposure in a paddy soil (Fetzer et al., 1993). influenced by a large number of factors that are not The oxidized products of denitrihcation, NO and addressed in this review, including plant physi- N20, have a toxic effect on methanogens similar ology, community composition, and hydrology. to that of 02 (Kliiber and Conrad, 1998). A nega- In addition to their direct effects on methanogen tive correlation between redox potential and CH4 physiology, these factors influence the process production is often observed in the absence of 02, indirectly by regulating the flow of methanogenic but this probably reflects competition between substrates from fermenting and syntrophic micro- methanogens and their competitors for reductants, organisms. Methanogenesis can be limited by any not a physiological requirement for a certain redox single link in the chain of reactions that begins potential. with detrital inputs. 8.08.4.3.2 Nutrients and pH 8.08.4.3.1 O2 and oxidant inhibition Laboratory incubations of wetland soils with Methanogenesis is inhibited by 02. This is nitrogen and phosphorus amendments have evident from field studies that show no overlap in shown either no effect or an inhibitory effect the depth distributions of 02 penetration in soils on methanogenesis (Bodelier et al., 2000a,b; or sediments and net CH4 production. Inhibition Bridgham and Richardson, 1992; Wang and by 02 is one reason that water table depth in Lewis, 1992). A low rate of phosphate supply to soils is often a strong predictor of CH4 emissions rice roots stimulated CH4 emission (Lu et al., (Kettunen et al, 1999; Roulet and Moore, 1995; 1999), while phosphate concentrations >20mM Sundh et al., 1995), the other reason being its specifically inhibited acetotrophic methanogenesis influence on CH4 oxidation. (Conrad et al, 2000). Methanogens are not necessarily as oxidant sen- Methanogens require a somewhat unique suite sitive as it was once believed. Some methanogens of micronutrients that include nickel, cobalt, iron, 336 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes and sodium (Jarrell and Kalmokoff, 1988). 30 Methanogenesis was stimulated by molybdenum, nickel, boron, iron, zinc, vanadium, and cobalt in a 25- rice paddy soil (Banik et ah, 1996), and by a cocktail of nickel, cobalt, and iron in Sphagnum- 20 derived peat (Basiliko and Yavitt, 2001). The 2 15 micronutrient cocktail did not stimulate C02 CD production in the same soils, suggesting that 10 methanogens, rather than fermenters, were 5 • directly limited by trace elements. Freshwater . ••••' methanogens required at least 1 mM Na+ to drive ATP formation by an Na+/K+ pump (Kaesler and 10 100 1,000 Schonheit, 1989). Trace element availability C02/CH4 could limit methanogenesis in peatlands that are isolated from inputs and sea salt Figure 5 The relationship between the temperature sensitivity (i.e., Q ) of CH production and the ratio deposition. The latter effect could be important in l0 4 C02 to CH4 produced in anaerobic incubations (van bogs in the interior of continents. Hulzen et al, 1999) (reproduced by permission of Most methanogenic communities seem to be from SoilBiol. Biochem. 1999,31,1919-1929). dominated by neutrophilic species. Some acidic peats have responded to an increase in pH with higher CH4 production (Dunfield et ah, 1993; assumed to be negligible during the winter. This is Valentine et ah, 1994), while other peats have not clearly not the case in peatlands where winter (Bridgham and Richardson, 1992). A substantial CH4 emissions contribute 2-21% of annual fluxes portion of the acetate pool may not be available to (Aim et ah, 1999; Disc, 1992; Melloh and Crill, methanogens at low pH because it is prevented 1996). Zimov et ah (1997) estimated that 75% of from dissociating (Fukuzaki et ah, 1990). CH4 emissions from Siberian lakes is emitted in the winter. 8.08.4.3.3 Temperature 8.08.4.3.4 Carbon quantity and quality Methanogenesis is often more sensitive to temperature than other biological processes, The quantity and quality (i.e., chemical com- which typically double in rate with a 10 °C increase position) of organic carbon compounds is a master in temperature (i.e., Ql0 = 2). A compilation of variable regulating methanogenesis. Organic car- temperature-response studies using wetlands soils bon fermentation is the ultimate source of both the reported an average Qw of 4.1 for CH4 production electron donors and the electron acceptors and 1.9 for aerobic CH4 oxidation (Segers, 1998). required by the two major methanogenic pathways van Hulzen et ah (1999) argued that such high Q10 (Section 8.08.4.2). There are inorganic sources of values are often caused by the changing avail- C02, but this substrate is abundant compared to ability of alternative electron acceptors or metha- H2 and does not limit hydrogenotrophic metha- nogenic substrates over the course of an nogenesis. Organic carbon also exerts a strong experiment. The time required for methanogen influence on another key regulator of methano- competitors (e.g., Fe(III) reducers) to deplete the genesis, namely the supply of inorganic com- pool of alternative electron acceptors decreases pounds that are toxic to methanogens (02, NO^) with increasing temperature. Thus, if CH4 pro- or more energetically favorable as electron duction is measured cumulatively over a fixed acceptors for competing organisms (NO^~, period of time, it can appear to increase at a Ql0 >2 Fe(III), humic acids, S04). Methanogenesis is even though the underlying processes increased at most vigorous when the consumption rate of a far lower Q10. This would explain an apparent alternative electron acceptors exceeds their supply correlation between Q10 and the ratio of C02 to rate. Because the organisms that consume these CH4 in anaerobic incubations (Figure 5). It is likely alternative substrates are heterotrophic and that temperature-sensitive steps during fermenta- usually carbon limited, such conditions are most tion or acetogenesis contribute to the temperature likely to develop in systems with a high supply of sensitivity of CH4 production. It is more difficult labile organic carbon compounds (Section 8.08.8). to explain high CH4 Ql0 values in some pure A wide variety of evidence suggests that carbon methanogen cultures (Segers, 1998). Methano- availability limits methanogenesis in situ. The fact genesis can also be insensitive to temperature that methanogenesis is often inhibited by the if fermentation is limited by other factors such presence of alternative electron acceptors such as as carbon quality (Yavitt et ah, 1988). Fe(III) and S04 is evidence of competition for Because methanogenesis is severely suppressed fermentation products and thus widespread carbon by low temperature, CH4 emissions are often limitation of the process (Section 8.08.8). Methane 337 Methanogenesis is stimulated by plant detritus photosynthates that originated in the root zone, and amendments in rice paddy soils (Dannenberg and were transported by groundwater through the soil Conrad, 1999; Inubushi et al, 1997; Kludze and profile (Chasar et al, 2000a). DeLaune, 1995; Tanji et al., 2003) and peat soils Soil organic matter quality reflects the chemical (Valentine et al, 1994), and by amendments of H2 characteristics of the dominant plant species, and it (Bridgham and Richardson, 1992; Yavitt et al., is quite poor in peatlands dominated by Sphagnum 1987;Yavitt and Lang, 1990) and dissolved moss. Sphagnum tissue has an exceedingly low organic carbon (Lu et al., 2000b). Methane nitrogen content (Aerts and de Caluwe, 1999; emissions are stimulated by organic amendments Hobble, 1996; Johnson and Damman, 1991), (Watanabe et al, 1995; Bronson et al, 1997), but which is one reason for lower CH4 emissions this effect may be due partly to lower CH4 from northern bogs than fens (Bridgham et al, oxidation caused by more rapid heterotrophic 02 1995; Dise et al, 1993; Moore and Knowles, consumption at the soil surface. 1990). Although Sphagnum is abundant in both The chemical composition of organic com- types, fens also support a variety of pounds (i.e., carbon quality) influences methano- vascular plant species with comparatively high genesis by regulating the production of tissue quality. The of Sphagnum is fermentation products (Section 8.08.3). This itself a reflection of low nutrient availability, and explains two common observations about changes is ultimately caused by the geomorphologic and in methanogenesis with increasing depth in hydrologic characteristics of the landscape. freshwater wetland soil profiles: (i) rates of Organic carbon indirectly influences methano- potential CH4 production decline with depth genesis by governing the rate that alternative below the aerobic zone (Clymo and Pearce, 1995; electron acceptors are consumed. A combination Kettunen et al., 1999; Megonigal and Schlesinger, of carbon content and alternative electron accep- 2002; Moore and Dalva, 1997; Sundh et al, 1994; tor availability explained 90% of the variation in Updergraff et al, 1995; Valentine et al, 1994; CH4 production in 10 rice paddy soils (Gaunt Yavitt et al, 1988,1990) and (ii) methanogenesis is et al, 1997). The sediments at Cape Lookout increasingly dependent on hydrogenotrophy (Hor- Bight support exceptionally high rates of CH4 nibrook et al., 2000). These depth-dependent production for a marine sediment because organic patterns reflect a combination of changes in the carbon inputs are rapid enough to deplete the pore- soil organic carbon quality, and increasing distance water S04~ pool before exhausting the labile from labile carbon substrates deposited near the organic carbon pool (Martens and Klump, 1984). surface. The mineralization of fresh organic High sedimentation rates also contribute to S04 matter is thought to favor acetotrophic methano- depletion at Cape Lookout Bight by increasing the genesis (Schoell, 1988; Sugimoto and Wada, diffusion path length, thus slowing S04 resupply 1993). In bogs, common measures of carbon from the water column. Salt marshes tend to have quality, such as lignin content, lignin: N ratio and higher CH4 emissions than marine sediments per C: N ratio, show soil organic matter becoming unit area (Bartlett et al., 1987) because they progressively older, more decomposed and recal- receive considerable organic inputs from plants, citrant with depth (Yavitt and Lang, 1990; thus relieving the substrate competition between Updergraff et al, 1995; Valentine et al, 1994). In methanogens and other microorganisms. Blair (1998) demonstrated the effectiveness of combin- a fen, 90% of the CH4 produced from soil organic matter mineralization came from particles with a ing the labile carbon flux rate and oxidant diameter >2 mm, which represents recent plant availability (e.g., 02 and S04 ) into a ratio for detritus (van den Pol-van Dasselaar and Oenema, modeling the proportion of organic carbon 1999). Because the >2 mm soil organic carbon metabolized to CPL, in marine sediments. fraction decreased rapidly with depth, 70% of the total CH production occurred in the top 5 cm of 4 8.08.4,3,5 Plants as carbon sources the soil profile. In other cases, the depth-dependent decline in potential CH4 production does not Plants and phytoplankton are the most abundant correlate to changes in the quality of the soil and labile organic carbon sources in ecosystems, organic carbon pool. For example, potential CH4 supplying photosynthates either in the form of production, but not the quality of soil organic exudates or fresh detritus. Several isotope tracer matter, decreased strongly with depth in two studies have demonstrated a tight coupling northern fens (Valentine et ah, 1994). Here, the between plant photosynthesis and methanogenesis methanogens were apparently using labile carbon in pot studies designed to represent rice paddies compounds that were produced in the root zone and (Lu et al, 2002; Dannenburg and Conrad, 1999; were less abundant with increasing distance from Minoda and Kimura, 1994; Minoda and Kimura, 14 the soil surface. This is consistent with CH4 1996), freshwater marshes (Megonigal et al., measurements that showed microbial respiration in 1999), and arctic tundra (King and Reeburgh, fens, but not in bogs, were based on recently fixed 2002; King et al, 2002). A full cycle of C02 338 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes assimilation by plants, release into soils, and emission as CH4 requires as little as 2 h, and up to 6% of the assimilated C02 is emitted as CH4. Photosynthate-derived carbon was estimated to contribute 10-100% of the CH4-C in rice paddies depending on the growth stage of the crop (Minoda and Kimura, 1994). Lu et al. (2000a,b) found that differences in the root exudation rates of three rice cultivars corresponded to variation in dissolved organic carbon in the rhizosphere and CH4 emissions. It should be noted that isotopic labeling studies alone cannot determine whether there is an energetic link between the plants and microbes (Megonigal et al., 1999). In most instances, it is difficult to apply a radiocarbon label in situ because of environmental regulations (for an exception see Wieder and Yavitt, 1994). However, a 14C label was applied worldwide during aboveground nuclear bomb testing, which peaked in 1965. Radiocarbon dates indicate that a significant amount of the r2 = 0.65 pore-water CH4 that currently exists in peatlands (Chanton et al, 1995; Charman et al, 1994; 0 5 10 15 20 25 30 35 Martens et al, 1992b), and an (Harvey (b) GPP (umol nr2 s"1) et al., 2002) was produced since 1965. In peatlands, pore-water CH4-C was significantly 1.5 younger than the bulk (Aravena et al, 1993), indicating that recently assimilated organic carbon compounds are carried downward into the 1.0 soil profile by advection. An interesting exception o 0 = 0.89 to this pattern was reported in Siberian lakes 3. where the radiocarbon age of CH4 indicated that 0.5 68-100% of the CH4-C had been assimilated • 700 ppm, wet during the Pleistocene (Zimov et al, 1997). u • 700 ppm, dry However, the contribution of this "old" carbon 0.0 fell to 23-46% in the summer with inputs of more 0.0 0.5 1.0 1.5 recent carbon sources. (c) Whole plant photosynthesis (umol s ) There are several additional lines of evidence Figure 6 The relationship between wetland CH4 for a link between methanogenesis and photosyn- emissions and various measures of primary pro- thesis. One is the observation that CH4 emissions ductivity: (a) emissions versus net ecosystem pro- are often strongly related to . duction (NEP) in North-American ecosystems ranging Such a relationship was first suggested by from the subtropics to the subarctic; (b) emissions Whiting and Chanton (1993), who reported a versus GPP in fen peatland mesocosms with high or low positive correlation between CH4 emissions and water table depths; and (c) emissions versus whole- net ecosystem exchange of C02 across North plant net photosynthesis in marsh microcosms exposed American wetlands distributed from the subarctic to elevated and ambient concentrations of atmospheric to the subtropics (Figure 6(a)). Net ecosystem COz. (after Whiting and Chanton, (1993); Updergraff et al, (2001); and Vann and Megonigal (2003), exchange (NEE) is the difference between gross respectively). primary production (GPP) and the respiration of plants (Rp) and heterotrophic organisms (Rh):

NEE = GPP -Rp-Rh (5) the relationship holds when photosynthesis is Similar relationships have been reported for many considered independently. Bridgham et al. (2001) individual peatland ecosystems (Aim et al, 1997; observed a positive correlation between GPP and Chanton et al, 1995; Whiting and Chanton, 1992; CH4 emissions in bog and fen mesocosms Whiting et al, 1991). Methanogenesis accounted (Figure 6(b)). Differences in CH4 emissions that for 4% of NEE when averaged across several of were caused by manipulating temperature and these studies (Bellisario etal, 1999). NEE is not a nitrogen availability could be explained by direct measure of photosynthetic activity because changes in photosynthesis. The effects of plant it includes Rh, but other studies have shown that community type (i.e., bog or fen) and water Methane 339 table depth needed to be accounted for separately. 8.08.4.4.1 Organic carbon availability Vann and Megonigal (2003) found that elevated The availability of labile organic carbon is atmospheric C02 stimulated CH4 flux in direct proportion to net photosynthesis in a greenhouse perhaps a more appropriate basis than salinity on study (Figure 6(c)), but again the regression which to generalize about the relative importance relationships varied by plant species and water of the two primary methanogenic pathways. table depth. Freshwater wetlands that are dominated by A final line of evidence for a direct link between acetotrophic methanogenesis also support high methanogenesis and photosynthesis is a report that rates of primary , and thus high rates of organic carbon mineralization. Labile carbon CH4 emissions from a group of peatlands were 13 pools are typically high in rice paddy soils, which correlated to S C-CH4 (Chanton et ah, 1995). The positive slope of this relationship indicated are highly productive and dominated (>50%) by that the highest fluxes occurred at sites where acetotrophic methanogenesis (Conrad, 1999). By labile carbon was relatively abundant, which comparison, Sphagnum bog peatlands have excep- favors acetotrophic methanogenesis and 13C tionally low nutrient availability, low primary enrichment (Section 8.08.4.4.1). Similar obser- productivity, and highly recalcitrant soil organic vations have been reported from other types of matter pools. Isotopic evidence and direct rate wetland and aquatic ecosystems (Chanton and measurements indicate that these systems are Martens, 1988; Martens et ah, 1986; Tyler et ah, dominated by hydrogenotrophic methanogenesis 1994). Collectively, these studies suggest that wet- (Bellisario et al, 1999; Chasar et al, 2000a; Hines land methanogens depend on labile, high-quality et ah, 2001; Lansdown et al., 1992). Despite organic carbon supplied by plants in the form of deep accumulations of nearly pure organic soils root exudates or detritus. in bogs, the size of the labile organic carbon pool is small because of poor carbon quality (Section 8.08.4.3.4). Whereas bogs are isolated from groundwater, fen peatlands receive substan- tial groundwater inputs and are characterized 8.08.4.4 Contributions of Acetotrophy versus by comparatively high nutrient availability, her- Hydrogenotrophy baceous plant biomass, labile carbon availability, Acetotrophy and hydrogenotrophy are the CH4 emissions, and acetotrophic methanogenesis. dominant methanogenic pathways in situ, This is particularly evident at the soil surface. although trimethylamines and methylated sulfur The size of the labile carbon pool is an species may be important in some marine important predictor of methanogenic pathways in sediments (Oremland et ah, 1982). Because marine sediments (Blair, 1998). Sediments with acetotrophic methanogenesis is associated with low carbon availability are dominated by hydro- high rates of total CH4 production, it is useful to genotrophic methanogenesis, while carbon-rich understand the factors that govern the relative systems such as Cape Lookout Bight support dominance of these two pathways. The theoretical relatively more acetotrophic methanogenesis contribution of hydrogenotrophic methanogenesis (Martens et ah, 1986). The labile portion of the to CH4 production during anaerobic degradation organic carbon pool in marine sediments is of carbohydrates is 33%, and this value is often determined by several factors such as chemical observed in freshwater ecosystems (see Conrad composition, adsorption by mineral surfaces, and (1999) for a partial compilation). However, the burial efficiency (Hedges and Keil, 1995). Poor relative contributions of the two pathways can organic carbon quality is expected in coastal vary considerably across ecosystems, seasons, and sediments that receive large inputs of terrestrially depths. An early generalization was that aceto- derived, highly degraded particulate organic trophic methanogenesis is dominant in freshwater matter from rivers (Hedges et ah, 1994; Martens ecosystems, while hydrogenotrophic methanogen- ef a/., 1992a). esis is dominant in marine systems (Whiticar, Stable isotopes provide a nondestructive 1999). This is a reasonable first approximation alternative to radiolabeled substrates for inferring because S04" reducers readily outcompete the relative importance of methanogenic path- methanogens for the limited acetate supply in ways. Whiticar et al. (1986) recognized that marine systems. However, there are examples of acetotrophic methanogenesis yields CH4 that is freshwater systems dominated by hydrogeno- 13C-enriched (S13C = -65%o to -50%o) com- trophic methanogenesis (Lansdown et al., 1992), pared to hydrogenotrophy (813C = — 110%c to marine systems with a relatively large contri- —60%c). The reason for this difference is a 12 13 bution from acetotrophic methanogenesis (Martens larger C/ C fractionation during CH4 pro- et ah, 1986), and wide variation in time and space duction by hydrogenotrophy (ac = 1.055-1.090) within a given site (Martens et al., 1986; Sugimoto than acetotrophy (ac = 1.04-1.055). Similarly, and Wada, 1993). CH4 from acetotrophy is deuterium-depleted 340 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

(SB = -400%c to -250%c) compared to hydro- 50 genotrophy (SD =-250%o to -170%o). Sub- sequent studies have suggested somewhat 40 broader ranges for these fractionation factors (Tyler, 1991). Plots of SD versus S13C can be used to infer the extent of aerobic or anaerobic 30 oxidation of CH4 (Alperin et ai, 1988; Coleman et ah, 1981). If oxidation has a substantial .3 20 influence on the CH4 pool, the relationship between <5D and S13C is positive. In the absence 10 of oxidation, and provided that the relative contributions of the methanogenic pathways varies in time or space, the relationship between 3D and 513C is negative. In such cases, the 10 20 30 40 50 13 (a) Lignin content (w/w%) variability in <5 C-CH4 can often be interpreted as a shift in the contribution of acetotrophy and -15 hydrogenotrophy to total methanogenesis (Burke et al, 1988, 1992). There are several factors other -20 than the methanogenic pathways and oxidation that can affect isotope fractionation factors (Happell et al, 1993; Avery and Martens, 1999; 59 -25 Bergamaschi, 1997; Tyler, 1992), and these make U it difficult to interpret stable isotope data in terms -30 of absolute rates. Additional fractionations occur when CH4 passes through plants, which must be < -35 accounted for when interpreting the stable isotope ratios of CH4 emitted from wetlands (Chanton et al, 1999a,b; 2002; Chanton and Dacey, 1991; -40 Harden and Chanton, 1994). 5 10 15 20 25 30 35 13 13 Fraction mineralized (C/C%) Plots of S C-CH4 versus S C-XC02 in peat- (b) land soils typically indicate a progressive tran- Figure 7 Influence of (a) lignin content on leaf litter sition from acetotrophic methanogenesis at the decomposition rates, and (b) leaf litter decomposition surface to hydrogenotrophic methanogenesis at rates on methanogenic pathway as reflected in the S l3C l3 depth (Chasar et al, 2000b; Hornibrook et al, of CH4. In (b), the y-axis is the difference in S C-CH4 l3 1997; 2000), that coincides with increasingly and 8 C of total mineralized carbon (CH4 + C02) recalcitrant soil organic matter pools and distance (Miyajima et al., 1997) (reproduced by permission of from plant-derived labile carbon at the soil surface Elsevier from Geochim. Cosmochim. Ada, 1997, 61, (Section 8.08.4.3.4). In a survey of the SD and 3739-3751). 13 8 C content of CH4 in gas bubbles collected from sediments in western Alaska, there was a marine sediments when the labile carbon flux is larger contribution from acetotrophic methano- high enough to deplete pools of 02, Mn(IV), genesis at vegetated sites than those that were Fe(III) and S04 . However, the relationship did unvegetated (Martens et al, 1992b). Miyajima not hold for deep-sea sediments (Blair, 1998). 13 et al. (1997) reported that C-CH4 enrichment The contribution of acetotrophy to methano- increased in proportion to increasing leaf litter genesis can be quite stable in time (Avery and decomposition rate; decomposition rate was in Martens, 1999; Burke et al, 1992; Hines et al, turn negatively related to lignin content (Figure 7). 2001), or it can vary strongly with the seasons. A Thus, this study made a direct link between carbon mid-latitude bog was dominated by hydrogeno- quality and the relative contributions of hydro- trophic methanogenesis for 10 months of the year, genotrophic and acetotrophic methanogenesis to then switched to acetotrophic methanogenesis overall CH4 production. This study was also (Avery et al, 2002). Because the switch was notable because it was a rare investigation of accompanied by a large increase in CH4 emis- anaerobic metabolism in a tropical peatland. In a sions, acetotrophy contributed —50% of total variety of shelf and slope marine sediments, 813C- annual emissions. Seasonal shifts in methanogenic CH4 is positively related to the flux of labile pathways were observed in several peatlands organic carbon to the sediment surface (Boehme located at mid-latitudes (Kelley et al, 1992; et al, 1996) and to the rate of pore-water S04" Lansdown et al, 1992; Shannon and White, depletion (Figure 8). This relationship indicates 1996). By comparison, no such switch occurred that acetotrophic methanogenesis makes a rela- in a group of bogs and fens located at high latitudes tively large contribution to CH4 production in (Duddleston et al, 2002; Hines et al, 2001). Methane 341

-50 Chin et al, 1999a; Conrad et al, 1987; Fey and Conrad, 2000; Schultz and Conrad, 1996; Schultz -55- et al, 1997; Yao and Conrad, 1999). Although changes in the relative contributions of the two -60 pathways coincide with shifts in the structure of methanogenic communities (Chin et al., 1999b; ,-65- Fey and Conrad, 2000), they are probably not a U direct response of methanogens to temperature. Rather, the methanogens appear to be responding to changes in substrate availability (Figure 9(b)), -75- suggesting that the actual source of the observed temperature limitation was organisms that produce or consume H2 and acetate. Indeed, H2 amend- ments stimulated hydrogenotrophic methano- -85- genesis at low (15 °C) temperature in a paddy soil, but not cellulose amendments, which would -90" 0 12 3 4 have required fermentation to methanogenic sub- Sulfate gradient (mM cm-1) strates (Schultz et al., 1997). This result is consistent with temperature limitation of syn- Figure 8 The relationship between that rate at which trophic bacteria, which produce H2 by fermenta- SO|~ is consumed down-core in shallow marine l3 2 tion (Section 8.08.3.2.2). It is more difficult to sediments and the 8 C-CH4. The r fit of the regression explain an increase in acetate concentrations at low line is 0.98 (Blair, 1998) (reproduced by permission of Elsevier from Chem. GeoL, 1998, 152, 139-150). temperatures. One proposal is that low temperature favors the fermentation of multicarbon substrates directly to acetate (Conrad, 1996; Fey and Conrad, These sites were completely dominated by hydro- 2000), but this has not been demonstrated. Aceto- genotrophic methanogenesis despite an accumu- genesis was shown to be <10% of the acetate lation of acetate in the pore water. Incubating sources at low temperature (Rothfuss and Conrad, these soils at 24 °C for 5 months did not trigger 1993; Thebrath et al, 1992). This body of work acetotrophic CH4 production, even though acetate demonstrates that building a mechanistic under- continued to accumulate during the period. standing of the temperature responses of methano- Apparently, the methanogens in these systems gens will require comprehensive studies that are limited by factors other than temperature, such include fermenting, syntrophic, and homoaceto- as microbial community composition, pH, or genic bacteria. perhaps trace nutrient availability (Section Physiological differences in the two methano- 8.08.4.3.2). genic genera that can use acetate may influence Chapelle et al. (2002) recently described a temperature-driven changes in methanogenic subsurface microbial community in which hydro- pathways (reviewed by Liesack et al, 2000). The genotrophic methanogens were >90% of the 16S members of the Methanosaetaceae that have been ribosomal DNA sequences and geothermal H2 was isolated to date use acetate exclusively and have a the primary energy source. Since geothermal H2 is relatively low threshold for the substrate (typically likely to be an abundant energy source for <100 |xM), while members of the Methanosarci- microbial metabolism on other planets, hydroge- naceae use both acetate and H2/C02, and have a notrophy may be an important pathway of relatively high threshold for acetate (typically anaerobic metabolism elsewhere in the universe. between 200 and 1200 pM) (Jetten et al, 1992; GroGkopf et al, 1998). The relative abundance of the two genera in paddy soils appears to be a function of both temperature and acetate concen- 8.08.4.4.2 Temperature tration (reviewed in Liesack et al, 2000). When Temperature is one of several factors that acetate concentrations were above the threshold of influence the contributions of acetotrophic and both genera, the Methanosaeta were dominant, hydrogenotrophic methanogenesis to overall CH4 presumably because they were more tolerant of production. The most comprehensive work on this suboptimal (15 °C) temperature than the Methano- topic has been done by Conrad and colleagues in sarcina (Chin et al, 1999a,b). The Methanosar- rice paddy soils and lake sediments. They have cina became dominant at high (30 °C) temperature, repeatedly observed that the contribution from which is consistent with the fact that they can also hydrogenotrophic methanogenesis declines at low use H2. The reverse of this pattern was observed temperatures, while the contribution of aceto- when acetate levels were maintained below the trophic methanogenesis shows the opposite threshold concentration for both genera. In this pattern (Figure 9(a); Chin and Conrad, 1995; case, the Methanosaeta were dominant at the high 342 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

40 35 & 30 o 25

X 90 oH 15 uX 10 experiment 1 experiment 2

10 20 25 30 35 40 (a) Temperature (°C) 3.5 70

3.0 60

2.5 50 ' /^" 2.0 40

—o— hydrogen 1.5 "\ cr°*^ 30 3

1.0 20 0.5 /^^V\ / v.-*-*---. 10 0.0 10 15 20 25 30 35 40 (b) Temperature (°C) Figure 9 The influence of temperature on: (a) methanogenic pathways and (b) methanogenic substrates (after Fey and Conrad, 2000). temperature because acetate concentrations were Martens and Berner, 1977). The process was below the threshold concentration of the Metha- inferred from pore-water S04 and CH4 profiles nosarcina (Fey and Conrad, 2000). because it imparted the CH4 profile with a strong The extensive work on methanogenic pathways "concave up" shape (Martens and Berner, 1977). in profundal lake sediments and rice paddy soils Further geochemical evidence for this process 14 14 does not appear to apply to a broader sample of included conversion of CH4 to C02 under anaerobic ecosystems. The proportions of CH4 anoxic conditions, and a zone of relatively 13 produced from hydrogenotrophic and acetotrophic C-enriched CH4 that coincided with a zone of 13 methanogenesis did not vary seasonally in a C-depleted C02 (citations in Valentine, 2002). temperate tidal freshwater estuarine sediment Although anaerobic CH4 oxidation was shown to (Avery and Martens, 1999; Avery et al., 1999) be microbially mediated long ago (Reeburgh, or a northern peatland soil (Hines et al., 2001) 1982), recent advances in molecular biology and despite changes of 25 °C or more. Temperature stable isotope techniques have provided compel- alone could not explain transient changes in the ling new biological insights on the process. methanogenic pathways in a mid-latitude (45° N) Nevertheless, there are no pure cultures or co- peatland (Avery et al, 1999, 2002). Clearly, cultures of organisms that oxidize CH4 anaerobi- detailed mechanistic studies on temperature cally, and several competing hypotheses about the regulation of fermentation and methanogenesis metabolic pathways involved are still considered need to be done in a wider variety of anaerobic viable explanations of the existing data (reviewed ecosystems (e.g., Kotsyurbenko et al, 1996). by Hinrichs and Boetius, 2002; Valentine, 2002). Three primary mechanisms have been offered to explain anaerobic CH oxidation in marine sedi- 8.08.4.5 Anaerobic Methane Oxidation 4 ments. The process may be carried out by a

Anaerobic CH4 oxidation in marine sediments single organism that couples methanotrophy to was first observed around the mid-1970s S04~ reduction. Anaerobic methane oxidation (Barnes and Goldberg, 1976; Reeburgh, 1976; coupled to sulfate reduction is thermodynamically Methane 343 favorable under conditions found in marine in a molar ratio of 1:1 in marine sediments sediments (Martens and Berner, 1977): (Nauhaus et al., 2002). 13 The fact that CH4 is highly depleted in C has CH4 + SO4" — HS - +HC03~ + H20 (6) proven useful for constraining the nature of the microorganisms that oxidize methane anaerobi- A second possibility yields the same net reaction 13 but is driven by a syntrophic relationship between cally. Because the S C of CH4 ranges from about an organism that oxidizes CH to C0 /H and an -50%c to -110%o PDB (Whiticar et al., 1986), the 4 2 2 fate of CfLrderived carbon can be traced through SO4 -reducing bacterium that consumes H2, thereby maintaining a thermodynamically favor- food webs by means of its distinctive isotopic able environment for the reaction (Hoehler et al., ratio. Such work has tended to support some form of reverse methanogenesis. Most isotope studies 1994). In this case, anaerobic CH4 oxidation is analogous to hydrogenotrophic methanogenesis in have been done in sediments overlaying CH4 reverse: hydrate deposits or seeps where the high pore- water CH4 concentration favors rapid anaerobic CH4 + 2H20 — CO, + 4H2 (7) CH4 oxidation. At such sites, Archea-specific lipids are abundant and highly depleted in 13C S0 " + 4H + H+ — HS Hco; 4H 0 (often <—100%c), indicating they could only be 4 2 2 derived from CH (Bian et al., 2001; Elvert et al., (8) 4 1999, 2000; Hinrichs et al, 1999, 2000; Orphan The ability to run metabolic pathways in forward or et al., 2001a; Peckmannera/., 1999; Pancostef al., reverse, thereby swapping end products for sub- 2000, 2001; Thiel et al, 1999, 2001). Interest- strates, has been documented previously in anaero- ingly, fatty acids that are characteristic of S04~- bic organisms. For example, the homoacetogen reducing bacteria are also highly 13C-depleted (up nicknamed Reversibacter oxidizes acetate to C02/ to —75%c) suggesting that S04~ reducers are H2 when the H2 concentration is low, and vice versa consuming an organic intermediate produced by (Lee and Zinder, 1988; Zinder and Koch, 1984). methane oxidizers (Boetius et al., 2000b; Hinrichs However, attempts to reverse methanogenesis in et al, 1999; Orphan et al, 2001a,b; Thiel et al, pure cultures by varying the H2 concentration have 2001). If so, this is evidence that anaerobic CH4 not been successful (Valentine et al, 2000). oxidation is performed by an Archea-Bacteria Thermodynamic and kinetic considerations consortium. suggest that acetogenic methanogenesis running Perhaps the most striking evidence of a in reverse is feasible only in high-CH4 environ- consortium between Archea and Bacteria is their ments such as CH4 seeps, as are pathways that spatial arrangement in hydrate seep sediments. require interspecies transfer of acetate (Valentine, Using fluorescent phylogenetic stains, Boetius 2002). Likewise, pathways the require interspecies et al. (2000b) observed clusters of —100 Archeal transfer of H2, formate or methanol are not cells surrounded by a 1-2 cell-thick shell of favorable for anaerobic CH4 oxidation under sulfate-reducing bacteria (Figure 10). Orphan et al. high- or low-CH4 conditions (S0rensen et al., (2001b) further showed that highly depleted S13C 2001). Nauhaus et al. (2002) were unable to stimulate S04~ reduction in sediments collected from a CH4 seep by adding H2, formate, acetate, or methanol. A modification of the reverse methanogenesis hypothesis was offered by Valentine and Reeburg (2000). This reaction would yield twice the amount of energy normally associated with anaerobic CH4 oxidation (reaction (6)), making it more thermodynamically favorable:

2CH4 + 2H20 — CH3COOH + 4H2 (9) In this case, the methanogens operating in reverse produce acetate and H2, which are then consumed by S04"-reducing bacteria:

z SOi • 4H2 — S 4H20 (10) Figure 10 In situ identification of microbial aggre- SO4 CH3COOH -^ H2S + 2HCO3 (11) gates consisting of Archea (red) in the center and S04~-reducing bacteria (green) on the periphery. Micro- All of these mechanisms conform to the obser- organisms were labeled with rRNA-targeted oligio- 2_ vation that CH4 is consumed and S is produced nucleotide probes, (source Boetius et al., 2000b). 344 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes ratios coincided with the Archeal core of these Anaerobic CH4 oxidation occurs in freshwater aggregates. Collectively, the evidence supports a environments where S04 concentrations are low syntrophic model of anaerobic CH4 oxidation in (Smith etal, 1991;Miuraef al, 1992; Murase and which methanogen-like Archea oxidize CH4, and Kimura, 1994a,b; Smemo and Yavitt, 2000; S04~-reducing bacteria consume end products, Grossman et al, 2002), suggesting that electron thereby making the reaction thermodynamically acceptors other than SO4~may play a role in the favorable. It remains to be determined whether the process. Thermodynamic considerations suggest Archeal member is a methanogen capable of both that anaerobic CH4 oxidation pathways may vary CH4 production and oxidation, or an obligate between sites as a function of CH4 partial pressure methanotroph. Highly 13C-depleted Archeal lipids (Valentine, 2002). Further progress will undoubt- have also been observed in solitary organisms, edly require isolating or enriching the responsible suggesting a is not necess- organisms. arily a requirement for anaerobic CH4 oxidation Very little information exists about the ecologi- (Orphan et al., 2002). cal conditions that govern anaerobic CH4 oxi- Little is known about the organisms that oxidize dation. Manipulation of a CH4 seep sediment methane anaerobically except for the broad out- demonstrated a broad temperature optimum of lines of their phylogeny. Sequences of 16S rRNA 4-16 °C, and rates that varied predictably with cloned from several CH4 seep sediments are temperature (Nauhaus et al, 2002). The process is dominated by two groups of previously unknown widespread in marine sediments and water Archeal genes, ANME-1 and ANME-2 (Boetius columns provided there is a sufficient supply of et al, 2000b; Hinrichs et al, 1999; Orphan et al, labile organic carbon to permit substantial metha- 2001a). The ANME-1 group typically occurs as nogenesis (Section 8.0.8.4.3.4), or some other single filaments or monospecific mats, and does CH4 source exists. Anaerobic CH4 oxidation is the not appear require a bacterial partner (Orphan primary sink for CH4 in the world oceans et al, 2002). They have been recovered from (D'Hondt et al, 2002), and it accounts for a hydrothermal vents (Takai and Horikoshi, 1999; large fraction of the S04 reduced in some marine Teske et al, 2002), methane hydrates (Lanoil sediments (Table 6), particularly in the narrow zone et al, 2001), and shallow marine sediments where CH4 and S04~ profiles intersect. Sulfate -2 : (Thomsen et al, 2001). Massive ANME-1 reduction rates as high as 140 mmol m d~ have "reefs" were recently discovered in the Black been reported above decomposing CH4 hydrate Sea (Michaelis et al, 2002). The cores of deposits (Boetius et al, 2000b), suggesting that anaerobic methane-oxidizing aggregates have so anaerobic oxidation significantly reduces CH4 flux far been composed of ANME-2 type Archea. This to the water column from these vast carbon group is closely related to known methanogens in reservoirs (Gornitz and Fung, 1994). the Methanosarcinales, an order with members that ferment acetate to CH4, but may also oxidize 8.08.4.6 Aerobic Methane Oxidation small amounts of acetate to CH4 (Zehnder and Brock, 1979). The proposition that some members Methane is subject to aerobic oxidation by of the anaerobic consortium are capable of methanotrophic bacteria when it diffuses across an oxidizing CH4 to acetate and H2 (reaction (9)) is anoxic-oxic interface before escaping to the attractive because it provides a mechanism to atmosphere (King, 1992): explain the presence of highly 13C-depleted lipids CH + 20 -+ C0 + 2H 0 (12) in S04-reducing bacteria (Valentine and Reeburg, 4 2 2 2 2000). It is known that acetate is consumed by members of the same S0 -reducing group found Methanotrophs occur at the oxic-anoxic interface 4 of methanogenic habitats, in symbiotic association in CH4 seep sediments (i.e., Desulfosarcinales). Judging by the microbial diversity of methane with animals (Kochevar et al, 1992), and inside wetland plants (Bosse and Frenzel, 1997). seeps, anaerobic CH4 oxidation may not be fully explained by any single mechanism proposed so Although methanotrophs dominate aerobic CH4 far. Many of the SRB associated with anaerobic oxidation, NH4-oxidizing bacteria may account for a small amount of the CH4 oxidation activity in CH4 oxidation environments are unknown and unique (Thomsen et al, 2001), and biomarker data soils and sediments (Bodelier and Frenzel, 1999). suggest a great deal of taxonomic diversity among both Archea and Bacteria in sediments that support 8.08.4.6.1 Methanotroph diversity anaerobic CH4 oxidation (Orphan et al, 2001a; Thiel et al, 2001). Specific inhibitors intended for Methanotrophs are obligate aerobes that use methanogens, S04~ reducers, and acetate oxidi- CH4 as a sole carbon and energy source (reviewed zers have sometimes influenced rates of anaerobic by R. S. Hanson and T. E. Hanson, 1996). They CH4 oxidation (Hoehler et al, 1994) and had no are a subset of the methylotrophic bacteria, all of effect at other times (Alperin and Reeburgh, 1985). which oxidize compounds lacking C-C bonds Methane 345 Table 6 Selected estimates of the proportion of SO; reduction in marine sediments that is mediated by anaerobic CH4 oxidation. Site Peak contribution Depth-integrated Citation to S04~ reduction contribution to SO4 w reduction r%) Aarhus Bay, Denmark 47-52 9 Thomsen et al. (2001) Kattegat, Denmark 61 10 Iverson and J0rgensen (1985) Skagerrak 89 10 Iverson and J0rgensen (1985) Zone, Namibia 100 Niewohner et al. (1998) Amazon Fan Sediment 50-85 Burns (1998) Norsminde Fjorde 10-30 Hansen et al. (1998) Big , Nevada 2 Iversenef al. (1987) Kysing Fjord <0.1 Iversen and Blackburn (1981) Hydrate Ridge, Oregon^ 100 100 Boetius et al. (2000b)

' Inferred by comparing SRR above decomposing CH4 hydrates to nearby nonhydrate sites.

(i.e., Ci compounds) (Murrell and Kelley, 1993). fixation. The extent to which activity The ability to metabolize Ci compounds is a is expressed by type I methanotrophs in situ feature that methylotrophs have in common with remains to be determined. methanogens. Indeed, the two groups share many All known methanotrophs have a low affinity homologous genes for Ci metabolism, despite for CH4 (Km > 1 |xM) and none can maintain the large evolutionary distance between the growth on CH4 at atmospheric concentration (for Archea and the (Chistoserdova a compilation of kinetic data see Conrad, 1996). ef a/., 1998). Although there are no pure cultures of "high Methanotrophic species are separated into two affinity" CH4 oxidizers, these organisms are types that differ with respect to phylogeny, known to exist in upland soils where they account ultrastructure, lipid composition, biochemistry, for —10% of the annual global CH4 sink and physiology. The type I methanotrophs belong (Table 5). High-affinity CH4 oxidizers evidently to the family Methylococcaceae in the gamma occur in wetland soils as well because there have proteobacteria; the type II methanotrophs are in been reports of net CH4 consumption from the the family in the alpha proteo- atmosphere in response to falling water table bacteria. There is growing evidence that the two depth (Happell and Chanton, 1993; Harriss et al., groups also differ ecologically. Amaral and 1982; Pulliam, 1993; Roulet et al, 1993; Knowles (1995) used opposing gradients of Shannon and White, 1994), low temperature CH4 and 02 to determine that a type I (Megonigal and Schlesinger, 2002), and low methanotroph preferred a somewhat lower CH4 soil organic matter (Giani et al., 1996). The concentration than a type II methanotroph. An capacity for both types of oxidation kinetics in often-cited ecological distinction between the two saturated soils and sediments has been assumed groups is that only the type II methanotrophs are to be due to the presence of mixed methano- N2 fixers (R .S. Hanson and T. E. Hanson, 1996). trophic populations (Bender and Conrad, 1992). This is consistent with a report that a type II This was confirmed in a recent study that methanotroph out-competed a type I methano- demonstrated the power of combining lipid troph in a nitrogen-limited chemostat culture biomarker analysis with isotope pulse-labeling (Graham et al., 1993), and in a field study where (Boschker et al, 1998). Bull et al (2000) pulse- 13 NHj fertilization increased the proportion of type labeled a forest soil with CH4, and followed its I methanotrophs in planted rice paddy soils incorporation into phospholipid fatty acids (Bodelier et al, 2000b). In the latter study, the (PLFAs). At the soil surface, where ambient abundance of type II methanotrophs did not CH4 levels prevail, the label was incorporated change in response to nitrogen fertilization, into novel organisms with PLFAs similar to type suggesting that only the type I methanotrophs II methanotrophs. A buried soil horizon with _1 were nitrogen limited. However, a survey of higher CH4 levels (100|JLLL ) indicated the several type I and type II methanotrophic strains presence of both type I and type II methano- for N2-fixation genes (m/H) and nitrogenase trophs. Another promising technique is "stable- activity demonstrated that the capacity for N2 isotope probing" in which a microbial community fixation occurs in both groups (Auman et al., is exposed to an isotopically labeled substrate 13 2001). Thus, it is no longer valid to assume that (e.g., CH4), thereby changing the mass of type I methanotrophs lack the capacity for N2 organisms that assimilate the substrate. The 346 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes labeled and unlabeled microorganisms are separ- fraction of the 02 demand in some systems; it ated by density-gradient centrifugation and DNA accounted for 30% of 02 consumption in Lake Erie sequenced (Radajewski et ah, 2000). This (Adams et al, 1982) and >9% of 02 consumption technique demonstrated that novel proteobacteria in Lake Constance (Frenzel et al, 1990). contributed to high-affinity CH4 oxidation in a Root-associated methanotrophy has been peat soil (Morris et ai, 2002). demonstrated to be 02-limited in the rhizosphere Until recently, all cultured methanotrophs of emergent aquatic macrophytes such as Phrag- required a pH >5 and most were neutrophilic mites australis (van der Nat and Middelburg, (R. S. Hanson and T. E. Hanson, 1996). Yet, CH4 1998), Sparganium eurycarpum (King, 1996a), oxidation at pH <5 has been demonstrated and Typha latifolia (Lombard! et al, 1997). Plant in acidic Sphagnum-bogs (Born et al, 1990; species with high rates of root 02 loss also support Dunfield et al., 1993; Fechner and Hemond, high rates of rhizosphere CH4 oxidation (Calhoun 1992; Yavitt et al., 1990). Relatively recently, and King, 1997), further suggesting the process is acidophilic methanotrophic bacteria have been generally 02 limited. Differences in root 02 loss isolated and two species have been described with rates among plant species should cause species- an optimum activity between pH 4.5 and 5.5 specific differences in rhizosphere CH4 oxidation (Dedysh etai, 1998a,b, 2000, 2002). This demon- rates if the process is 02 limited. Indeed, soil-free strates that some methanotrophs are adapted to root CH4 oxidation potentials vary among plant the ambient pH conditions of their environment. species (Sorrell et al, 2002). Because >75% of The relative abundance of various methano- the CH4 efflux from wetlands may pass through trophic species or types has been investigated in plants (Chanton and Dacey, 1991; King, 1996a; peatlands, rice paddies, lakes and elsewhere with Shannon and White, 1994), 02 limitation in the PLFA analyses, community DNA analyses rhizosphere should translate into 02 limitation on (McDonald and Murrell, 1997), and fluorescent an ecosystem basis. in situ hybridization (). A suite of FISH By comparison with uncultivated plants, the probes has been developed for methanotrophs case for 02 limitation in the rice rhizosphere is not with a range of specificity for types, genera and so clear. Linear correlations between rhizosphere species (Amann et al., 1990b; Bourne et al., 2000; CH4 oxidation and CH4 concentration suggested Dedysh et al, 2001; Filer et al, 2001; Gulledge that oxidation is CH4-limited (Bosse and Frenzel, et al, 2001; Rossello-Mora et al, 1995). The use 1998; Gilbert and Frenzel, 1995; Gilbert and of such probes in an acidic northern peatland Frenzel, 1998). This result is consistent with a showed that <1% of the methanotrophic commu- study by Denier van der Gon and Neue (1996) nity was type I. Although the remainder could in which the atmospheric 02 concentration was tolerate low pH, <5% were known acidophiles doubled to 40% around rice shoots, but CH4 oxida- (Dedysh et al, 2003). A single , Methylo- tion increased by only 20%. However, kinetic and cystis, contributed 60-95% of all detectable theoretical considerations have also suggested methanotrophic cells. that 02 limits rhizosphere CH4 oxidation in rice (Bosse and Frenzel, 1997; van Bodegom et al, 2001). 8.08.4.6.2 Regulation of methanotrophy To the extent that 02 limitation exists, 02 Methane oxidation rates can be limited by 02, competition is an important factor regulating CH4 CH4, or nitrogen. Limitation by 02 is most likely to oxidation efficiency. Based on kinetic modeling, occur in continuously flooded or submerged van Bodegom et al. (2001) concluded that systems that are strongly methanogenic. Methane methanotrophs were only able to out compete oxidation is limited by 02 flux to the sediment heterotrophic bacteria for 02 when 02 concen- surface in lakes (Frenzel et al, 1990) and flooded trations were <10|JLM. The competitive advan- wetlands (King, 1990a,b; King et al, 1990). For tage of the methanotrophs at low 02 this reason, the presence of algal mats or sub- concentrations was due to their high affinity for merged aquatic vegetation can cause diurnal 02. There are other microaerophilic bacteria with variations in CH4 oxidation, with the highest a high affinity for 02 (e.g., Fe(II)-oxidizing rates during daylight due to photosynthetic 02 bacteria, Section 8.08.6.5.2) that compete with production (King, 1990b). Heilman and Carlton methanotrophs for 02, as well as abiotic oxidation (2001) reported a particularly dramatic example of reactions involving Fe(II) or H2S. The continued diel variation in CH4 emissions driven by 02 development of simulation models, such as the release from submerged aquatic macrophytes. one offered by van Bodegom et al. (2001), will In this case, the sediment was a net CH4 sink help integrate the effects of the multiple factors in daylight and a CH4 source at night. However, affecting rhizosphere CH4 oxidation. they concluded that 02 was inhibiting methano- Methane availability is likely to limit oxidation genesis rather than stimulating methanotrophy. in systems that are weakly methanogenic, such as Methanotrophy itself can account for a substantial marine sediments or dry-end wetlands that have Methane 347 exposed soils and subsurface water tables. substantially to NH^ oxidation. Competition Methane-limited methanotrophy was recently between NH^ and CH4 for the active site on demonstrated in a pair of tidal forested wetlands MMO is one explanation for the common that were aerobic in the top 5-10 cm (Megonigal observation that nitrogenous fertilizers inhibit and Schlesinger, 2002; Figure 11). Indirect CH4 oxidation (Steudler et ah, 1989; Conrad and evidence for CH4-limited methanotrophy in peat- Rothfuss, 1991; Crill et ah, 1994; King, 1990a), land ecosystems is the observation that potential and it is consistent with fact that inhibition often

CH4 oxidation tends to peak near the water table varies with the relative concentrations of CH4 and boundary, where CH4 concentrations are high and NHj (Bosse et ah, 1993; Boeckx and Van Cleemput, 1996; Dunfield and Knowles, 1995; 02 concentrations are low (Kettunen et ah, 1999; Sundh et ah, 1995; Sundh et ah, 1994). Such a King and Schnell, 1994a,b; Schnell and King, relationship should be common in bog-type peat- 1994; van der Nat et ah, 1997). A second lands that are typically dry at the soil surface, have explanation for depressed CH4 oxidation is NO2 toxicity (King and Schnell, 1994b; Kravchenko, relatively low potential CH4 production rates, and are dominated by nonvascular plants. Methano- 1999). Both of these factors should be relatively trophy in bogs can also be limited by high rates of unimportant in anaerobic environments because diffusivity through the soil profile due to low bulk CH4 concentrations are high and nitrification rates density (Freeman et ah, 2002). are low due to limited 02 availability. Indeed, nitrogen fertilization unexpectedly stimulated Methane oxidation can be limited by nitrogen CH oxidation in rice paddy soils (Bodelier et ah, due to enzyme-level inhibition or unmet metha- 4 2000b), where both CH concentrations and plant notroph nitrogen demand. The oxidation of CH 4 4 nitrogen demand are typically high. These authors by methanotrophic bacteria and NH^ by used a combination of radioisotope labeling and ammonium-oxidizing bacteria is initiated by PFLA analysis to determine that type I methano- similar enzymes— trophs were nitrogen-limited in the absence of (MMO) and ammonium monooxygenase (AMO). fertilizer application. Using the same technique, The membrane form of MMO is evolutionary Nold et ah (1999) observed that NH| additions related to AMO (Holmes et ah, 1995); there is also decreased CH4 incorporation into methanotroph a soluble form of MMO. MMO and AMO fortui- lipids in a lake sediment. Although CH4 oxidation tously oxidize a variety of additional compounds is clearly influenced by fertilization, the variety of (Hanson and Hanson, 1996), including NH^ different responses that have been observed and (MMO) and CH4 (AMO). Bodelier and Frenzel the number of mechanisms that have been (1999) determined that the contribution of nitri- proposed for these effects makes it difficult to fiers to CH4 oxidation was negligible in rice paddy generalize about the process (Gulledge et ah, microcosms; however, methanotrophs contributed 1997). In a recent compilation of the wetland CH4 oxidation literature, Segers (1998) concluded that pH is not an important factor governing aerobic 40 CH4 oxidation, salt concentrations of —40 mM inhibit the process, and aerobic CH4 oxidation -0.623 +0.73 b:, ^ = 0.962 responds to temperature with Q10 = ~2. The Ql0 for CH4 oxidation can be <2 when there is a 30 phase-transfer limitation on CH4 diffusion (e.g., King and Adamsen, 1992).

uX M 20 0 8.08.4.6.3 Methane oxidation efficiency

The amount of CH4 emitted to the atmosphere 10 is just a fraction of that produced because of X u methanotrophy. Globally, it accounts for —80% of gross CH4 production (Reeburgh, 1996; Reeburgh et ah, 1993). The wetland CH4 ~l— —I— —I— 10 20 30 40 oxidation sink has been estimated at 40% 2 Gross CH4 emissions (mg CH4 m d ') (King, 1996b) to 70% (Reeburgh et ah, 1993) of gross CH4 production, or roughly 100- _1 Figure 11 Relationship between rates of gross CH4 400Tgyr . Methane oxidation efficiency is emission and CH4 oxidation measured over a 13-month generally higher in wetland forests than marshes, period in two tidal freshwater wetlands (source perhaps because forests occupy drier positions on Megonigal and Schlesinger, 2002). the landscape and develop a relatively deep oxic 348 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes zone at the soil surface (Megonigal and Schle- estimate methanotrophy in the rice rhizosphere. singer, 2002). It follows that wetland forests are They applied acetylene gas to the shoot, and more likely than marshes to be CH4 limited. In allowed it to diffuse through the stem, along some cases, CH4 oxidation efficiency is 100% with 02, into the root zone where it blocked and the site becomes a net CH4 sink (Harriss methanotrophy. Acetylene has the unfortunate ef of., 1982). property of causing irreversible inhibition, but it The path taken by a CH4 molecule en route to is far less expensive and equally effective as the atmosphere affects the chance it will be some alternative inhibitors (King, 1996a). oxidized. Oxidation is perhaps most likely when MethylHuoride (CH3F) is an attractive inhibitor CH4 diffuses across a soil or sediment surface because the effects are reversible, so that it can because of the potential for a wide aerobic zone, be used many times on the same site (Epp and and a correspondingly long residence time in the Chanton, 1993; Oremland and Culbertson, aerobic zone. One of the highest CH4 oxidation 1992a,b). However, CH3F inhibits acetotrophic rates that has been reported was measured in a methanogenesis and may therefore underestimate landfill soil that had high pore space CH4 CH4 oxidation. This is not necessarily a problem concentrations overlaid by >1.5m of aerobic when the gas is applied to shoots (King, 1996a; topsoil (Whalen et al., 1990). In most cases, the Megonigal and Schlesinger, 2002), but the potential for CH4 oxidation in the rhizosphere concentration and duration of application should will be limited by a narrow aerobic zone of be considered beforehand (Lombard! et al., 1997). perhaps 1-2 mm, or an aerobic zone may be absent Recently, diHuoromethane was shown to rever- altogether (e.g., King et al., 1990; Roura-Carol sibly inhibit CH4 oxidation without severely and Freeman, 1999). However, there is a notable affecting hydrogenotrophic or acetotrophic exception in which 0% from roots can penetrate methanogenesis (Miller et al., 1998). Stable a few centimeters into the soil (Pedersen et al., isotope ratios offer a nonintrusive means of 1995). Methane oxidation efficiency in the rhizo- estimating CH4 oxidation (e.g., Popp and sphere is likely to be lower than at the soil surface Chanton, 1999). because the residence time of CH4 in the rhizo- sphere is relatively short. For example, King (1996a) determined that 43% of the CH4 crossing the soil surface was oxidized compared to 27% of 8.08.4.7 Wetland Methane Emissions and Global Change the CH4 crossing the rhizosphere in a freshwater marsh. Methane bubbles (i.e., ebullition) bypass The influence of global-scale changes in climate, the CH4 oxidation zone altogether. Ebullition nutrient availability and land use on CH4 efflux occurs when the CH4 partial pressure exceeds a from freshwater wetlands is of much current critical threshold. It is a significant process in interest because they are one-third of all CH, aquatic sediments (Chanton and Martens, 1988; sources to the atmosphere. Methane emissions from Martens and Klump, 1980; Zimov et al, 1997) wetlands per se are not considered in detail here and is generally under-appreciated as a source of because the topic falls outside the primary scope of CH4 emissions from wetlands (Romanowicz et al., this review and many excellent reviews already 1993). Because plants efficiently evacuate CH4 to exist (Matthews and Fung, 1987; Aselmann and the atmosphere by passive diffusion or mass flow Crutzen, 1989; Bouwman, 1991; Neue, 1993; through stems (Chanton and Dacey, 1991), they Bartlett and Harriss, 1993; Bubier and Moore, improve CH4 oxidation efficiency by decreasing 1994; Bridgham et al., 1995; Denier van der Gon ebullition (Bosse and Frenzel, 1998; Grunfeld and and Neue, 1995; Vourlitis and Oechel, 1997; Brix, 1999). Understanding the proportion of CH4 Minami and Takata, 1997; Le Mer and Roger, efflux that occurs by these three pathways is 2001). In a compilation of 48 published studies of essential for modeling the response of natural wetland CH4 emissions (Le Mer and Roger, 2001), 1 l wetlands, rice paddies and landfills to manage- median rates were 0.43 kg CH4 ha~ d~ for peat- -1 _1 ment or climate change (Bogner et al., 2000). lands, 0.72 kg CH4 ha d for other nonagricul- J : A variety of techniques have been used to ture freshwater wetlands, and 1.0 kg CH4 ha~ d~ estimate the proportion of gross production that is for rice paddies. This ranking matches that consumed by CH4 oxidation (i.e., CH4 oxidation for primary production and is consistent with efficiency), most of which involve some degree of the large body of evidence that suggests manipulation. A common approach is to measure methanogenesis is closely coupled to the photo- methanotrophy in the presence and absence of synthetic carbon supply (Sections 8.08.4.3.4 02. However, this approach will underestimate and 8.08.4.3.5). In this section, we briefly the process if it simultaneously stimulates CH4 consider the influence of global change on CH4 production. An alternative approach is to apply a emissions. specific inhibitor of methanotrophy. This tech- Rising atmospheric C02 will influence CH4 nique was used by de Bont et al. (1978) to emissions from wetlands even in the absence of Methane 349 changes in climate. Temperate and tropical wet- and simultaneously lengthen the growing season land ecosystems, including rice paddies, consist- (Myneni et ah, 1997). Verville et ah (1998) ently respond to elevated C02 with an increase in concluded that plant community composition had photosynthesis and CH4 emissions (Dacey et ah, a larger effect on CH4 emissions than direct 1994; Megonigal and Schlesinger, 1997; Ziska changes in temperature in wet tussock tundra. A et ah, 1998), although one exception has been similar conclusion was reached in a study of global reported (Schrope et ah, 1999). One study change variables in a fen peatland (Granberg et ah, reported a strong linear relationship (r = 2001). These results are consistent with evidence 0.87-0.97) between photosynthesis and CH4 that the influence of plants on substrate availability emissions, suggesting that the elevated C02 (Section 8.08.4.3.5), CH4 ventilation, and metha- response is due to an increase in the supply of notrophy (Section 8.08.4.6.3) vary strongly among labile carbon compounds (Vann and Megonigal, plant species. 2003). By comparison, the responses of northern The effects of global warming on hydrology are peatland ecosystems have been inconsistent, with relatively uncertain compared to temperature. The several studies reporting negligible effects on long-term effects of a falling water table will be a photosynthesis, primary production or CH4 emis- decrease in CH4 emissions (e.g., Nykanen et ah, sions (Berendse et ah, 2001; Hutchin et ah, 1995; 1998) caused by lower CH4 production, lower Hoosbeek et ah, 2001; Kang et ah, 2001; Saarnio CH4 oxidation or both. Freeman et al. (2002) et ah, 1998, 2000). A likely reason for the absence concluded that reduced CH4 emissions following of a photosynthetic response to elevated C02 in a simulated drought in a peatland were due mainly these systems is nitrogen limitation. In a study of to its effect on methanogenesis, and only secon- tussock tundra, Oechel et ah (1994) found that the darily to methanotrophy. They proposed that CH4 elevated C02 treatment did not enhance photo- diffusion through the unsaturated soil surface was synthesis for more than a few months unless it was too rapid to permit an increase in CH4 oxidation crossed with a 4°C increase in air temperature. efficiency; this is plausible in peat soils because They suggested that warming relieved a severe they typically have a low bulk density at the nutrient limitation on photosynthesis by increas- surface, but is unlikely to apply to mineral soil ing nitrogen mineralization. Thus, the effects of wetlands unless they are on sands. Methane elevated C02 alone are expected to interact with emissions often recover slowly following drought nitrogen deposition and temperature. The influ- because of reduced methanogen populations and ence of elevated C02 may be particularly regeneration of alternative electron acceptors. important in tropical wetlands in which nitrogen Human activity has substantially increased the mineralization and methanogenesis are not tem- hydrologic import of oxidized nitrogen and sulfur perature-limited. These systems account for about into wetland ecosystems. Emissions of sulfur are 60% of all CH4 emitted from natural wetlands expected to double over the next 50 years (Rodhe, (Table 5). 1999), and anthropogenic N2 fixation will also Temperature changes are anticipated to be the increase dramatically (Section 8.08.5.1). This largest at latitudes above 45° N that hold 57% of all raises the possibility that methanogenesis will be wetland areas and 33% of the global soil carbon suppressed due to substrate competition between pool. Northern hemisphere wetlands contribute methanogens and denitrifiers or S04 reducers. It about 10% of all CH4 emissions (Table 5). Tropical is well known that methanogenesis is low in areas emit amounts of CH4 comparable to north- S04 -rich marine sediments, and S04 amend- ern-hemisphere wetlands (Aselmann and Crutzen, ments to rice paddies suppress CH4 production 1989; Matthews and Fung, 1987). This region is and emissions. However, even the low doses of also expected to warm in the coming decades S04 in can suppress methanogenesis (Kattenberg et ah, 1996), but efforts to quantify the (Disc and Verry, 2001). Gauci et ah (2002) contribution of southern hemisphere peatlands to suggested that 1990 levels of S04 deposition global carbon storage and fluxes are just beginning depressed CH4 emissions from northern wetlands (Thompson et ah, 2001). Increasing temperature by 5-17%. The effects of nitrogen fertilization on stimulates the decomposition and fermentation of CH4 emissions are variable and depend on the soil organic matter (Section 8.08.4.3.3), and many integrated responses of plants, methanogens and studies have reported temperature-dependent methanotrophs, which can offset one another increases in CH4 emissions as a result (Bartlett (Bodelier et ah, 2000b). Several studies have and Harriss, 1993). However, the most important shown a net increase in CH4 emissions with impacts of rising temperature on CH4 emissions nitrogen fertilization (Banik et ah, 1996; Lindau are likely to be realized through changes in the etah, 1991). species composition and metabolic activity of It is important to test how various factors that plants (Schimel, 1995). A temperature-induced force global changes interact because the results increase in nitrogen mineralization is expected to may be nonadditive or counterintuitive. Granberg stimulate plant productivity in northern bogs, et al. (2001) evaluated CH4 emissions from a fen 350 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes in response to crossed treatments of elevated air production than previously believed (Zehr and temperature, nitrogen deposition, and sulfur Ward, 2002). Furthermore, completely new meta- deposition. They found that SO|~ additions bolic pathways have been discovered in the past depressed CH4 emissions at ambient temperature, decade that will require a re-evaluation of nitrogen but not at elevated temperature. The reason for sinks in many ecosystems. this unexpected result was that pore-water S04" Although processes are the focus of this review, concentrations did not increase at elevated there have been significant strides made in temperature, presumably because of stronger quantifying N2 fixation, denitrihcation and other SO4- sinks such as plant uptake. nitrogen transformations at large scales (e.g., van Global changes in climate and nutrient avail- Breemen et al., 2002). There exist several recent ability are expected to alter the existing source- compilations and reviews of nitrogen oxide emis- sink balance of several radiatively active trace sions (Olivier et al., 1998) and denitrihcation rates gases including C02, CH4, and N20. Elevated (Barton et al, 1999; Herbert, 1999), including one 15 atmospheric C02 and falling water tables have the focused on studies that used the N isotope pairing potential to increase in technique (Steingruber et al., 2001). biomass if forests replace herbaceous commu- nities (Trettin and Jurgensen, 2003). However, an increase in primary production will translate into 8.08.5.1 Nitrogen in the Environment an increase in CH4 emissions due to improved Microbial transformations of nitrogen have a substrate supply (Whiting and Chanton, 2001). dramatic influence on the structure and function of Falling water tables will decrease CH4 emissions, ecosystems because nitrogen availability often but may increase N20 emissions (Martikainen establishes the upper limit of productivity, par- et al., 1995). At present, it is uncertain whether ticularly in terrestrial and marine ecosystems. feedbacks between climate change and wetland Widespread nitrogen limitation is somewhat metabolism will exacerbate or mitigate radiative ironic because the atmosphere is a vast reservoir forcing. This can be determined only by account- of N2, yet no can use nitrogen in this ing for changes in the current fluxes of C02, CH4, form. Plants require nitrogen in the form of NO^~ and N20, allowing for differences in their global (nitrate), NH^ (ammonium) or organic-N, and warming potential (e.g., Liikanen et al., 2002). most animals ultimately derive nitrogen from plants (except perhaps rock- snails (Jones and Shachak, 1990)!). These compounds have single or double covalent bonds that require far 8.08.5 NITROGEN less energy to sever than the triple bonds of N2, and they are referred to collectively as fixed, All forms of anaerobic metabolism are influ- combined or biologically available nitrogen. enced directly or indirectly by 02, but this is Prokaryotes are unique in their ability to convert particularly true of nitrogen metabolism. Many of N2 to fixed nitrogen (i.e., perform N2 fixation), and the organisms that anaerobically transform nitro- they participate in nearly every other important gen perform best under aerobic conditions, resort- nitrogen transformation (Figure 12). ing to anaerobic metabolism in order to cope with Among several factors that favor persistent low 02 availability (Tiedje, 1988). This physio- nitrogen limitation is the fact that prokaryotes logical versatility is favored by the fact that also convert fixed nitrogen back to N2 in the nitrogen oxides can replace 02 as terminal electron process of denitrification (Vitousek and Howarth, acceptors with only a small loss of energy yield 1991), a dominant form of anaerobic microbial (Table 1), and it is the reason that nitrogen-based metabolism. The global inventory of fixed nitrogen anaerobic metabolism is so cosmopolitan. Deni- is negligible compared to the atmospheric N2 trification occurs in even the driest ecosystems on reservoir, because denitrification has roughly Earth (Peterjohn, 1991). Whereas most of this balanced N2 fixation over eons. Denitrification is review is concerned with saturated soils and also a source of (N20) and nitric sediments as venues for anaerobic metabolism, oxide (NO), which are far less benign gases. the section on nitrogen includes work in upland Nitrous oxide is a greenhouse gas that is —300 soils. Here we focus on anaerobic metabolism times more effective at radiative forcing than C02 involving nitrogen, but it should be recognized that on a basis (Ramaswamy et al., 2001), and aerobic metabolism is important to the physiologi- accounts for —6% of the radiative forcing since ca. cal of many of the organisms concerned. 1750. Because its lifetime in the atmosphere is very Research on nitrogen-based anaerobic metab- long, N20 mixes in to the stratosphere where it olism has addressed denitrihcation to the exclusion promotes 03 destruction. is an of several other pathways. Recent advances make it ingredient in smog and promotes tropospheric O3 clear that some less-studied pathways make a production, which is a danger to human health. larger contribution to NO^ consumption or N2 Emissions of N20 and NO have increased in the Nitrogen 351 from sewage effluents and is used widely in wastewater treatment facilities. Wetland and aquatic ecosystems often provide the same service for the cost of protecting these areas as a (Bowden, 1987; Groffman, 1994).

8.08.5.2 Nitrogen Fixation 60 O In a sense, all biological N2 fixation is anaerobic because the nitrogenase enzyme is strongly inhibited by 0 . As a result, many aerobic 3 1 2 diazotrophs (N2-fixing microorganisms) have specialized structures (i.e., heterocysts) for keep- 2- ing the site of N2 fixation 02-free. In the absence of such structures, N2 fixation varies with 02 concentrations, often peaking at night when oxygenic photosynthesis is absent (Bebout et al., 1987; Currin et al, 1996). Dinitrogen fixation is an energetically expens- ive reaction and is inhibited by micromolar levels - exchange of NH4", which requires comparatively little energy to assimilate. The energetic demands of - abiological reaction the process is one reason why many diazotrophs are either photosynthetic themselves or occur in Figure 12 Major reduction-oxidation reactions symbiotic relationships with plants. However, involving nitrogen. The reactions are numbered as diazotrophy is certainly not restricted to these follows: (1) mineralization, (2) ammonium assimila- groups; it occurs widely among heterotrophic and tion, (3) nitrification, (4) assimilatory or dissimilatory nitrate reduction, (5) ammonium oxidation, (6) chemoautotrophic microorganisms, and both aero- oxidation, (7) assimilatory or dissimilatory nitrate reduc- bes and anaerobes (Lovell et al., 2001). tion, (8) assimilatory or dissimilatory nitrite reduction, A provocative link between diazotrophy and (9) denitrification, (10) chemodenitrification, (11) anae- methanogenesis was recently proposed by Hoehler robic ammonium oxidation, and (12) dinitrogen fixation et al. (2001), who reported that cyanobacteria (after Capone, 1991) (reproduced by permission of (formerly blue-green algae) in a coastal mudflat ASM Press from Microbial Production and Consump- were a "hot spot" of H2 production, presumably tion of Greenhouse Gases: Methane, Nitrogen Oxides, due to a side-reaction of the nitrogenase enzyme and Halomethanes, 1991). system. In some cases, H2 accumulated underneath the mats to levels that were high enough to support methanogenesis, despite >50 mM SO4 . How- past decades from aerobic and anaerobic sources ever, the more important link to anaerobic due to the widespread use of fertilizers and fossil metabolism was the possibility that H2 production fuel combustion. by diazotrophs promoted the oxidation of the Human activity has roughly doubled the rate of primordial Earth. The loss of diazotroph-produced N2 fixation (Vitousek et al., 1997), and thereby hydrogen to space may have promoted the enhanced microbial nitrogen transformations (Gal- accumulation of oxidants, thereby initiating the loway et al., 1995). The enhanced metabolism of transition from an anaerobic to an aerobic N2-fixing bacteria through cultivation of planet before there were significant amounts of was an intended outcome of this activity, but many atmospheric 02. other effects have been unintended. Transport of nitrogen through water, air, and sediments has increased the productivity of adjacent ecosystems, 8.08.5.3 Respiratory Denitrification resulting in terrestrial and aquatic The term denitrification is often used in the (Meyer-Reil and Koster, 2000). For example, general literature to describe any process that nitrogen export from the Mississippi basin converts nitrogen oxides (NO^~ or NO^) to has been linked to the expansion of anaerobic reduced nitrogen gases (N20 or N2). In the context sediments in the Gulf of Mexico (Rabalais et al., of microbial metabolism the term is used more 2002). Elevated NO^~ levels are a direct human narrowly to describe a specific respiratory path- health threat in newborn infants and adults way. Denitrification is the most common form of deficient in glucose-phosphate dehydrogenase anaerobic respiration based on nitrogen. Energy is (Comly, 1945; Payne, 1981). Denitrification is conserved by coupling electron transport phos- an effective means of removing fixed nitrogen phorylation to the reduction of nitrogen oxides 352 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes located outside the cell. Because nitrogen is not assimilated into the cell, the process is dissim- N0 —^ NO^» N O^N ilatory. Respiratory denitrification is more ener- 2 2 2 getically favorable than Fe(III) reduction, SO4" Nitrification reduction or methanogenesis (Table 1), and it Denitrification tends to be the dominant form of anaerobic carbon NnO metabolism when NO^ or N02 are available. Many microorganisms also reduce nitrogen oxides without conserving energy for growth, in which NH3—*-NH2OH—>• N02 - -NO- -N,0- case the process is nonrespiratory. Nonrespiratory Nitrifier Denitrification nitrogen oxide reduction may be assimilatory (i.e., assimilated into the cell) or dissimilatory. Dis- Figure 13 Relationships between three pathways of similatory NO^~ reduction to NH4" is a potentially inorganic nitrogen oxidation and reduction (Wrage important sink for NO^ in most ecosystems et al., 2001) (reproduced by permission of Elsevier from (Section 8.08.5.4). SoilBiol. Biochem. 2001, 33, 1723-1732).

8.08.5.3.1 Denitrifier diversity and metabolism acceptor. A variety of intermediate compounds are produced as the terminal electron acceptor is Denitrifying bacteria are aerobes that substitute reduced stepwise to N2 (Figure 13), including the NO3" (or NO2") for 02 as the terminal electron solute nitrite (N02), and the gases nitric oxide acceptor when there is little or no 02 available (Payne, 1981; Firestone, 1982). Aerobic respir- (NO) and nitrous oxide (N20) (Ye et al, 1994). ation yields more free energy than NO^~ respir- With a few exceptions, a single organism can complete the entire series of reductions from NO^~ ation, and it is favored metabolically because 02 inhibits key denitrification enzymes. Although to N2 (Tiedje, 1988). Nonrespiratory denitrifiers tend to produce N20 rather than N2 (Tiedje, 1988). some organisms can denitrify at 02 concentrations up to 80% of air saturation (Robertson and Because these intermediate compounds are Kuenen, 1991), the process is most rapid in situ required for metabolism, it is reasonable to under anaerobic conditions provided there is a assume that denitrifiers can also consume them. Consumption of external NO and N20 creates supply of NO^. The critical 02 concentration where denitrifiers switch to mostly anaerobic many variations on the basic denitrification respiration is roughly <10 jjumol L_1 (Seitzinger, scheme. Nitrite, NO, and N20 can serve as 1988; Tiedje, 1988). Denitrification activity sur- terminal electron acceptors rather than intermedi- vives well in aerobic soils, even without new ates, and they can also replace N2 as terminal enzyme synthesis or cell growth (Smith and products. It was recently observed that nitrogen Parsons, 1985). It persists in habitats that lack dioxide gas (N02) is a terminal electron acceptor leading to N production by the nitrifying 02 or NOJ presumably because denitrifiers can 2 maintain themselves with a low level of fermenta- bacterium eutrophica (Schmidt tion (J0rgensen and Tiedje, 1993). and Bock, 1997). Denitrifying bacteria use all three energy Respiratory denitrification is widespread among sources available to bacteria including organic various physiologic and taxonomic groups of carbon compounds (), inorganic bacteria (Tiedje, 1988; Zumft, 1997), and may be compounds (lithotrophs), and light (phototrophs). considered the most versatile form of anaerobic The dominant populations of denitrifiers appear to metabolism. The only prokaryotic group that does be organotrophs such as and not include denitrifying members is the Entero- Alcaligenes. Others are able to ferment or use bacteriaceae (Zumft, 1997). Interestingly, most H2, sulfur, or NH4" as energy sources. Some members of this group are capable of dissimilatory denitrifiers are also N2 fixers that grow in nitrate reduction to ammonium (Tiedje, 1994), a association with plants, including members of process that competes with denitrification for the genus (Tiedje, 1988), although it is NO^T- With a few exceptions (Zumft, 1997), not clear if many of these strains are capable of denitrification is absent among the gram-positive growing solely on NO^ as an electron acceptor. bacteria and obligate anaerobes. Many denitrifiers are chemolithotrophs, including The use of 16S rONA gene sequences to detect the obligate chemolithotroph Nitrosomonas euro- the presence of denitrifying bacteria in microbial paea (Poth and Focht, 1985; Poth, 1986), which is communities is limited by the high phylogenetic better known as a nitrifying microorganism. diversity of this group. Prieme et al. (2002) In its most familiar form, denitrification characterized the heterogeneity of gene fragments requires organic carbon for the electron donor coding for two types of {nirK and NO3" or NO^ for the terminal electron and nirS) in an upland soil and a wetland soil. Nitrogen 353

They concluded that both soils, but particularly the denitrification in the presence of 02, it occurs only wetland soil, had a high richness of nir genes, most under anaerobic or microaerobic conditions in of which have not yet been found in cultivated situ. The presence of water is nearly a prerequisite denitrihers. This contrasts with a similar survey by for denitrification activity because it slows 02 Rosch et al. (2002) in which denitrification was not diffusion by a factor of 104 compared to air. In a genetic trait of most of the uncultured bacteria in a upland soils, abundant air-filled pore spaces hardwood forest soil. permit rapid gas exchange and 02 concentrations Fusarium oxysporum and a few other fungi are typically 21% to depths of 1 m or more have the capacity to denitrify NO^~ or N02 to (Megonigal et al., 1993). Water reduces the 02 N20 (Shoun and Tanimoto, 1991; Usuda et al., supply by blocking a fraction of the pores, thereby 1995). Their denitrifying enzyme system is increasing the effective distance (i.e., tortuosity) coupled to the mitochondrial electron transport an 02 molecule travels to a given microsite chain where it produces ATP (Kobayashi et al., (Renault and Sierra, 1994; Figure 14). As the 1996). They do not grow under strictly anaerobic soil water content increases, some sites are conditions, but require a minimal amount of 02 blocked completely so that there is no air-filled (Zhou et al., 2001). F. oxysporum can also pathway to the atmosphere. This explains why produce N2 using N02 as an electron acceptor denitrification rates in soils are often positively (Tanimoto et al, 1992). Zhou et al. (2002) have related to water content (Drury et al., 1992; reported that F. oxysporum is capable of at least Groffman and Tiedje, 1991). For example, three types of metabolism depending on 02 denitrification enzyme activity in a temperate availability. Aerobic respiration is favored when hardwood forest was higher in wet years than dry 02 is abundant, denitrification occurs under years (Bohlen et al., 2001). microaerobic conditions, and NOJ is reduced to Soil texture is a good predictor of denitrification NH^ during fermentation under anoxic con- rates at the landscape scale (Groffman, 1991), in ditions. There have been virtually no field studies part because it captures the interaction between of N03-based metabolism by these organisms. water content and soil porosity with respect to gas However, Laughlin and Stevens (2002) selec- and solute diffusion path length. At a given soil tively inhibited fungal activity using cyclohex- water content, the small pores found in clayey imide and observed a decrease in N20 fluxes of soils are more likely to be blocked than the 89% in a grassland soil; the antibiotic streptomy- relatively large pores found in loam and sand cin decreased N20 flux by 23%. Because fungi soils. Soil texture also influences temporal varia- represent a large portion of the microbial biomass bility in soil water content because it largely in soils (Ruzicka et al., 2000), they may be an establishes the water infiltration rate and water important source of N20 in some terrestrial holding capacity (de Klein and van Logtestijn, ecosystems. 1996; Sexstone et al., 1985). A particularly useful expression for soil water content is percent water- filled porosity (Williams et al., 1992), which is the ratio of volumetric water content to total soil 8.08.5.3.2 Regulation of denitrification porosity. Based on a compilation of denitrification Foremost on the long list of factors that regulate studies in agricultural and forest soils, Barton et al. denitrification are those that influence the avail- (1999) determined that denitrification increases ability of 02, NO^~, and organic carbon, the dramatically above 65% water-filled porosity on primary substrates that are metabolized by average, with higher values for sandy soils (74- denitrifying bacteria. Some factors (e.g., nitrifica- 83%) than clayey soils (50-74%). tion rate) influence substrate pools directly, while Carbon availability largely controls 02 con- other factors (e.g., soil water content) influence sumption rates, either directly by fueling aerobic the transfer of substrates to sites where denitrifica- heterotrophic respiration, or indirectly by support- tion occurs. These factors have been reviewed ing the anaerobic production of reductants, such as previously (Cornwell et al, 1999; Herbert, 1999; Fe(II), that subsequently react with 02. Even small Seitzinger, 1988). Although it is convenient to amounts of organic carbon can produce anaerobic consider the factors that regulate denitrification conditions in soils that are partially or completely individually, the ability to predict in situ rates saturated. In upland soils, anaerobic sites occur requires an understanding of their complex inside soil aggregates. Hojberg et al. (1994) used interactions (Parton et al., 1996; Strong and 02 and N20 microsensors to demonstrate simul- Fillery, 2002). taneous 02 consumption and denitrification in the Molecular oxygen (02) is the physiologically surface of soil aggregates (Figure 15). Anaerobic preferred terminal electron acceptor for denitrify- microsites are particularly pronounced near ing bacteria, and its presence represses denitrifica- sources of organic carbon such as roots or detritus. tion enzyme synthesis and activity. Although For example, Parkin (1987) found that a single some microorganisms are capable of performing leaf constituting 1% of the soil mass supported 354 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

%.g.;:%%%.

'•?:•'•'y t '*"''" ":'-^-: o. Diffusion Path

r» An "2 Tortuous Diffusion Path

Figure 14 The physical and chemical factors that regulate substrate diffusion to denitrifying bacteria (Strong and Fillery, 2002) (reproduced by permission of Elsevier from Soil Biol. Biochem. 2002, 34, 945-954).

Some forms of organic carbon support higher denitrification rates than others due to differences in carbon quality. For example, fresh pine needles supported higher denitrification rates than senes- cent pine needles in a riparian wetland (Schipper et al., 1994), presumably because the former had more labile compounds such as soluble carbo- hydrates, van Mooy et al. (2002) interpreted depth-dependent changes in the ami no acid content of sinking particulate organic carbon in the Pacific Ocean as evidence that denitrifiers preferred to metabolize nitrogen-rich amino acids. If this is the case, calculations of denitrification in the eastern tropical North Pacific Ocean may be 9% higher than previous estimates that were based 100 200 300 400 on typical Redfield ratios. 02 or N20 (uM) Denitrification can be limited by carbon avail- ability when 02 is absent and NO^ is abundant. Figure 15 Profiles of 02 and N20 with depth from the Additions of glucose stimulated denitrification in surface of a soil aggregate. The profiles were 11 of 13 agricultural soils that were presumably determined with microelectrodes (source H0jberg ef af., 1994). fertilized (Drury et al., 1991). Similar observations have been made in water columns (Brettar and Rheinheimer, 1992), marine sediments (Slater and 85% of the denitrification. Such "hot-spots" are Capone, 1987), river sediments (Bradley et al., one explanation for the notoriously high spatial 1995), aquifers (Smith and Duff, 1988; Obenhuber variability in N20 production measured in upland and Lowrance, 1991), wastewater treatment wet- soils (Parkin, 1987) and, to a lesser extent, in lands (Ingersoll and Baker, 1998), and forested estuarine sediments (Kana et ah, 1998). Although wetlands (DeLaune et al, 1996). Tiedje (1988) microsites in upland soils are a small fraction of proposed that the major influence of carbon on the total soil volume, they contribute perhaps 70% in situ denitrification is to promote anaerobic of global N20 emissions (Conrad, 1996). conditions. Nitrogen 355 Due to a high demand for nitrogen by all (Section 8.08.5.3.5). Heterotrophic oxidation of organisms, the NO^ pool in upland ecosystems is NH4 or organic-N to N02 and NO3" has been commonly small. In wetland and aquatic ecosys- observed in fungi, but bacteria are generally tems, NO^" availability may be limited further by considered the dominant in situ source of NO^ anaerobic conditions and low nitrification rates. in most ecosystems. The factors regulating Nitrate addition studies have demonstrated that nitrification in marine sediments were reviewed NO^ availability limits denitrification in a variety by Henriksen and Kemp (1988). of aquatic ecosystems (Lohse et ah, 1993; Because nitrification occurs only under aerobic Nowicki, 1994). In upland soils, NO^~ limitation (or microaerobic) conditions and denitrification can be an indirect effect of low soil water content, under anaerobic conditions, the two processes are which increases the length of the diffusion path- spatially separated. However, if the sites where way between aerobic sites where nitrification these processes occur are sufficiently close takes place and anaerobic microsites (Figure 14). together, NO^ transport and consumption are Thus, low soil water content suppresses denitrifi- very rapid and the overall process is considered to cation in upland soils by simultaneously increas- be coupled nitrification-denitrification. On the ing the 02 supply and decreasing the NO^~ supply. basis of a literature review, Seitzinger (1988) Seitzinger (1994) concluded that NO^-limitation concluded that nitrification is generally the major of denitrification in eight riparian wetlands was an source of NO^ for denitrification in river, lake, indirect response to limited organic carbon and coastal sediments. The same is likely to be availability. Organic carbon governed mineraliz- true of nonagricultural soils that are largely ation rates and, therefore, the NH4" supply to dependent on mineralization and atmospheric . The relationship between NO^ deposition for fixed nitrogen. concentration and denitrification rate often Denitrification can be uncoupled from nitrifica- approximates a Michaelis-Menten function. tion when NO3" is delivered to soils and sediments Half-saturation concentrations generally range from outside sources such as overlying water, from 27 pM to 53 pM for stirred marine sedi- fertilizers, and atmospheric deposition. Fertilizers ments (Seitzinger, 1988); the values for intact are a source of NO3" to groundwater (Spalding and soils can be 30-fold higher because of diffusion Exner, 1993), which in turn is a source of NO3" to limitation (Schipper et ah, 1993; Strong and riparian forests (Hill, 1996), tidal marshes (Tobias Fillery, 2002). et ah, 2001a), and estuarine sediments (Capone and Bautista, 1985). Surface water runoff carries NO3" directly to rivers, lakes, estuaries and oceans (Jordan and Weller, 1996; Joye and Paerl, 1993). 8.08.5.3.3 Nitrification - denitrification Because of high NO^ loading, denitrification rates coupling in estuarine sediments are often directly pro- Microbial metabolism and anthropogenic portional to the NO^ concentration in the activity are the primary sources of NO^ used by overlying water column (Kana et ah, 1998; denitrifying bacteria. Nitrate is a waste product of Nielsen et ah, 1995; Pelegri et ah, 1994), chemoautotrophic nitrification, a series of two indicating that denitrification is controlled by the dissimilatory oxidation reactions, each performed NO^ diffusion rate across the aerobic sediment by a distinct group of bacteria. In the first step, surface. NH4" is oxidized to NO^ by Nitrosomonas The response of denitrification to certain europaea and other "Nitroso-" genera including environmental factors varies according to how Nitrosococcus and Nitrosospira. Although >95% strongly the process is coupled to nitrification. An of the total NH3 + NHJ" pool is in the form of increase in 02 penetration into sediments can NHJ" at pH < 8, these organisms are also known stimulate coupled nitrification-denitrification as oxidizers because NH3 is used at the by enhancing NO^ availability (Rysgaard et ah, enzymatic level. In the second step, N02 is 1994), but simultaneously depress water oxidized to NO^ by nitrite-oxidizing bacteria column-supported denitrification because of the belonging to "Nitro-" genera such as increased distance that NO3" must diffuse to reach and . The two genera constitute the the anaerobic layer (Cornwell et ah, 1999). Nitrobacteriaceae (Buchanan, 1917). Ammonium Risgaard-Petersen et ah (1994) reported that a is also oxidized by microorganisms that specialize photosynthesis-induced increase in 02 penetration in other types of metabolism. The most important into an estuarine sediment doubled the coupled of these are methanotrophic bacteria (Section nitrification - denitrification rate, but reduced the 8.08.4.6.1), which were estimated to mediate 44- amount of water column-supported denitrification 85% of the NH4" oxidation in a rice paddy soil by 50%. Because the contribution of coupled (Bodelier and Frenzel, 1999). In addition, metha- nitrification - denitrification to total denitrifica- notrophic bacteria produce and consume NO (Ren tion at this site was small (—16%), the net effect et ah, 2000), a common nitrification product of deeper 02 penetration was to reduce 356 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes denitrification. An increase in 02 penetration in Nftt (Lin et ah, 2002). In upland systems, plants North Sea sediments had an entirely different are more likely to have an overall negative effect outcome (i.e., higher overall denitrification) on denitrification due to competition for NH^ and because the sediment, rather than the water NO^~, and consumption of water (Tiedje, 1988). column, was the dominant NO^ source supporting denitrification (Lohse et ah, 1993).

8.08.5.3.5 N20 and NO fluxes

8.08.5.3.4 Animals and plants Classical coupled nitrification-denitrification requires NFEj", organic carbon, aerobic conditions, Animals and plants influence denitrification in and anaerobic conditions. It involves three distinct wetland and aquatic ecosystems by altering the populations of microorganisms, some of which are availability of 02, NO^, and carbon. In organic- heterotrophic and others autotrophic. As a result, rich sites, nitrification rates tend to be limited by regulation of the process is rather complex and the the depth of 02 penetration (Kemp et ah, 1990). relative proportions of NO^~, NO, N20, and N2 as Benthic invertebrates enhance pools of 02 and end products varies widely with environmental NO^ by burrows and irrigating them with and ecological conditions. overlying water (Pelegrf et ah, 1994). Macro- Several factors can cause the reduction of NO^~ faunal tubes and burrow walls support higher to N2 to be incomplete, resulting in the pooling of potential rates of nitrification than nearby surface NO and N20. Denitrification and nitrification sediments, and contribute >25% of benthic enzymes are inhibited by H2S (Joye and Holli- nitrification is some coastal ecosystems (Black- baugh, 1995) and 02. The sensitivity of denitrify- burn and Henriksen, 1983;Kristensenef ah, 1985). ing enzymes to 02 inhibition is inversely The degree to which infauna enhance potential proportional to the oxidation state of the nitrification varies among worm species due to nitrogen substrate (Figure 12), increasing in the variations in irrigation behavior (Mayer et ah, order: NO-T reductase < N02 reductase < NO 1995). Nitrification and denitrification are coupled reductase < N20 reductase (Dendooven and in burrow wall sediments just as they are in Anderson, 1994; McKenney et ah, 1994). Pooling surface sediments (Sayama and Kurihara, 1983). can be caused by an excess of NO^ relative to Animals also enhance denitrification by concen- organic carbon, or an imbalance in the kinetics of trating organic matter into faecal pellets, thereby the various steps. Some organisms lack key creating "hot-spots" of 02 demand. enzymes in the sequence and release NO or N20 Aquatic plants influence sediments in much the as waste products (Tiedje, 1982). Reduction to N2 same manner as animals. Reddy et ah (1989) appears to be favored over N20 at circumneutral applied : N-NH4" to the saturated root zone of pH (Simek et ah, 2002). three emergent aquatic macrophytes and recov- Nitrogenous gas emission from soils varies 15 ered -25% of the label as N2 over 18 d. Since strongly with soil water content (Williams et ah, 15 N2 was not recovered from the unplanted 1992). The water content at which efflux from soils controls, it was apparent that coupled nitrifica- peaks generally increases in the order: tion-denitrification was stimulated by the rhizo- NO > N20 > N2. Yang and Meixner (1997) sphere. Similar observations have been reported reported that NO fluxes peaked at —20% water- for submerged aquatic plants (Caffrey and Kemp, filled pore space in a grassland soil, which is in 1991; Christensen and S0rensen, 1986) and qualitative agreement with other studies (Otter (Law et ah, 1993). Risgaard-Petersen et ah, 1999; Potter et ah, 1996). Nitrous oxide and Jensen (1997) reported that denitrification emissions peak at higher levels of soil moisture was enhanced in the presence of a submerged than NO emissions (Drury et ah, 1992), and N2 aquatic macrophyte by six-fold. This impressive emissions are highest in saturated soils. This effect was due in part to the fact that 02 was pattern is caused by differences in the 02 released well below the sediment surface, creat- sensitivity of denitrifying enzymes, and by the ing a zone of nitrification capped above and influence of soil water content on gas diffusion. below by zones of denitrification. The result was High soil water content restricts the diffusion of a highly efficient coupling between the two gases and enhances the diffusion of solutes processes. The influence of plants on anaerobic (Section 8.08.5.3.2). Because nitrifying bacteria metabolism is apparently species- or site-specific require both a gas (02) and a solute (NH4), the because studies of other marine angiosperms optimal availability of substrates occurs where have found no effects on denitrification soils are wet, but not saturated (Williams et ah, (Rysgaard et ah, 1996). In addition to introducing 1992). Such conditions favor N20 production 02, plants introduce organic carbon in the form of because nitrification and denitrification can be root exudates (Lynch and Whipps, 1990) and simultaneously producing N20 (Stevens et ah, detritus, and compete with microorganisms for 1997). The influence of soil water on NO and N20 Nitrogen 357 emissions is complicated by that fact that these (Rice and Tiedje, 1989), and not regulated by gases are also consumed by microorganisms. 02. These factors have the opposite effect on Water promotes microbial consumption of NO dissimilatory NO^~ reduction to NH4. Assimila- and N20 by restricting diffusion to the atmosphere, tory NO^ reduction by microorganisms is usually thus increasing their residence time in the soil assumed to be negligible due to inhibition by (Davidson, 1992; Skiba etal., 1997). NH|. It is difficult to distinguish between nitrification and denitrification as sources of NO and N 0. Selective inhibition of nitrification by 8.08.5.4.1 Physiology and diversity of 2 DNRA bacteria 10 Pa C2H2 (acetylene) and denitrification by 10 kPa C2H2 (Davidson et al, 1986) has widely Dissimilatory nitrate reduction to ammonium been used, but it suffers from a number of is an anaerobic pathway that is insensitive to problems. These include the possibility of C2H2- NKt and yields energy. The first step of the insensitive denitrifiers (Dalsgaard and Bak, 1992), process is termed nitrate respiration because it is and NO scavenging by 02 (Bollmann and Conrad, coupled to electron transport phosphorylation that 1997; McKenney et al., 1997). Techniques based generates ATP: on * N or 13N tracing present different problems (Boast et al., 1988; Arah, 1997), but are favored in NO, + H, ^ No; H?0 (13) more recent studies. The contribution of nitrifica- tion to NO and N 0 production is highest under 2 Nitrate respiration is widely found among micro- the conditions that favor nitrification, namely, organisms, some of which further reduce NO^ to moderate 0 partial pressure and high NH^ 2 NH|: concentrations (Avrahami et al., 2002). Nitrifica- tion is generally considered to be the dominant + source of NO upland in soils, but there are studies NO; 3H2 + 2H — NHJ + 2H20 (14) that suggest the opposite (reviewed by Ludwig et al., 2001). The dominant source may vary with This second step is not coupled to energy environmental factors such as pH (Remde and production except in a few species that are not Conrad, 1991). expected to be abundant in situ, such as Campy- Nitric oxide and N20 are also produced lobacter, Deulfovibrio, and (Tiedje, abiotically by chemical decomposition of N02 1988). The overall DNRA reaction transfers 8e~ (Hooper and Terry, 1979). The reaction is and yields 600 kJ mol-1 NO^~ (Tiedje, 1994): favorable at low pH and yields mainly NO (van Cleemput and Baert, 1984), although N 0 can also + 2 NOf + 4H2 + 2H — NHj + 3H20 (15) be produced (Martikainen and De Boer, 1993). Abiotic production of NO and N20 is assumed to The second step in which NO^ is reduced to NH4" be relatively unimportant in most ecosystems is ecologically significant because it ensures that (e.g., Webster and Hopkins, 1996). the nitrogen can be retained in the ecosystem for plant uptake, microbial assimilation or adsorption 8.08.5.4 Dissimilatory Nitrate Reduction to cation exchange sites; otherwise, NO^ would to Ammonium (DNRA) accumulate and be subjected to denitrification. Because NO^ reduction to NH4" does not Nitrate reduction studies have focused over- normally yield energy, it presumably has other whelmingly on denitrification at the expense of physiological or ecological advantages. Perhaps other NO^ sinks such as dissimulatory NO^ its most likely role is to serve as an electron sink reduction to NHJ" (also known as DNRA or nitrate for the regeneration of NADH from NADH2 ammonification). The ecological implications of (Bonin, 1996). reducing NO^ to NFEj", versus N2 are vastly DNRA bacteria can be aerobic, facultatively different because NHf" is more readily retained in anaerobic or obligately anaerobic. As a group they the ecosystem, and it is a form that is readily are unlike the denitrifying bacteria in that most assimilated by biota. Thus, DNRA contributes to species are fermentative (Bonin, 1996; Tiedje, eutrophication by reducing the quantity of fixed 1988). DNRA bacteria are abundant in aerobic nitrogen that is returned to the atmosphere as N2. soils and other environments that do not favor There are two pathways by which NO^ can be DNRA activity per se. This suggests that they can reduced to NKJ\ In assimilatory NO^ reduction, compete with other fermentative bacteria or NO^~-N is taken up by microorganisms or plants, aerobes for carbon substrates (Tiedje, 1988). reduced to NH4, and assimilated into organic Much less is understood about the diversity and nitrogen compounds such as amino acids. The physiology of DNRA bacteria than denitrifiers, process is inhibited by low concentrations despite the fact that they are sometimes the larger (0.1 JJLLL-1) of NH4" or organic nitrogen of the two dissimilatory NO^ sinks (Table 7). 358 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes Table 7 Examples of the contribution of dissimilatory nitrate reduction to ammonium (DNRA) to total dissimulatory nitrate reduction in various freshwater, marine, and terrestrial ecosystems. The ranges are inclusive of all sites and treatments reported for which appropriate data could be drawn, including those that manipulated substrate levels. Ecosystem %DNRAa Method Citation

Marine sediment 79-93 ABT + estimate of DNRA Bonin (1996) Marine sediment 18-97 15N assay + ABT Bonin et al. (1998) Marine sediment 98 l5N assay + ABT Gilbert et al. (1997) Marine sediment 33 l5N assay Goeyens et al. (1987) Marine sediment 46-68 l5N assay + ABT S0rensen (1978) Marine sediment 0-18 l5N assay + ABT Kaspar et al. (1985) Estuarine sediment 5-30 l5N assay + ABT Kaspar (1983) Estuarine sediment 0-85 l5N assay Christensen et al. (2000) Estuarine sediment 73-82 ABT + estimate of DNRA J0rgensen and S0rensen (1985) Estuarine sediment 66-75 ABT + estimate of DNRA J0rgensen and S0rensen (1988) Estuarine sediment 6-75 15N assay + ABT J0rgensen (1989) Estuarine sediment 10-61 l5N assay Koike and Hattori (1978) Estuarine sediment 2-3 l5N assay Felegrietal. (1994) Mangrove soil 3-100 l5N assay Riviera-Monroy (1995) Brackish marsh soil 7-70 l5N assay Tobias et al. (2001a) Brackish marsh soil 1-7 l5N assay Tobias et al. (2001b) River sediment0 6-10 l5N assay Kelso ef a/. (1997) Aquifer sediment 22-60 l5N assay Bengtsson and Annadotter (1989) Rice paddy soil 1-56 l5N assay Buresh and Patrick (1978) Rice paddy soils 4-12 l5N assay Yin et al. (2002) Boreal forest 40-93 l5N assay + ABT Bengtsson and Bergwall (2000) Calcareous clay soil 4-19 l5N assay Fazzolari et al. (1998) Clay loam soil 1-2 l5N assay Chen etal. (1995) Silt loam soil 1-77 l5N assay + ABT deCatanzaro et al. (1987) Silt loam soilc 11-30 l5N assay Stanford et al. (1975) Wet tropical forest 71-83 l5N assay Silver et al. (2001)

* Percent DNRA calculated as (NO, reduction to NH4)/(N03 reduction to NH4 + N2 + N20). ' ABT is the acetylene block technique. c Used last reported incubation time point.

8.08.5.4.2 DNRA versus denitrification perhaps because of increased carbon availability due to root exudates. Nonetheless, competition DNRA and denitrification occur simultaneously between the two groups of organisms has not been and can be expected to compete for carbon and adequately studied and there are exceptions to this NO^. Tiedje et al. (1982) proposed that the generalization that suggest the influence of other partitioning of NO^~ to N2 versus NH^ is a factors. function of the carbon: NO^ ratio. They reasoned A series of laboratory and field studies showed that a combination of abundant electron donors that low temperatures favor denitrifying bacteria (i.e., carbon) and limited electron acceptors (i.e., while high temperatures favor DNRA bacteria in NO^) should favor organisms that use electron temperate salt marsh and estuarine sediments of acceptors most efficiently. In this case, DNRA has the Colne estuary, UK (King and Nedwell, 1984; a competitive advantage because it transfers 8 Ogilvie et al., 1997a,b). It is not yet clear whether moles of electrons per mole of NO^~ reduced, this effect can be generalized to other systems. while denitrification transfers five moles of Field estimates of NO^ partitioning to the two electrons. Nitrite respiration alone transfers two processes in a North Sea estuary were best moles of electrons. Thus, denitrifying bacteria explained when it was assumed that DNRA should out compete DNRA bacteria for organic activity was favored by both high and low carbon when carbon is limiting. Several studies temperatures (Kelly-Gerreyn et al., 2001). Tem- support the generalization that high labile carbon perature effects may explain why the relative availability and/or low NOJ availability favor importance of DNRA versus denitrification DNRA over denitrification (Bonin, 1996; decreased from May to October in a temperate Fazzolari et al, 1998; King and Nedwell, 1985; tidal marsh despite a three-fold increase in Nijburg et al, 1997; Yin et al, 2002). In wetlands, dissolved carbon concentrations (Tobias et al., the presence of emergent aquatic macrophytes 2001a). appeared to stimulate DRNA activity (Nijburg DNRA activity has been assumed to be et al, 1997; Nijburg and Laanbroek, 1997), important only under highly reduced conditions Nitrogen 359

(Eh = - 200 mV; Buresh and Patrick, 1981). (Figure 12), the substrate and environmental However, there is growing evidence that DNRA controls on anaerobic NKJ~ oxidation are likely can be important in relatively oxidized environ- to be quite different from those for the classical ments. DNRA accounted for -75% of NO^ sequence of reactions. The contribution of these reduction in a humid tropical rainforest where processes to overall denitrification is largely the soil 02 content was 15% (Silver et al., 2001). unknown, but the limited evidence suggests they DNRA activity was stimulated by the amphipod may be quantitatively important in both natural Corophium volutator (Pelegri et al., 1994), an ecosystems and wastewater treatment processes. organism that increases redox potential through burrowing. Fazzolari et al. (1998) suggested that DNRA bacteria are less 02 sensitive than 8.08.5.5.1 Anammox denitrifying bacteria. Clearly, highly reduced conditions are not a prerequisite for DNRA to be Ammonium oxidation linked to N02 (nitrite) an important NO^ sink. reduction was first recognized in a wastewater DNRA has an abiotic equivalent reaction that treatment system and patented under the process can proceed at rates comparable to the biotic name anammox for anaerobic ammonium oxi- reaction in the presence of "green rust" (Hansen dation (Mulder et al., 1995). The discovery was et al., 1996) or trace metals such as Cu(II) (Ottley motivated by observations of simultaneous NH4" et al., 1997). Green rusts are Fe(II)-Fe(III) and NO^ losses balanced by N2 production under precipitates that form in nonacid, Fe(II)-rich anaerobic conditions. The existence of a "miss- soils and sediments (Hansen et al., 1994). A ing" chemolithotrophic organism capable of similar reaction may have made a significant coupling NKt oxidation to either N07 or N02 contribution to the NKJ~ inventory on the prebiotic reduction had been predicted nearly two decades Earth (Summers and Chang, 1993). earlier based on thermodynamic considerations In the past, DNRA has been considered to be (Broda, 1977). The anammox reaction was inconsequential in nonmarine ecosystems (Tiedje, initially proposed to involve NO

8.08.5.5.2 Nitrifier denitrification only under conditions of low 02 concentration or anaerobiosis in N. europaea. NH3-oxidizing bacteria are typically character- Very few studies have attempted to establish ized as a NO^ source for other bacteria that whether these alternative metabolic pathways in ultimately produce NO^~ (Henriksen and Kemp, NH3-oxidizing bacteria contribute to N20 or N2 1988; Schlesinger, 1997). However, it is important production in situ. Webster and Hopkins (1996) to recognize the metabolic versatility of NH3- estimated that nitrifier denitrification contributed oxidizers (i.e., nitritiers) and to consider the other 29% of the N20 produced in a dry soil and 3% in a roles they may play in nitrogen cycling. Some wet soil, but an earlier study concluded that their nitrifying bacteria produce N2 from NH4" using O2 contribution to N20 production was negligible or N02 () as oxidants. The (Robertson and Tiedje, 1987). In a review of the process has been named nitrifier denitrification topic, Wrage et al. (2001) proposed the highest to indicate that it involves autotrophic NH3- contributions from nitrifying denitrifiers are likely oxidizers with an enzyme system similar to that to occur under conditions of low carbon and of the heterotrophic denitrifying bacteria (Poth nitrogen content. The contribution of nitrifier and Focht, 1985; Wrage et al, 2001). In fact, there denitrification to anaerobic metabolism and N2 may be an evolutionary linkage between denitri- production will ultimately be limited by the low fying NH3-oxidizers and denitrifiers, as suggested availability of N02 (nitrogen dioxide) in anoxic by a high degree of similarity in their nitrite soils and sediments. reductase gene sequences (Casciotti and Ward, Autotrophic microorganisms are of interest in 2001). During nitrifier denitrification, NH3 is wastewater treatment because they can denitrify oxidized to NO;T, as in typical nitrification, then without an organic carbon supplement (Jetten et al., reduced to NO, N20, or N2 (Figure 13). Thus, the 1999; Verstraete and Philips, 1998). The microbial process couples NH3 oxidation and denitrification ecology of these artificial systems is better under- within a single organism, such as Nitrosomonas stood than natural systems at present. The anammox europaea or N. eutropha (Ritchie and Nicholas, bacterium B. anammoxidans and the nitrifier 1972; Bock et al., 1995). N. eutropha were able to coexist in a laboratory- The most recent metabolic pathway described scale reactor (Schmidt et al., 2002). Indeed, the in Nitrosomonas is anaerobic NH3 oxidation in specific anaerobic NH3 oxidation activity of which 02 is replaced by nitrogen dioxide (N02) or N. eutropha was 10 times higher than in co-culture its dimeric form, N204: than pure culture. Rather than the two groups producing N2 simultaneously, their metabolism is NH, N204 — N02 +2NO + 3H+ (18) more likely to be coupled in a nitrification-denitrifi- cation reaction that bypasses nitrite-oxidizing This pathway has been described in N. eutropha bacteria (Schmidt etal., 2002). Wastewater reactors and a few other related species (Schmidt and Bock, that favor the anammox process contain aerobic and 1997; Schmidt et al, 2002). It is superficially anaerobic NH -oxidizers, but not nitrite oxidizers similar to aerobic NH oxidation in that hydro- 3 3 such as Nitrobacter. This suggests that the nitrite xylamine is an intermediate, N02 is a product, and oxidizers cannot compete effectively with the NH3 a portion of the NO^ may be reduced further N2: oxidizers for 02, nor can anaerobic NH{~ oxidizers compete with NH oxidizers for N0 (Schmidt NO2 + 4H+ — 0.5N + 2H 0 (19) 3 2 2 2 et al., 2002). The discovery of anammox-like Schmidt and Bock (1997) demonstrated that metabolism in marine sediments (Thamdrup and anaerobic nitrifier denitrification supports growth. Dalsgaard, 2002) raises many ecological questions The anaerobic and aerobic nitrifier denitrifica- about the nature of interactions between these tion pathways differ in that NO is an end product groups of organisms in natural ecosystems. under anaerobic conditions rather than an inter- mediate compound. In addition, nitrogen dioxide- 8.08.5.5.3 Abiotic and autotrophic dependent NH3 oxidation by N. eutropha does not denitrification require ammonium monooxygenase (Schmidt et al., 2002), demonstrating that the two pathways Denitrification can be supported by electron are enzymatically different. In the absence of donors other than organic carbon such as Fe(II), NH3, N. eutropha can use H2 or simple organic Mn(II) and H2S, and they can proceed by both compounds as electron donors (Abeliovich and abiotic and biotic pathways. Abiotic reduction of Vonhak, 1992; Bock et al, 1995). In contrast to NO^ to N2 coupled to Fe(II) oxidation may occur the anammox process, which is strictly anaerobic, at low rates in the pH range 2-7 (Postma, 1990): 02 does not inhibit N02-dependent NH3 oxidation and N production can occur even under aerobic 2+ 2 10Fe + 2N03 + 14H20 conditions (Zart and Bock, 1998). However, 4 — lOFeOOH + N2 + 18H (20) Shrestha et al. (2002) observed N2 production 362 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes The reaction is quite sensitive to temperature rice paddy soils (Kamura et al., 1963; Takai et al., (S0rensen and Thorling, 1991), and is catalyzed 1963a,b), the process was thought to be either by freshly precipitated Fe(III) oxides and Cu2+ unimportant or dominated by abiotic mechanisms (Buresh and Moraghan, 1976; Postma, 1990). In (Lovley, 2000c). Similarly, biological Fe(II) contrast to the abiotic reaction, chemolithoauto- oxidation was assumed to be negligible except trophic denitrihcation coupled to NO^~ proceeds under acidic (pH < 3) conditions. It is now clear rapidly under the low temperature, circumneutral that microorganisms mediate both Fe(III) pH conditions that are typical of the Earth's reduction and Fe(II) oxidation across a broad surface (Weber et al., 2001) (Section 8.08.6.5.2). range of pH, temperature, and salinity conditions. Two pathways have been proposed that form N2 Dissimilatory Fe(III) reduction is the dominant from abiotic reactions between fixed nitrogen and form of anaerobic carbon metabolism in many mangenese, and both are thermodynamically ecosystems, and may be one of the earliest forms feasible in typical marine pore water. Reduction of microbial metabolism to evolve on Earth or of NO3" by Mn2+ (i.e., dissolved Mn(II)) was elsewhere (Lovley, 2000c). In contrast to metab- proposed by Aller (1990): olism based on oxygen, nitrogen, or carbon terminal electron acceptors, metabolisms invol- ,2+z+ 2N03 + 5Mn + 4H20 — N2 ving iron, manganese, and sulfur (Section 8.08.7) are regulated strongly by both geochemical and + 5MnQ2 + 8lf (21) biochemical phenomena. Geochemistry is funda- Because Mn reacts rapidly with 02, reaction mental to understanding the use of metals in (21) may be restricted to anoxic zones in anaerobic metabolism, and several of the most manganese-rich sediments. Although this reaction intriguing questions about iron and manganese is similar to a microbially mediated reaction bacteria concern their interactions with mineral involving Fe(II) (reaction (26)), there is no surfaces or their influence on geochemistry. evidence at present to suggest it is performed Previous reviews have addressed many of the biotically. Luther et al. (1997) proposed a mech- salient interactions between iron cycling and other anism for producing N2 from NHJ" that should be biogeochemical processes (Figure 17). favored in aerobic zones where Mn02 is rapidly Because Mn(IV) reduction generally makes a regenerated from Mn2+: small contribution to carbon metabolism, it is not considered in detail in this review. However, there + 2+ 2NH3 + 3Mn02 + 6H — N2 + 3Mn + 6H20 are many similarities between the cycles of the (22) two metals that we allude to in the text. In fact, many Fe(III)-reducing bacteria are more aptly The reaction was demonstrated in anaerobic described as metal-reducing bacteria because they marine sediments amended with freshly syn- also reduce Mn(IV) and a variety of other metals. thesized Mn02 and NH| (Luther et al., 1997). A recent and excellent review of both Fe(III) and However, it could not be detected with a sensitive 15 Mn(IV) cycling was provided by Thamdrup N-labeling technique in an unamended, manga- (2000), and Tebo et al. (1997) reviewed microbial nese-rich sediment (Thamdrup and Dalsgaard, Mn(IV) oxidation. 2000). The reaction may be favorable only with certain forms of Mn02 (Hulth et al., 1999) or high levels of NH|. 8.08.6.1 Iron and Manganese in the Environment Chemoautotrophic denitrification coupled to Fe(III) and Mn(IV) respiration influences the H2S, S°, or S2Of~ occurs in some bacteria of the genus Thiobacillus such as T. denitrificans (Hoor, cycling of many other elements that are of 1981; Section 8.08.7.9). Nitrate reduction by such a concern to environmental scientists. Fe(III)- and pathway increased with FeS additions and fol- Mn(IV)-reducers interfere with the metabolism of lowed Michaelis - Menten kinetics in a marine SO4 -reducers and methanogens by competing sediment (Garcia-Gil and Golterman, 1993). for organic carbon. Because anaerobic metab- olism is often organic carbon limited and Fe(III) reduction yields more free energy than methano- genesis, Fe(III)-reducing bacteria suppress fresh- 8.08.6 IRON AND MANGANESE water emissions of CH4, an important greenhouse gas (Section 8.08.4). Roden and Wetzel (1996) Iron and manganese oxides are the most concluded that CH4 emissions from a freshwater abundant components of Earth's surface that can wetland were reduced by 70% due to Fe(III) serve as anaerobic terminal electron acceptors in cycling. microbial metabolism, yet it was recognized only Some Fe(III)-reducing bacteria oxidize recently that microorganisms play a key role their organic pollutants (Anderson et al., 1998; cycling. Despite early reports that suggested Heider et al, 1999; Lovley, 2000a; Lovley and biological Fe(III) reduction was important in wet Anderson, 2000; Gibson and Harwood, 2002). Iron and Manganese 363

Fe(II) solids (, carbonates, ) Oxidized organics (RN02, RCl) Sorption/co-precipitation _ and inorganics (metals and rads, organics) (Se(VI), Cr(VI), As(V), Mn(IV)) C02, H20 Fe(II) Oxidized S Oz + COz (3) (sof-, S°) Solubilized species NOf + COz (4) (metal and rads, PO|~) Light + CO; ^

(2; Reduced S Organic matter (HS" FeS, FeS2) (cells) Fe(III) '1) Organic carbon, H2, organic contaminants Reduced organics (RNH2, RH+ACI") Sorption/co-precipitation and inorganics (metals and rads, organics) (Se(IV), Cr(III), Fe(III) solids As(lll), Mn(II)) (oxides, phosphates, sulfides)

Figure 17 Substrates and processes coupled to Fe reduction-oxidation. Circled numbers refer to recent reviews of the role of microbial processes in various iron transformations. (1) Lovley and Anderson (2000), Thamdrup (2000), Johnson (1998), Blake and Johnson (2000), Kiisel et al. (1999), and Peine et al. (2000). (2) Thamdrup et al., (1993), Lovely and Phillips (1994a, Schipper and J0rgensen (2002), Blake and Johnson (2000), Pronk and Johnson (1992). (3) Emerson (2000), Johnson (1998), Blake and Johnson (2000), Edwards et al. (2000b), and Roden et al. (in press). (4) Straub et al. (2001). (after Tebo and He, 1999; Roden et al., in press).

Benzene degradation is stimulated in the labora- (Raven et al, 1998; Senn and Hemond, 2002), and tory by substances that enhance Fe(III) reduction vice versa (Cummings et al, 1999; Harvey et al, (Lovley et al, 1994b, 1996b), and benzene- 2002; Zachara et al, 2001). Fe(III) oxides degrading sediments are enriched in Geobacter complex and co-precipitate phosphorus (Gunnars species (Anderson et al, 1998; Rooney-Varga et al, 2002; Bjerrum and Canfield, 2002), perhaps et al., 1999), the only genus of microorganism the biosphere's ultimate limiting nutrient (Tyrrell, known to couple Fe(III) reduction to the oxidation 1999), and P is released upon Fe(III) reduction of aromatic compounds. Because petroleum- (Smolders et al, 2001). contaminated sediments are often anaerobic, Fe(III)-reducing bacteria are attractive candidates for bioremediation. Some Fe(III)-reducing organisms, particularly 8.08.6.2 Iron and Manganese Geochemistry members of the Geobacteraceae, have the ability Iron and manganese total 5% of the continental to reduce heavy metal pollutants such as U(VI) crust, with iron contributing 98% and manga- (Holmes et al., 2002). Because reduction changes nese the remainder (Weaver and Tarney, 1984). uranium from a soluble to an insoluble form, such They are subject to rapid changes in redox state organisms may be used to remove uranium from mediated by both geochemical and biological polluted waters (Lovley, 1997; Lovley and processes. Iron atoms in near-surface Earth Phillips, 1992a). Fe(III)-reducing bacteria can environments cycle between an oxidized or ferric reduce a long list of metals including gold, silver, state, Fe(III), and a reduced or ferrous state, chromium, cobalt, , and technetium Fe(II). Manganese exists in three redox states: (Lovley, 1997). Mn(II), Mn(III), and Mn(IV). In this review, the Iron and manganese have a strong influence on abbreviation Mn(IV) is understood to represent the availability of trace metal pollutants through Mn(IV) and Mn(III). precipitation-dissolution reactions (Burdige, Oxidized iron in equilibrium with even the 1993; Cornell and Schwertmann, 1996). Trace most unstable of iron minerals has a concen- metals form surface complexes or co-precipitate tration of about 10~8 M in seawater at pH 8 with Fe(III) and Mn(IV) oxides, and they are (Stumm and Morgan, 1981), and oxidized forms released upon Fe(III) and Mn(IV) reduction of manganese are nearly insoluble at neutral pH. (Zachara et al., 2001). For example, processes Due to their poor solubility, these elements are that oxidize Fe(II) retain in sediments conserved in soils and sediments derived from 364 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes rock . The solubility of Fe(III) and Shewanella oneidensis (formerly S. putrefaciens) Mn(IV) is greatly enhanced when completed by produced and deposited an unidentified iron ligands (Lovley and Woodward, 1996; Luther mineral intracellularly (Glasauer et al., 2002). et ah, 1996; Stone, 1997). Contrary to the Intracellular deposits of magnetite were pre- assumption that most dissolved iron is Fe(II), viously known only from pore-water concentrations of Fe(III) can be and a few higher organisms (Bazylinski and comparable to Fe(II) (Ratering and Schnell, Moskowitz, 1997). 2000), presumably due to organic complexation (Luther et al., 1996; Tallefert et al., 2000). Thus, iron cycling is partially mediated by the production and degradation of organic chelators. 8.08.6.3 Microbial Reduction of Iron and Iron and manganese solubility increases dra- Manganese matically upon reduction, yet most of the Fe(II) Microorganisms that reduce extracellular and Mn(II) in sediments at any given moment is Fe(III) or Mn(III, IV) to support metabolism or not in a soluble form (Heron and Christensen, growth (i.e., dissimilatory Fe(III)- or Mn (IV)- 1995). Dissolved forms of both metals sorb reducing bacteria) can be classified into two broad strongly onto cation exchange sites. Fe(II) pre- physiological categories. The most important of cipitates as a variety of secondary minerals, the these for anaerobic carbon metabolism are the most common of which are iron monosulhde organisms that use metals as their primary (FeS), pyrite (FeS2), siderite (FeC03), vivianite II I II terminal electron acceptor for the partial or (Fe3(P04)2), and magnetite (Fe Fe 2 04), a mag- complete oxidation of organic compounds or H2, netic, mixed valance mineral. Sulfide minerals thereby conserving energy via Fe(III) or Mn(III, dominate estuarine and marine sediments (Section IV) respiration. The second group are fermenting 8.08.7), and are stable in neutral to weakly acidic bacteria that channel a small portion (<5%) of habitats in the absence of 02. Siderite formation is their organic-carbon-derived electron transport to favored in water with carbonate alkalinity > 1 mM Fe(III) or Mn(IV) reduction, thereby using metals (King, 1998), and it is the most abundant form of as nonrespiratory electron sinks (Lovley, 1987, Fe(II) in sediments dominated by Fe(III) reduction 1997). Most work on microbial metal reduction or methanogenesis (Coleman et al., 1993). These has focused on iron because it is abundant and minerals are also produced by Fe(III)-reducing interacts strongly with other elements, whether in bacteria during ferrihydrite reduction (Zachara an oxidized or reduced state. However, it is et al., 2002). In contrast to Fe(II), Mn(II) is quite recognized that metal-reducing microorganisms stable in the presence of 02 at circumneutral pH are often capable of using several alternative (<8) and therefore tends to accumulate in oxic elements as terminal electron acceptors, and environments. Mn(II) may precipitate as MnC03 likewise, a given element can be reduced by a (van Cappellen et al., 1998), but most is either wide variety of microorganisms. adsorbed, dissolved or organically complexed (Burdige, 1993). About 35% of the iron and 75% of the 8.08.6.3.1 Metabolic diversity in Fe(III)- and manganese in soils and sediments is in the form Mn(IV)-reducing organisms of free oxides (Canfield, 1997; Cornell and Schwertmann, 1996; Thamdrup, 2000). The The capacity for dissimilatory Fe(III) and remainder occurs as a minor constituent of silicate Mn(IV) reduction is widely distributed among minerals. The lattice structure of Fe(III) oxide the subdivisions of the Bacteria, including all minerals varies widely. Freshly oxidized Fe(III) subclasses of the Proteobacteria (Coates et al., precipitates rapidly as ferrihydrite (Fe(OH)3), a 2001) and some . Most of the reddish-brown, amorphous, poorly crystalline known species of Fe(III)-respiring bacteria are in mineral. Ferrihydrite is the dominant product of the delta subclass of the Proteobacteria. This Fe(II) oxidation whether it occurs by abiotic subclass includes the Geobacteraceae family, oxidation, aerobic microbial oxidation, or anaero- which is populated entirely by Fe(III)-reducing bic microbial oxidation (Straub et al., 1998). Over bacteria, including the well-studied species Geo- a period of weeks to months, amorphous ferrihy- bacter metallireducens (Lovley et al., 1993a). drite crystals undergo diagenesis to yield well- Metal reduction has also been reported in ordered, strongly crystalline, stable minerals such hyperthermophiles in the Archea (Vargas et al., as hematite(a-Fe203) and goethite (a-FeOOH) 1998). More than 40 isolates that couple anaerobic (Cornell and Schwertmann, 1996). growth to dissimilatory Fe(III) respiration have Fe(III)-reducing bacteria growing on ferrihy- been characterized (Coates et al., 2001). drite can produce extracellular fine-grained mag- The energy sources used by Fe(III)- and netite (Fe304) (Lovley et al., 1987; Fredrickson Mn(IV)-respiring organisms include H2 and et al., 1998). Recently, it was discovered that acetate, which are the same primary substrates Iron and Manganese 365 that support methanogenesis and S04~ reduction. aquifer sediments suggest that they are dominated Unlike the methanogens, some metal-respiring by the genera Geobacter or Geothrix (Anderson microorganisms also oxidize multicarbon organic etal., 1998; Rooney-Varga et al., 1999; Snoeyen- compounds such as lactate, pyruvate, long-chain bos-West et al., 2000). More ecosystems will need fatty acids, aromatic compounds, amino acids and to be surveyed before generalizations are made glucose (Kiisel etal, 1999; Lovley, 2000b). These about the dominant groups elsewhere (e.g., compounds can be completely oxidized to C02, or Rossello-Mora et al., 1995). they can be fermented to acetate. Some fermenta- tive organisms cannot grow solely on Fe(III), but reduce small quantities of Fe(III) as a nonrespira- tory electron sink (Lovley, 1987, 2000b). 8.08.6.3.2 Physical mechanisms for accessing oxides Complete oxidation of organic carbon to C02 is found in three of the four genera in the A fundamental difference between Fe(III) and Geobacteraceae family (Lovley, 2000c), and Mn(IV) reduction and the other major pathways of phylogenetically distinct genera such as Geovibrio anaerobic metabolism is that the oxidants are and Deferribacter (Greene et al., 1997). The sparingly soluble solids. In order to access the gamma and epsilon subclasses of the Proteobac- terminal electron acceptor a microorganism must teria tend to oxidize a more limited range of either: (i) make physical contact with an oxide organic acids, and often carbon oxidation is surface, (ii) reduce an intermediate compound that incomplete (Lovley, 2000b). A well-studied can shuttle electrons from the cell to a distant member of this group is Shewanella oneidensis oxide surface, or (iii) use the metal in a dissolved (formerly S. putrefaciens), one of the first or chelated form. All three adaptations are described nonfermentative Fe(III)-reducers observed among Fe(III)- and Mn(IV)-respiring (Myers and Nealson, 1988; Nealson and Myers, organisms. Metal-respiring organisms do not 1992), and notable because it is also a facultative necessarily require direct physical contact with anaerobe (Kieft et al., 1999). The ability to use H2 oxides as some early studies suggested (Lovley for an energy source is common among incom- and Phillips, 1986; Tugel et al., 1986), but such pletely oxidizing Fe(III)-reducers, and hyperther- contact may be common. Evidence of this is the mophilic Fe(III)-reducers in the Archea and observation that Fe(III) reductase activity tends to Bacteria (Slobodkin et al., 2001; Vargas et al., be localized in the outer membrane of both G. 1998). Relatively few Fe(III)-reducing bacteria sulfurreducens and S. oneidensis (Beliaev and outside these groups are known to oxidize H2 Saffarini, 1998; Gaspard et al, 1998; C. R. Myers (Coates et al., 1999). However, some species of and J. M. Myers, 1992). Using atomic force Geobacter can use both H2 and acetate (Caccavo microscopy, Lower et al. (2001) measured the ef a/., 1994, 1996). adhesion strength of S. oneidensis cells to mineral Most Fe(III)-reducing organisms have the surfaces and found that it increased rapidly upon capacity to reduce Mn(IV) and at least one other exposure to goethite under anaerobic conditions. common oxidant such as 02, NO^, SO4 , There was some evidence for a mobile iron elemental sulfur (S°), or humic substances reductase in the outer that (Thamdrup, 2000; Senko and Stolz, 2001). In facilitated electron transport from the cell to the contrast to the small penalty in free energy that oxide surface. Grantham et al. (1997) reported occurs when denitrifying bacteria switch from 02 that sites where S. oneidensis was observed to NOJ-respiration, growth on Fe(III) by the adhering to Fe(III) oxide surfaces corresponded facultative anaerobe S. oneidensis is far inferior to to pits caused by Fe(III) dissolution. The shape of that on 02 (Kostka et al., 2002a). Humic the pits suggested that the organisms were more substances may prove to be the most common mobile under anaerobic conditions than aerobic alternative electron acceptor for metal-reducing conditions, an adaptation that may enhance their bacteria, particularly in freshwater environments access to fresh Fe(III) oxide surfaces. A similar where S° is not abundant (Lovley et al., 1996a, adaptation was observed in G. metallireducens 1998). There are several SO4"-reducing bacteria which expressed Hagella and was chemotaxic that can also reduce Fe(III) and U(VI), although toward Fe(II) and Mn(II) when grown on they generally cannot support growth on transition insoluble oxides, but not when grown on soluble metals alone (Lovley et al., 1993b). In fact, Fe(III) or Mn(IV) (Childers et al, 2002). This desulfuricans will simultaneously adaptation had gone unnoticed because most reduce SO4" and Fe(III), or SO4" and U(VI), if H2 culture work had previously been based on is available (Lovley and Phillips, 1992b; Coleman complexed, soluble Fe(III) media, highlighting gfaZ., 1993). the importance of culturing metal-reducing organ- Relatively little is known about the abundance isms on substrates that are common in situ such as of Fe(III)-reducing species in situ. Attempts to ferrihydrite or goethite. More species must be quantify dissimilatory Fe(III)-reducing bacteria in tested to determine if chemotaxis is unique to 366 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes Fe(III)-reducing organisms that do not produce substances in an oxidized form. Reduced humics soluble electron shuttling compounds. and (i.e., reduced quinones) are rapidly oxidized by Fe(III) and Mn(IV), making them available again for microbial respiration. Some humic substances may be more efficient than microbial cells at accessing physical 8.08.6.4 Factors that Regulate Fe(III) Reduction locations on metal oxides because they are Because Fe(III) respiration is the most recent relatively small (Zachara et al., 1998). Finally, major pathway of anaerobic metabolism to be because enzymatically reduced humics may pass investigated, there has been relatively little work electrons to Fe(III), Mn(IV), U(VI), or other on the processes that regulate it in situ. Rather, oxidants (Stone, 1987), an organism can indirectly most advances have concerned the mechanisms by access several potential "secondary" electron which microorganisms access Fe(III) oxides and acceptors simultaneously rather than just one interact with the physicochemical processes that (Figure 18). The ability to shuttle electrons govern Fe(III) oxide dissolution. between microorganisms and inorganic oxidants explains why even low concentrations of AQDS (5 jjumol kg-1) and humics can stimulate Fe(III) reduction in sediments (Nevin and Lovley, 2000). 8.08.6.4.1 Electron shuttling compounds An actual contribution of humic substances to Humic substances are effective electron shut- metal oxide reduction in natural systems has not tling compounds that are ubiquitous in soils and been demonstrated, and there are processes such sediments (Stevenson, 1994). The redox-active as adsorption or decomposition that could limit moieties in these compounds are primarily their effectiveness. Kostka et al. (2002a) observed quinones and related aromatic reductants (Lovley that AQDS additions elicited a larger increase in and Blunt-Harris, 1999; Scott et al., 1998; Stone Fe(III) reduction by S. oneidensis growing on et al., 1994). Because purified humic acids are ferrihydrite than smectite clay minerals. This expensive and it is difficult to evaluate their redox suggests that the influence of humic substances state, the synthetic quinone 2,6-anthraquinone may depend on soil or sediment mineralogy. disulphonate (AQDS) has been used as a model Nevertheless, there is ample evidence to suggest compound for studying the potential for humic that a portion of the anaerobic metabolism that acids to serve as terminal electron acceptors was previously attributed to direct enzymatic (Lovley et al., 1996a). AQDS may not be an Fe(III) and Mn(IV) reduction was actually none- appropriate substitute for native humic acids in nzymatic reduction by microbially reduced humic studies of reaction kinetics (Nurmi and Tratnyek, substances. 2002), but it has proved to be useful for isolating Variability in the effectiveness of humic sub- humic-reducing microorganisms from a variety of stances as electron shuttling compounds is ecosystems including lakes, wetlands and estu- expected due to differences in their chemical aries (Coates et al., 1998). It appears that most, if structure. For example, the electron-accepting not all, Fe(III)-reducing bacteria can obtain capacity of humic substances from three distinct energy by the reduction of AQDS or humic sub- sources varied nearly 700-fold in the following stances (Lovley et al., 2000). The strains tested order: soils > sediments > dissolved river-borne to date include members of the Geobacteraceae (Scott et al., 1998; Figure 19). Microorganisms family and hyperthermophiles. The metabolism are a source of quinone moities and other electron is present in fermentative and nonfermen- shuttling compounds. For example, S. oneidensis tative organisms, and may be particularly widespread among thermophilic organisms (Lovley, 2000c). There are several reasons why the ability to reduce humic substances could be an advantage to Fe(II) metal-respiring microorganisms. Higher rates of reduction should be possible with humic acids and other soluble oxidants than with solid-phase oxidants that require the organism to continually establish physical contact in order to access fresh Fe(III) oxide surfaces. Humic substances are ubiquitous and accumulate to levels that could make them significant alternative electron acceptors in some ecosystems. As with other anaerobic electron Figure 18 The variety of electron shuttles that acceptors, the capacity to support respiration is promote Fe(III) reduction, (after Nevin and Lovley, enhanced by processes that regenerate humic 2000). Iron and Manganese 367

Q 8001 with S. oneidensis, the stability of soluble ligand- | 700. Fe(III) complexes was inversely related to Fe(III) reduction rates (Haas and DiChristina, 2002). Of fr 600' the chelators tested, the highest Fe(III) reduction I 500. rates were produced by citrate, which is a weakly « 400' complexing ligand similar to most naturally V occurring organic chelators. «T 300' Organic ligands also form complexes with | 200, Fe(II) (Luther et al., 1996) and may enhance 1 loo- Fe(III) dissolution by removing the Fe(II) that 53 o. E*l1^1 FA FA [IA HA HA HA coats and passively blocks access to fresh Fe(III) a. oxides surfaces. Urrutia et al. (1999) considered Water Sediment Solids this mechanism and concluded that it was more Figure 19 Ecosystem-level differences in the ability important than Fe(III) complexation. Bacterial of humic substances to accept electrons. Samples were cells adsorbed to mineral surfaces also inhibit taken from the water column of freshwater lakes and Fe(III) reduction rates (Urrutia et al., 1999). rivers, aquatic sediments and soils (FA = fulvic acid, HA = humic acid) (after Scott et al, 1998).

8.08.6.4.3 Mineral reactivity excretes an unidentified quinone-based compound Metal oxides in soils and sediments are that can reduce ferrihydrite and Mn02 (Newman mixtures of poorly crystalline, strongly crystalline and Kolter, 2000), and G. sulfurreducens excretes and silicate-bound minerals. Differences in the a c-type with similar capability intrinsic reactivity of these broad categories is (Seeliger et al., 1998). These studies raise the sufficient to explain much of the variation in Fe(II) prospect that some organisms can synthesize their and Mn(II) reduction rates observed in nature, own electron shuttle compounds if the supply of which spans several orders of magnitude. Mix- oxidants is limiting. tures of Fe(III) oxides are typically separated by exposing them to a sequence of chemical extrac- tions involving increasingly strong solutions of 8.08.6.4.2 Fe(III) chelators acids, ligands, and reductants. For example, the poorly crystalline fraction can be extracted with a There are a variety of humic and nonhumic combination of aerobic and anaerobic organic chelating agents that enhance the dissol- solutions (Thamdrup and Canfield, 1996) or a ution and solubility of metals (Stone, 1997; Luther dilute solution of HC1 (0.5 M), while dithionate- et al., 1992). Some forms of chelated Fe(III) are citrate-bicarbonate is used to extract the strongly more rapidly reduced by bacteria than insoluble crystalline fraction. A limitation of such tech- Fe(III) oxides (Lovley, 1991). Amending aquifer niques is that the fractions are operationally sediments with the synthetic chelator nitrilotria- defined and not mineral specific. Efforts have cetic acid (NTA) increased dissolved Fe(III) and been made to calibrate wet chemical extraction stimulated the enzymatic reduction of Fe(III)- protocols against minerals of known composition oxides (Lovley and Woodward, 1996). The (Canfield, 1989; Haese et al, 1997; Kostka and increase in Fe(III) reduction was attributed to Luther, 1994). Nevertheless, no extraction tech- enhanced availability of dissolved Fe(III) rather nique is entirely discriminating for a specific than a stimulation of growth caused by trace metal phase or mineral (Zachara et al., 2002). elements or metabolizable organic carbon. How- Recognizing the wide range of reactivities rep- ever, the effect could also be explained by NTA resented in a single operationally defined Fe(III) complexation of Fe(II), which would prevent oxide pool, Postma (1993) proposed characteriz- Fe(II)-poisoning of the mineral surface (see next ing reactivity with a single continuous extraction. paragraph). Other complexing agents such as The "reactivity continuum" method allows hydroxamate have no effect on Fe(III) dissolution straightforward comparisons of the bulk reactivity rates (Holmen et al., 1999). As with humic acids, of pure and mixed iron (hydr)oxide pools there is very little evidence that chelators (Figure 20). substantially influence microbial iron metabolism The susceptibility of metal oxides to reduction in situ, and there are reasons that model chelators and dissolution depends on mineralogy, crystal- such as NTA are not good surrogates for those in linity, surface area, the effectiveness of reducing natural systems (Straub et al., 2001). The stability and chelating agents, and microbial activity. of ligand-Fe(III) complexes influence Fe(III) Early culture studies with Fe(III)-respiring reduction rates because the cell must be able to bacteria demonstrated that Fe(III) reduction out compete the chelator for Fe(III). In a study rates vary with mineral form or crystallinity 368 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes

100 200 300 400 500 600 (a) Surface area (m2 g~')

100 150 200 300 Time (xlO3 s)

Figure 20 Application of the reactivity continuum method to Fe(III) oxide pools in oxidized aquifer sediments. The ratio mlm0 is the fraction of undissolved Fe-oxide. The slope of the curve decreases with time because reactive minerals are increasingly depleted leaving behind relatively more recalcitrant forms 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 (source Postma, 1993). (b) Fe(II) (mM) Figure 21 Relationship of Fe(III)-reducing bacteria activity and growth to oxide surface area, (a) Percent (Lovley and Phillips, 1986). Minerals such as Fe(III) reduced as a function of oxide surface area. Surface ferrihydrite and lepidocrocite (-y-FeOOH) are area corresponded to different mineral types and included generally reduced more rapidly than relatively hematite, goethite, and ferrihydrite. (b) The density of Shewanella oneidensis cells as a function of the amount of stable minerals such as goethite and hematite structural Fe(III) reduction to Fe(II) in smectite clay, a (Postma, 1993). Amorphous manganese oxides strongly crystalline, high-surface-area Fe mineral. Differ- such as vernadite are more easily reduced than ences in Fe(II) content reflect different amounts of clay strongly crystalline forms such as pyrolusite, but particles inoculated into a minimal basal media (after the overall influence of crystallinity on reduction Roden and Zachara, 1996 and Kostka et al., 2002a, kinetics appears to be weaker for manganese and respectively). iron oxides (Burdige et al., 1992). It has now become evident that the primary (Stucki et al., 1996), and may therefore support reason for mineral-related differences in reactivity a substantial portion of the planet's anaerobic is not thermodynamic stability, but the fact that microbial carbon metabolism. amorphous minerals have a far higher surface area The reactivity of Fe(III) and Mn(IV) minerals is than crystalline minerals (Figure 21). Kostka et al. often highest immediately after they precipitate (2002a) reported that Fe(III)-reducers grew nearly following Fe(II) or Mn(IV) oxidation. Because as well on smectite clay as on ferrihydrite, even metal reduction is usually limited by electron though smectite clay is considered a crystalline acceptor availability (or co-limited with . Although smectite clay is more highly carbon), ecosystem processes that favor Fe(II) and crystalline than ferrihydrite, the two minerals 2 _1 Mn(II) oxidation also promote Fe(III) and Mn(IV) have a comparable surface area (~700m g ) reduction. Regeneration by O2 occurs slowly in (Schwertmann and Cornell, 1991). As a result, rivers, lakes, and oceans because 02 diffusion in Fe(III)-reducing bacteria were able to remove water is exceedingly slow and sediments have a 20-50% of the total clay-bound iron and thereby high biological oxygen demand. Oxidation is alter the physical and chemical characteristics of much faster when the sediments are mixed by these ubiquitous minerals (Kostka et al., 1999a,c). currents or bioturbation (Aller, 1990). Periodic Although Fe(III) reduction rates are reduced by drawdown of the water table in intertidal sedi- one to two orders of magnitude in the presence of ments and wetlands rapidly introduces 02 and minerals with low- versus high surface area, there accelerates metal oxidation. Wetland plants is ample evidence demonstrating that dissimila- enhance metal oxidation by introducing 02 tory Fe(III)-reducing bacteria can gain energy for directly in the soil from porous root systems growth using goethite, hematite, and magnetite (Kostka and Luther, 1995), a process known as (Kostka and Nealson, 1995; Roden and Zachara, radial oxygen loss, and by removing soil water by 1996), in some cases consuming up to 30% of transpiration (Dacey and Howes, 1984). Metal the oxide-bound Fe(III). Fe(III)-rich clays and oxides are also regenerated anaerobically through other highly crystalline forms account for much a variety of enzymatic and nonenzymatic mech- of the Fe(III) mass in soils and sediments anisms coupled to redox cycles of sulfur, iron, and Iron and Manganese 369 manganese. Because all of these processes can contributes to nonenzymatic Fe(III) reduction in occur in wetland soils, a larger proportion of the water columns or sediments. total iron pool may be reactive in wetlands than The relative contributions of enzymatic and non- nonwetland ecosystems (Weiss, 2002). Thus, enzymatic pathways to Fe(III) reduction in marine wetlands may be "hot spots" of iron cycling. environments is difficult to measure and highly variable. Because nonenzymatic reduction is rapid enough to compete with microorganisms for Fe(III)-oxide substrates (Thamdrup, 2000), non- 8.08.6.4.4 Abiotic versus biotic reduction enzymatic reduction should be favored over enzy- Iron and manganese differ from other common matic reduction in environments where sulfides are terminal electron acceptors such as 02, NO3", abundant and reactive Fe(III) oxides are scarce. In SO4 , and HCO^ in that they are subject to rapid such sediments, biological reduction may be nonenzymatic reduction. For example, Fe(II) restricted to microsites where sulfide levels are reduces Mn(IV) according the generalized reac- low and poorly-crystalline Fe(III) is abundant tion (Lovley and Phillips, 1988; Postma, 1985): (Canfield, 1989). However, H2S does not accumu- late in the porewater of many marine environments 2Fe(II) + Mn(IV) — 2Fe(III) + Mn(II) (23) (i.e. they are not sulfidic) because it reacts with metal This reaction makes manganese the ultimate oxides. In sediments where sulfide is produced but electron sink for a portion of the organic carbon does not accumulate, enzymatic Fe(III) reduction consumed by dissimilatory Fe(III)-reducing bac- rates are often substantial. Stoichiometric consider- teria, and it reduces diffusive losses of iron from ations suggest that enzymatic reduction should be a sediments. Nonenzymatic Mn(IV) reduction by large part of the overall Fe(III) reduction in such Fe(II) can be significant in high-Mn(IV) sediments sediments (Thamdrup, personal communication). (Aller, 1990), but its global significance is For example, the oxidation of one mole of organic ultimately limited by the low content of manganese carbon to C02 can support the reduction of either 4 moles of Fe(III) to Fe(II) or 0.5 moles of SO4" to in the Earth's crust compared to iron. 2_ Nonenzymatic reduction by sulfides is widely S . Assuming that SO|~ reduction produces H2S, considered to be the primary mechanism for Fe(III) the H2S could oxidize as little as 1/3 moles of Fe(III) and Mn(IV) reduction in systems where SO4 is (i.e. 3H2S + 2FeOOH — 2FeS + S). In this case, abundant and sulfate reduction dominates anaero- the relative contributions of enzymatic and non- bic metabolism (Aller and Rude, 1988; Burdige enzymatic processes to Fe(III) reduction would be and Nealson, 1986; Jacobson, 1994; Kostka and comparable even if Fe(III) reduction contributed Luther, 1995; Postma, 1985; Pyzik and Sommer, just 10% of total organic carbon oxidation. It is 1981). Yet, Fe(III)-reducing bacteria are abundant necessary to quantify the proportion of total Fe(III) in marine soils and sediments (Lowe et al, 2000; reduction contributed by Fe(III)-reducing bacteria Kostka et al, 2002c), and several workers have in order to understand the ecology of these noted that Fe(II) continues to accumulate when ubiquitous organisms. However, from the perspec- tive of overall microbial metabolism, it should be H2S production is blocked by molybdate, appar- ently because of enzymatic Fe(III) reduction noted that Fe(III) reduction is ultimately a result of (Canfield, 1989; Canfield et al, 1993b; Mines microbial respiration regardless of the mechanism. et al, 1997; Jacobson, 1994; Joye et al, 1996; That is, Fe(III) serves indirectly as a terminal Kostka et al, 2002c; Lovley and Phillips, 1987; electron acceptor for SO|~-reducing bacteria when S0rensen, 1982). Fe(III) undergoes non-enzymatic reduction by H2S. Organic compounds have the potential to abiotically reduce Fe(III) and Mn(IV) (Luther et al, 1992; Stone, 1987). Phenols and a variety of 8.08.6.4.5 Separating enzymatic and other aromatic compounds reduce Fe(III) rapidly nonenzymatic Fe(III) reduction at acidic pH, but slowly at circum neutral pH (LaKind and Stone, 1989). Humics can reduce A common approach to estimating the contri- Fe(III) effectively at circumneutral pH and they bution of microorganisms to Fe(III) and Mn(IV) are abundant in soils and sediments. Because reduction based on carbon mass balance. Total humics and other organic compounds often serve anaerobic microbial respiration is estimated by as electron shuttles between metal-reducing bac- measuring the carbon mineralization rate (i.e., teria and metal oxides (Lovley et al., 1996a), it J]C02 + CH4 production). The carbon that may be difficult to separate microbial and was respired by SO4 -reducing bacteria and nonmicrobial sources of electrons. Finally, methanogens is subtracted from the total, and aerobic photoreduction of Fe(III) has been the difference is assumed to be the contribution of observed in freshwater and marine environments Fe(III)- and Mn(IV)-reducing bacteria (Canfield (Barbeau et al, 2001; Emmenegger et al, 2001), et al, 1993b; Thamdrup, 2000). Typically, the but it is unknown to what degree this process contribution of aerobic metabolism to carbon 370 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes mineralization is excluded by design, denitrifica- 8.08.6.4.6 Fe(IH) reduction in ecosystems tion is assumed to be negligible because of limited NO^~ availability, and it is assumed that non- Field studies of enzymatic Fe(III) reduction are methanogenic fermentation does not completely scarce, but they generally indicate that the availability of reactive Fe(III) and organic carbon oxidize organic carbon to C02. The approach is attractive in marine systems because methanogen- govern rates. Thamdrup (2000) reported a positive esis is negligible, which simplifies the calculation, relationship between the content of poorly crystal- line Fe(III) in marine sediments and the fraction of and SO4 reduction can be measured accurately 35 carbon oxidation mediated by Fe(III)-reducing using a S radioisotope technique (J0rgensen, bacteria (Figure 22). The range of values indicates 1978a,b,c). Sulfate reduction measurements in salt that Fe(III) reduction can account for up to 90% of marsh soils are more problematic because severed the carbon oxidation in some sediments. Roden roots can introduce fermentable organic carbon and Wetzel (2002) found a similar relationship in compounds, and spatial variability introduces a freshwater marsh, also with a contribution from error into the difference calculation. Fe(III) reduction of between 20% to 90% to the A second method for separating enzymatic and overall anaerobic carbon metabolism. It is note- nonenzymatic Fe(III) reduction by H S is to block 2 worthy that Fe(III) reduction was never Fe(III)- SO|" reduction with molybdate (M0O4"). The saturated in either study, suggesting that Fe(III) technique has been used effectively to demonstrate availability may limit Fe(III) reduction within the the importance of enzymatic reduction in marine range of labile Fe(III) concentrations that are and freshwater sediments (Section 8.08.6.4.4). As generally found in soils and aquatic sediments. By with all inhibitor techniques, there is the possibility analogy to denitrification, SO4 reduction and that molybdate additions directly or indirectly methanogenesis, organic carbon availability is affect processes other than SO4" reduction. For expected to also limit Fe(III) reduction. Roden and example, it could overestimate biotic Fe(III) Wetzel (2002) reported a strong linear relationship reduction if the enzymatic process was stimulated (r2 = 0.99) between microbial respiration by a cessation of competition with H S for Fe(III) 2 (a measure of carbon availability) and initial substrates, or underestimate if SO4" reduction was Fe(III) reduction rate, and proposed that overall not completely blocked. Despite these potential Fe(III) reduction was regulated by both carbon limitations, the molybdate method produces pat- and Fe(III) availability according to the equation: terns that are consistent with other types of geochemical data, and it is therefore widely used. Rpe(ni) = aM Fe(III) (24) Perhaps the most elegant method for separating oc ri enzymatic and nonenzymatic Fe(III) reduction where ^Fe(iii) and R are rates of Fe(III) reduction would exploit a difference in natural Fe-isotope oc and organic carbon decomposition, respectively. fractionation. However, such a method has proved Fe(III) is the concentration of reactive Fe(III), to be elusive. Experiments with the Fe(III)- reac a is the stoichiometric ratio of Fe(III) atoms reducing bacterium Shewanella algae yielded a reduced per carbon atoms oxidized (4:1 in this 1.3%c fractionation between the ferrihydrite case), and b is a rate constant. Thus, for any given substrate and soluble Fe(II), suggesting that initial amount of reactive Fe(III), organic carbon variations in the natural abundance 8 Fe could be interpreted biologically (Beard et ai., 1999). Unfortunately, even larger iron fractionations (up 100 to 3.6%c) are possible by nonbiological mechan- isms (Anbar et ai, 2000). The 160/180 ratio of 80 siderite (FeCO^) was similar in the presence and • * absence of Fe(III)-reducing bacteria, suggesting that oxygen isotopes will not be a useful signature 60 of enzymatic Fe(III) reduction (2001). However, the isotopic composition of biogenic iron minerals 40 could prove to be useful as paleothermometers. 18 Temperature-dependent fractionation of 5 0 in 20 biogenic siderite occurred in cultures of Fe(III)- •••• reducing bacteria (Zhang et al., 1997; Zhang et al., uO 2001), and in the intracellular magnetite produced by magnetotactic bacteria (Mandernack et al., 0 10 20 30 40 50 60 Poorly crystalline Fe(III) (mmol cm-3) 1999). Radiolabeling the pool to more precisely measure Fe(III) reduction rates proved Figure 22 The fraction of carbon metabolism due to unsuccessful because of rapid isotope exchange Fe(III) reduction in marine sediments as a function of between the Fe(III) and Fe(II) pools (Roden and the poorly crystalline Fe(III) content (after Thamdrup, Lovley, 1993). 2000). Iron and Manganese 371 availability regulates Fe(III) reduction rates until strain SG-1 has also been well studied. This the reactive Fe(III) pool has been exhausted. species and related strains have been shown to oxidize Mn(II) even while in a spore stage (Francis and Tebo, 2002). All of the known organisms that perform Mn(II) oxidation are 8.08.6.5 Microbial Oxidation of Iron heterotrophic, and the process has not yet been and Manganese linked to chemolithoautotrophy. Further details on Fe(II) and Mn(II) oxidation regenerate high- Mn(II) oxidation are provided by Tebo et al. quality (i.e., poorly crystalline) substrates for (1997) and Emerson (2000). Fe(III) and Mn(IV)-reducing microorganisms. In contrast to other terminal electron acceptors, these elements rapidly precipitate upon oxidation 8.08.6.5.2 Anaerobic Fe(II) oxidation and settle so that they are efficiently recycled in the Anaerobic Fe(II) oxidation is one of the most ecosystem. Oxidation occurs under both aerobic recently recognized categories of anaerobic and anaerobic conditions, and it can be biologically metabolism. The first report of such metabolism mediated or autocatalytic in either case. Mechan- described anaerobic, phototrophic organisms that isms for chemical oxidation of Fe(II) under required only Fe(II), C02, and light for growth anaerobic conditions include reaction with (Widdel et al., 1993): Mn(IV) (Postma, 1985) and NO% (Hansen et al., 2+ 1994; Weber et al., 2001). Organisms that oxidize 4Fe + C02+llH20 Fe(II) and Mn(II) to support growth must compete H with chemical oxidation for substrates. — 4Fe(OH)3 + (CH20) + 8H (25) Since then, several such bacteria have been isolated from freshwater and marine environments 8.08.6.5.1 Energetics of Fe(II) and Mn(II) (Ehrenreich and Widdel, 1994; Heising and oxidation Schink, 1998; Heising et al., 1999; Straub et al., 1999). One implication of this metabolism is that Microbial growth based on Fe(II) oxidation microorganisms may have contributed to the proceeds readily at acidic pH (<4) because banded iron formations, which resulted from Fe2+ (i.e., aqueous Fe(II)) is stable (Patrick and widespread Fe(III)-oxide deposition in oceans at Henderson, 1981). In contrast, biological Fe(II) a time when there was very little atmospheric 02 oxidation at circumneutral pH has been (Kump, 1993). Little is known about the physi- assumed to be unimportant because autocataly- ology or ecology of these organisms (Straub et al., tic oxidation is rapid, and the reaction produces 2001), but most strains can couple Fe(II) oxidation relatively little free energy (Straub et al., 2001). to NO3" reduction. However, neutralophilic Fe(II)-oxidizing bac- Even more recent was the first report of teria growing on Fe(II) may actually be able anaerobic Fe(II) oxidation coupled to NO^ to generate more free energy than acidophilic reduction (Straub et al., 1996; Hafenbradl et al., bacteria when one considers the forms of Fe(II) 1996): and Fe(III) in natural ecosystems. At a pH of 6 or 7, Fe(II) will often be in the form of FeC03, 10FeCO3 + 2NOr + 24H20 — 10Fe(OH)3 and Fe(III) will be in the form of an insoluble + N2 + 10HCO3~ + 8H+ (26) amorphous hydroxide, such as Fe(OH)3 or FeOOH, which removes the product from the Although NO^T reduction to NHJ" is thermodyna- solution (Emerson, 2000). Under these con- mically feasible, all known strains produce ditions, Fe(II) oxidation coupled to 02 respir- primarily N2 (Straub et al., 2001), making this a ation can generate substantial free energy. novel metabolic pathway for N2 production Because Mn(II) is stable at pH <8, significant (Section 8.08.5.5.3). Cultures from a variety of contributions from microorganisms to Mn(II) environments grow solely on aqueous or solid- oxidation has been relatively easy to demonstrate, phase Fe(II) and C02 (i.e., they are chemolithoau- and a number of Mn(II) oxidizing bacteria have totrophs), while others require Fe(II), C02 and an been isolated. Microbial oxidation is considered organic co-substrate such as acetate (i.e., they are the primary mechanism of Mn(II) oxidation in putative ). Organisms capable of circumneutral freshwater (Ghiorse, 1984; Nealson anaerobic Fe(II) oxidation accounted for up to et al., 1988). The sheathed bacterium Leptothrix 58% of the total cultivatable denitrifying commu- discophora is perhaps the most-studied species of nity in the profundal sediments of a lake (Hauck Mn(II)-oxidizing bacteria, and rate laws have et al., 2001), but not more than 0.8% of denitrifiers been developed to describe Mn(II) oxidation as a in other aquatic sediments (Straub and Buchholz- function of pH, temperature, dissolved 02, and Cu Cleven, 1998). Freshwater organisms can couple concentration (Zhang et al., 2002). sp. NO^ reduction to the oxidation of microbially 372 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes reduced goethite and other solid-phase Fe(II) where it would have little influence on pyrite (Weber et ai, 2001), and enrichment cultures dissolution (Schrenk et ah, 1998). In fact, from marine sediments oxidize FeS, but not FeS2 A. ferriooxidans may serve a beneficial role (Schippers and J0rgensen, 2002). The contribution by reducing aqueous loads of Fe(II) and other of these organisms to N2 production and Fe(II) metals. As these studies illustrate, there is much oxidation activity in ecosystems has not been insight to be gained from investigating the evaluated. It has been suggested that their role ecology of microorganisms in situ. in Fe(II) oxidation is ultimately limited by NOJ The only neutrophilic Fe(II)-oxidizer that had availability, which is typically low due to NO^ been cultured in the laboratory previous to the assimilation by plants and microorganisms. 1990s was Gallionella ferruginea, a microaerobe However, many of these organisms are facultative that forms a helical stalk (Emerson, 2000). anaerobes and can use 02 as an alternative G. ferruginea is a member of the beta-Proteobac- electron acceptor for heterotropic growth. The teria, and is capable of chemolithoautotrophic or capacity to switch between 02 and NO^ as mixotrophic growth (Hallbeck and Pederson, terminal electron acceptors may be a common 1991; Hallbeck et ah, 1993). A number of novel trait among chemoautotrophic bacteria that link strains of chemolithoautotrophic Fe(II)-oxidizing Fe(II) oxidation to N2 production, just as it is bacteria have been isolated from groundwater among strictly heterotrophic denitrifying bacteria. (Emerson and Moyer, 1997) and hydrothermal vents (Emerson and Moyer, 2002). These strains are unicellular and do not form extracellular stalks. 8.08.6.5.3 Aerobic Fe(II) oxidation They are members of the gamma-P roteobacteria subclass in the family Xanthomonaceae. Another Iron oxidation occurs at the interface of aerobic recent isolate of lithotrophic Fe(II)-oxidizing and anaerobic environments according to the bacteria, strain TW-2, was recovered from a following generalized reactions: freshwater wetland and determined to be in the 2+ + 3+ beta-Proteobacteria (Sobolev and Roden, 2003). Fe i02+H ^Fe ^HzO (27) It is often assumed that Fe(II) oxidation is a ,3+J_f nonenzymatic process at the circumneutral pH Fe " + 3H20 — Fe(OH)3(s) + 3IT (28) conditions that prevail in most ecosystems because The relative contributions of enzymatic versus autocatalytic oxidation is quite rapid. This may be nonenzymatic oxidation to overall Fe(II) oxi- the case when 02 levels are high due to physical dation rates is influenced by a variety of factors mixing, bioturbation, or rapid changes in water such as pH, concentrations of 02 and Fe(II), table depth. But when the position of the oxic- and the rate of Fe(II) delivery (Neubauer et ah, anoxic interface is stable, opposing diffusion 2002). At pH <4, the nonenzymatic oxidation gradients of Fe(II) and 02 intersect to form of Fe(II) by 02 is very slow, and microbial microaerobic zones ([02] <0.5%), the ideal activity can increase oxidation rates by a factor niche for organisms that require both reduced >10 (Singer and Stumm, 1970). The practical inorganic elements and 02 for respiration. The implications of this observation for the abate- half-life of nonenzymatic Fe2+ (i.e., dissolved ment of has motivated a Fe(II)) oxidation at low 02 concentration is up to great deal of research on chemolithotrophic, 300-fold longer than at high 02 concentration acidophilic prokaryotes such as Acidothiobacil- (Roden et ah, in press). A long Fe2+ half-life lus ferrooxidans (formerly Thiobacillus ferroox- should favor microbial oxidation. Indeed, chemo- idans) and Lepto spirillum ferrooxidans lithoautotrophic bacteria competed successfully (Nordstrom and Southam, 1997). The influence with nonenzymatic processes for 02 and Fe(II) of such organisms on the pH of mine drainage when grown in pure culture at circumneutral pH may vary, depending on the niche they occupy (Emerson and Revsbech, 1994). in an acid mine waste . L. ferrooxidans Pure culture studies have alluded to some of the tolerates extremely low pH (<1.0) conditions factors that may regulate enzymatic Fe(II) oxi- typical of solutions in contact with pyrite ores, dation in complex natural environments such as and it may be responsible for initiating a series sediments, soils, and the rhizosphere. Neubauer of reactions that ultimately enhance pyrite et ah (2002), used microcosms fed with envir- dissolution and generates acidity (Schrenk onmentally relevant concentrations of 02 and et ah, 1998). Archea in the order Thermoplas- Fe(II) to investigate the metabolism of an Fe(II)- males are also abundant at such sites and oxidizing strain isolated from the wetland rhizo- presumably contribute to Fe(II) oxidation sphere. They found that both biotic and abiotic (Edwards et ah, 2000b). Contrary to its Fe(II) oxidation increased linearly (r2 > 0.90) presumed role of enhancing pyrite dissolution, with the rate of Fe(II) addition (Figure 23). Since it has been shown that A. ferriooxidans actually the experimental Fe(II) addition rate approxi- occurs in somewhat less acidic environments mated in situ Fe(III)-diffusion rates in freshwater Iron and Manganese 373 Fe(II) oxidation (Neubauer et al., 2002). Indirect -"-0.5% 02 lmMFe(TI) evidence that aerobic, chemolithotrophic Fe(II)- -»-0.5%O2 10mMFe(E) oxidizing bacteria contribute to Fe(III) deposition -»-3.0%O2 lmMFe(II) in circumneutral environments is the observation —*-3.0%O2 lOmMFelll) that they are ubiquitous and can account for 1 % or more of the total microbial population (Weiss ef a/., 2003).

8.08.6.6 Iron Cycling Nealson (1983) articulated the first detailed description of iron cycling in which microbial 100 200 300 400 500 600 _1 metabolism was a central theme. Research since Fe(H) addition rate (uM h ) then has confirmed the more speculative elements of a microbial , and contributed new Figure 23 The influence of a chemolithoautotrophic bacterium on Fe(II) oxidation rates as a function of insights on the processes that link Fe(III) reduction Fe(II) addition rate (x-axis) as determined in stirred and Fe(II) oxidation in a "ferrous wheel," particu- microcosms. The effects varied depending on the 02 larly in wetland soils. Microbial iron cycling concentration and the total Fe(II) oxide content in the promotes organic carbon oxidation via Fe(III) microcosm. The bacterium was isolated from the reduction, and it suppresses the metabolism of rhizosphere of wetland plants (reproduced from sulfate reducers and methanogens that do not Neubauer et al, 2002). compete effectively with Fe(III)-reducing bacteria for carbon substrates (Section 8.08.1). Many of the new developments in this area have come from the wetland soils, these data suggested that microbial work of Roden and colleagues, who investigated a Fe(II) oxidation in wetlands is Fe(II)-limited. freshwater wetland where nonenzymatic Fe(III) The bacteria fared best in competition with reduction by sulfides was relatively unimportant nonenzymatic oxidation when the total iron that (see synthesis by Roden et ah, in press). They had accumulated in the chambers was low, proposed a model in which Fe(II)-oxidizing and suggesting that competition is mediated in part Fe(III)-reducing bacteria are spatially organized by factors such as particle size distribution, into aggregates reminiscent of those that mediate texture, and mineralogy. Each of these factors anaerobic methane oxidation (Figure 10; Section 8.08.4.5). The aggregates have an Fe(III)-coated influences the abundance and nature of surfaces particle at the center, surrounded by a mass of that can adsorb Fe(II) and coordinate oxidation Fe(III)-reducing bacteria, and then an outermost reactions. Finally, the presence of Fe(II)-oxidizing layer of Fe(II)-oxidizing bacteria (Sobolev and bacteria actually slowed the rate of overall Fe(II) Roden, 2002). The Fe(II)-oxidizing bacteria oxidation in some circumstances. This presum- enhance Fe(III) reduction by two mechanisms: (i) ably was caused by the inhibition of nonenzymatic they generate a steep 02 gradient that maintains Fe(II) oxidation via chelation and stabilization of anaerobic conditions inside the aggregate, and (ii) aqueous Fe(II), perhaps by exopolymers or other they produce compounds that chelate Fe(III) so that organic extracellular molecules that are produced it can diffuse back into the anoxic center of the by Fe(II)-oxidizing bacteria (Emerson and Moyer, aggregate where it will be re-reduced to Fe(II). The 1997). Similar compounds have been proposed to Fe(II) produced by Fe(III)-reducing bacteria dif- chelate Fe(III), allowing it to diffuse into anaero- fuses toward the periphery of the aggregate, bic zones after microbial oxidation (Sobolev and completing the iron cycle. The end result would Roden, 2001). be the oxidation of organic carbon without the net There is currently no satisfactory way to consumption or depletion of Fe(III) oxides. Certain separate enzymatic from nonenzymatic Fe(II) details of this model are supported by experimental oxidation in situ. However, pure culture studies evidence (Roden et al., in press), but this model have shown that bacteria mediated up to 90% of should be considered speculative. It is interesting Fe(II) oxidation (Emerson and Revsbech, 1994; to further speculate that these aggregates might Neubauer et al, 2002; Sobolev and Roden, 2001). also include bacteria that are capable of linking In contrast to these studies, van Bodegom et al. anaerobic Fe(II) oxidation to denitrification, with (2001) reported that microbial Fe(II) oxidation the NOJ provided by nitrifying bacteria at the was insignificant compared to nonenzymatic periphery. oxidation in a model rice system. However, their Iron cycling may also be favored by processes experiment included periods of vigorous sample that operate at larger spatial and temporal scales agitation that may have artificially favored abiotic than those proposed by Roden et al. (in press). 374 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes In wetland soils, iron cycling is promoted by several factors related to the presence of plants (Kostka and Luther, 1995; Roden and Wetzel, 1996; Frenzel et al, 1999). Roots introduce 02 into anaerobic soils (Schiitz et al, 1991; Hines, 1991; Holmer et al, 2002), which supports biotic and abiotic Fe(II) oxidation (Neubauer et ah, H,S 2002), and the deposition of poorly-crystalline Fe(III) oxides (Mendelssohn et al, 1995). Due to repeated annual cycles of root growth and senescence, non-rhizosphere soils (i.e. mm or more from a root surface) in wetlands are enriched Organic S in poorly-crystalline Fe(III) oxides compared to sediments that lack plants altogether (Weiss, Figure 24 A simplified biological redox cycle for 2002). The combination of abundant poorly- sulfur. crystalline Fe(III) minerals and labile organic carbon from roots favors rapid Fe(III) reduction the abundance of sulfate (S04) in seawater. during periods of anoxia. Because variations in Many specialized, small-scale environments, plant development and photosynthetic activity can including hypersaline and geothermal ( cause the rate of 02 release by roots to vary and ) ecosystems also thrive on hourly, daily and seasonally (Risgaard-Petersen microbial sulfur transformations (e.g., Baas Beck- and Jensen, 1997), parts of the root system may ing, 1934; Ehrlich, 2002). In marine environ- support net deposition of poorly-crystalline Fe(III) ments, a large fraction (up to 80%) of the organic oxides at times when 02 is released, then switch to carbon is respired through S04 reduction net Fe(III)-reduction when 02 is absent. Indeed, (Canheld et al, 1993a, b), so copious amounts Fe(III)-reducing and Fe(II)-oxidizing bacteria of sulhde are available for geochemical and occur on the same 1-cm lengths of plant roots microbial transformations. (Weiss et al, 2003). Burrowing animals influence sediments in much the same manner as roots (Kostka et al, 2002c; Gribsholt et al, 2003), and water table fluctuations can cause sediments to 8.08.7.1 Sulfur Geochemistry shift rapidly between reducing and oxidizing The average crustal abundance of sulfur is conditions. Although the processes discussed 260 (jLg g_1 and most on Earth is present as metal here have been demonstrated, their individual sulfide, gypsum, anhydrite, and dissolved sulfate. contributions to regulating Fe(III) cycling at an Sulfur can be found in a range of valence states ecosystem level is largely unknown. from the highly reduced sulfide (—2) to the most oxidized form in S04 (+6). There are several intermediate valence forms of sulfur that can serve 8.08.7 SULFUR as both electron donors and acceptors for bacteria depending on environmental conditions, the most Sulfur (S) is an important element biochemi- important being elemental sulfur and thiosulfate cally and geochemically. It is the fourteenth most (Odom and Singleton, 1993). Sulfate is highly abundant element in the Earth's crust and is —1% soluble and is the second most abundant anion in of the dry mass of organisms where it serves many freshwater (after bicarbonate) and in seawater structural and enzymatic functions. Sulfur acts as (after chloride) with the ocean acting as a major a significant electron donor and acceptor in many reservoir of S04 . Sulfate is also found in bacterial metabolisms (J0rgensen, 1988). evaporite deposits, primarily as gypsum Microbial sulfur transformations are closely (CaS04-2H20). Gypsum dissolution is a source linked with the carbon cycle; sulfur reduction of SO4- for microorganisms (Machel, 2001). coupled to organic carbon oxidation is a major Sulfate-containing minerals are common oxi- mineralization pathway in anoxic habitats, and dation products of sulfide minerals and can autotrophic sulfur oxidation can occur aerobically include anhydrous such as barite and anaerobically. Sulfur compounds are often (BaS04) and hydroxylated sulfates such as alunite highly reactive, which results in a tight coupling (KAl3(S04)2(OH)6) and jarosite (KFe3(S04)2 of the oxidative and reductive portions of (OH)6) (Deer et al, 1992). Elemental sulfur (S°) the biological , particularly at the is formed hydrothermally and as an oxidation redoxcline where sulfur cycling can be extremely product of sulfide weathering. However, S° is also rapid (Figure 24). Although sulfur cycling occurs a common product of sulfide oxidation by bacteria in terrestrial and freshwater aquatic environments, and can serve as an electron acceptor for bacterial it is most prominent in marine ecosystems due to respiration (J0rgensen, 1982a). Sulfur 375 Reduced S(II), or sulfi.de, is present in a variety Polymers of forms, most of which are solids. Dissolved (e.g., polysaccharides, proteins) sulfide exists as bisulfide ion (HS~) at neutral pH, sulfide ion (S2_) at alkaline pH, and hydrogen sulfide (H2S) at low pH. H2S is the only form of Monomers sulfide that is volatile and it imparts a character- (e.g., monosaccharides, istic "rotten egg" smell to sediments at low tide. amino acids) Dissolve sulfides react strongly with base and transition metal ions to form insoluble sulfide Fermentation minerals, the most prominent of which are iron sulfides that make anoxic sediments black. Freshly precipitated iron sulfides are amorphous, but tend to crystallize rather quickly into other acid-soluble authigenic minerals such as Mackinawite (tetra- V | H2S [>H2S gonal FeS) and Greigite (cubic Fe3S4). Many metals form insoluble sulfide minerals, which CO, Acetate H90 result in economically important deposits, and as SOj many as 95 sulfide minerals appear in standard S°^(c)(c)kS' H S

2 lactate + SO4 —• 2 acetate + 2C02 In essence, SO4 reducers in this case are 2 fermenting bacteria that depend on H2-consuming + 2H20 + S ~ (29) methanogens to maintain a low H2 partial pressure, or thereby making fermentation thermodynamically feasible. In addition, methanogenic bacteria can consume the acetate produced by the SO4 4 propionate + 3O4 —» 4 acetate + 4C02 reducers. 2 + 2H20 + 3S ~ (30) (ii) Electron donors. Energy sources used by SO4 -reducing bacteria are similar to those Some incomplete oxidizers are also capable of 2 - utilized by metal-reducing and methanogenic fermenting substrates without SO . respiration, bacteria. It is generally accepted that SO4- for example: reducing bacteria oxidize fermentation products- such as fatty acids, alcohols, and H (Christensen, 3 lactate —• 2 propionate + acetate + C02 (31) 2 1984; Parkes et al, 1989; S0rensen et al, However, some strains of SO2.- reducers are 1981), but this list is incomplete and is continually capable of fermentations that yield H2 as an end growing as new metabolisms are discovered. product that must be removed quickly by a Sulfate reducers can also metabolize chemolitho- bacterial partner. It was noted early that strains trophically (autotrophically) when utilizing H2 as produced acetate as an end product, so SO2," an electron donor. Since most of the substrates reducers were regarded as fermenting bacteria for used by SO4" reducers are provided by a variety Sulfur 377 of other bacteria, these compounds represent Pfennig, 1982). Thus, S04 reducers contribute to rapidly recycled intermediates that accumulate DNRA (Section 8.08.5.4). quickly when SO4 reduction is inhibited by Iron reduction has been observed in SO4- the use of SO4" analogues such as the group reducing bacteria (Bale et al, 1997; Knoblauch IV oxyanions molybdate and selenate (Oremland et al, 1999b; Lovley et al, 1993b), but this ability and Capone, 1988; Smith and Klug, 1981; Taylor is more ubiquitous in the non-SO^" -reducing and Oremland, 1979). members of the 8 Proteobacteria (Kostka et al, Besides the common fermentation products 2002a; Nielsen et al, 2002), and cell growth has listed above, SO^-reducing bacteria have been not been observed. Sulfate reducers are capable of shown to consume a variety of other substrates reducing a variety of other metals such as including xenobiotics and other aromatic com- uranium, chromium, technetium, and gold, but pounds (Bolliger et al, 2001; Elshahed and like iron, no growth occurs (Lovley and Phillips, Mclnerney, 2001b; Kniemeyer et al., 2003; Kuever 1992b, 1994b; Lovley et al, 1993b). The et al., 2001; Lovley et al., 1995), carbohydrates reduction of arsenate to arsenite supports growth such as fructose and sucrose (Sass et al., 2002), of SO^-reducing bacteria and, in some cases, amino acids (Burdige, 1989; Coleman, 1960), arsenate is the preferred electron acceptor (Macy et al, 2000; Newman et al, 1997a,b; Stolz and alkanes and alkenes up to C20 (Aeckersberg et al., 1991, 1998), phosphite (Schink et al, 2002), Oremland, 1999). Sulfate reducers can reduce aldehydes (Tasaki et al., 1992), dicarboxylic acids selenate (Stolz and Oremland, 1999) and Mn(IV) (Postgate, 1984), glycolate (Friedrich and Schink, (Tebo and Obraztsova, 1998). Other entries on the list nonsulfur inorganic 1995), methylated nitrogen and sulfur compounds electron acceptors used by SO|~ reducers are car- (Finster et al., 1997; Heijthuijsen and Hansen, bonate, which is reduced to acetate (Klemps et al, 1989; Kiene, 1988; van der Maarel et al, 1996a), 1985), and 02 (Billing and Cypionka, 1990). acetone (Platen et al, 1990), and sulfonates such Oxygen reduction is a relatively common feature as taurine and cysteate (Visscher et al, 1999). of SO4 -reducing bacteria (Cypionka, 2000; Sulfate-reducing bacteria are able to use a Dannenberg et al, 1992). However, aerobic variety of organic compounds as both electron growth in pure cultures is poor or absent and it acceptors and electron donors. For example, appears that 0 reduction, despite being enzy- Desulfovibrio species can ferment fumarate or 2 matic, is primarily an 02 removal mechanism malate, but will reduce these species to succinate (Cypionka, 2000). Many of the SO4"-reducing if an additional electron donor is available (Miller bacteria isolated from oxic environments belong and Wakerley, 1966). to the genus Desulfovibrio, although other genera (iii) Electron acceptors. Although the dominate in some instances (Krekeler et al, 1997; reduction of SO|~ is considered to be the classic Sass et al, 1997; Wieringa et al, 2000). Organic _ role of SC>4 -reducing bacteria, these organisms compounds that can serve as electron acceptors are capable of utilizing a wide variety of inorganic for SO4 reduction include malate, aspartate, sulfur compounds as electron acceptors including cysteine, sulfonates, pyruvate, acrylate, and oxi- sulfite, bisulfite, metabisulfite, dithionite, tetra- dized glutathione (Rabus et al, 2000). thionate, thiosulfate, dimethylsulfoxide, sulfur Sulfate-reducing bacteria couple the reductive dioxide, and elemental sulfur (S°) (Fitz and dehalogenation of aromatic compounds to growth Cypionka, 1990). Although some species (DeWeerd et al, 1990; Dolfing and Tiedje, 1987), can grow with S° as an electron acceptor, the and both chlorinated benzoates and bromophenols growth of many species is inhibited in the are used as electron acceptors (Boyle et al, 1999; presence of S° (Bak and Widdel, 1986; Widdel Mohn and Tiedje, 1990). Separate populations of and Pfennig, 1982), presumably due to increases SO4 reducers can be either sources (indirectly) in redox potential (Rabus et al, 2000). or sinks for acrylate. Acrylate is formed during Nitrate and NO^ reduction are rather wide- the breakdown of the marine osmolyte dimethyl- spread in SO4 reducers (Keith and Herbert, 1983; sulfoniopropionate, a process that is catalyzed by McCready et al, 1983; Mitchell et al, 1986; SO4 reducers in sediments, and then reduced by Momaetal, 1997; Seitz and Cypionka, 1986), and other SO^-reducing species (van der Maarel in some cases NO^~ is preferred over SO|~ (Seitz ef a/., 1998, 1996d). and Cypionka, 1986). The bisulfite reductase enzyme in SO4 reducers displays some activity toward N0 , which partially explains the ubiquity 2 8.08.7.3 Taxonomic Considerations of NOJT reduction. However, it appears that specific N02 reductases are also present (Liu and Sulfate-reducing bacteria are a complex physio- Peck Jr., 1981). Unlike denitrifying bacteria that logic group and classifying them has traditionally reduce NO^~ to N2, the end product of NO^ required the consideration of several properties, reduction in SO4" reducers is NH^ (Widdel and the most important of which are motility, cell 378 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes shape, the guanine plus cytosine content of DNA, phylogenetically distinct from all other S04~- the presence of desulfoviridin and cytochromes, reducing bacteria. The genus exhibits a great growth temperature, use of various electron nutritional versatility comparable to that of non- donors, and the ability to conduct complete or spore-forming sulfate reducers, including the use incomplete oxidation. New analysis of ribosomal of H2, alcohols, fatty acids, other aliphatic mono- RNA (rRNA) sequences has allowed for a more carboxylic or dicarboxylic acids, alanine, hexoses, thorough organization of the SO4 -reducing or phenyl-substituted organic acids as electron bacteria into four major groups: gram-negative donors for dissimilatory SO4" reduction (Widdel mesophilic, gram-positive spore forming, thermo- and Pfennig, 1999). Desulfotomaculum species are philic bacterial, and thermophilic Archeal (Castro not considered to be particularly important in etal., 2000). The gram-negative mesophilic group marine sediments compared to the non-spore- - of SO4 reducers is placed within the delta (S) forming SO|~ reducers. However, they are subdivision of the Proteobacteria and includes easily isolated from organic-rich sediments and two major families, the Desulfovibrionaceae and they possess metabolic capabilities that are known Desulfobacteriaceae although many genera fall to be important in sediments. Their ability to form outside of these families. The Desulfovibriona- spores gives them a competitive advantage in some ceae includes the genera Desulfovibrio and habitats and they are found to dominate environ- Desulfomicrobium, and this family appears to be ments like rice paddy sediments that undergo rather phylogenetically distinct (Devereux et al., wetting-drying cycles that could harm nonsporing 1990). The Desulfobacteriaceae family is much cells (Rabus et al., 2000). Dtm. acetoxidans was the less distinct and includes several genera first SO4" reducer isolated that was capable of (perhaps >20), including many of the complete acetate oxidation (Widdel and Pfennig, 1977). oxidizing species (Castro et al., 2000; Widdel and Since the early 1980s, many new lineages of Bak, 1992). The gram-positive spore-forming - - SO4 reducers have been isolated and described. SO4 reducers constitute primarily members of The phylogeny of SO4 reducers has expanded to the genus Desulfotomaculum. include the family Desulfobacteriaceae (Widdel Prior to the late 1970s, only two genera of and Bak, 1992), which includes many new genera SO^-reducing bacteria were known, Desulfovi- that are capable of complete and/or incomplete brio and Desulfotomaculum (Widdel, 1988; Wid- oxidation; the genera Desulfobacter, Desulfobac- del and Bak, 1992). The desulfovibrios have terium, and Desulfococcus to name a few. received the most attention because they are Members of the filamentous, gliding SO4"-redu- relatively easily isolated from the environment cing genus Desulfonema also appear to be and are not difficult to maintain in laboratory common inhabitants of organic-rich sediments, culture. They were originally described as gram- negative bacteria that are curved rods, do not especially within sharp redox gradients (Fukui et al, 1999). Sulfate reducers as a whole are produce spores, and utilize primarily H2 and lactate as electron donors (Postgate and Campbell, phylogenetically distinct from other bacteria, 1966). However, the group has been expanded to which has led to the discovery of signature DNA include members that are capable of using several sequences of ribosomal subunit genes that have other electron donors and electron acceptors, the been used as oligonucleotide probes and poly- merase chain reaction primers for the detection, latter including NO^, 02, and metal oxides (Barton et al, 1983; Cypionka, 2000; Lovley determination of relative abundance, and micro- et al, 1993b; Moura et al, 1997). They are the scopic visualization of members of the group classic examples of SO4 reducers that conduct (Amann et al, 1990a,b; Daly et al, 2000; incomplete metabolism with acetate as an import- Devereux et al, 1992; Devereux and Stahl, ant end product. Phylogenetic analyses have 1993). Probes specific for SO4 -reducing bacteria shown the desulfovibrios to be sufficiently diverse have been applied to marine water columns and distinct to warrant placement within their own (Ramsingef al, 1996; TeskeetaL, 1996), family (Devereux et al, 1990). Desulfovibrio and bacterial mats (Fukui et al, 1999; Minz et al, species are routinely isolated from marine sedi- 1999a; Ramsing et al., 1993; Santegoeds et al, ments and probably are an important component 1999), and sediments (Hines et al., 1999; of SO4 reduction in the sea. The desulfovibrios LlobetBrossa et al, 1998; Rooney-Varga et al., are the only group observed to enter into 1997; Sahm et al, 1999b; Sass et al, 1998). syntrophic relationships with methanogenic bac- In addition, reverse sample genome probing teria (Fiebig and Gottschalk, 1983; Pankhania (Voordouw et al, 1991), (Wawer ef a/., 1988). et al, 1997) and dissimilatory sulfite-reductase Desulfotomaculum species are gram-positive, genes (Minz et al, 1999b; Wagner et al, 1998), rod shaped, spore-forming SO^-reducing bacteria and denaturing gradient gel electrophoresis that as a group display wide phylogenetic (Okabe et al, 2002) have been used to study diversity (Rabus et al, 2000). They are also quite SO4- reducers in nature. Sulfur 379 8.08.7.4 Sulfate-reducing Populations acetate-utilizing sulfate reducers were most abun- dant in a marine (Visscher et al, The population composition of S04~-reducing 1992; Teske et al, 1998), while ethanol utilizers bacteria has been investigated using culturing, greatly outnumbered acetate utilizers in a salt biochemical, and genetic methods. In general, marsh sediment (Hines et al, 1999). estimates of abundance using viable counting Molecular analyses have furthered the descrip- methods such as colony counts on solid media or tion of SO2.- -reducing groups in depositional growth in liquid media after serial dilutions (most 2 environments. Both whole-cell in situ hybridiz- probable number (MPN) methods), are 10 - ations and hybridizations using RNA extracted 105 ml-1, which appear low when considered in 2- from sediments have been employed to investigate terms of the rate of in situ SO . reduction and the diversity of SO2.- reducers (Ramsing et al, culture estimates of rates per cell. This discre- 1993; Rooney-Varga et al, 1997). Incomplete pancy is undoubtedly due to the inability of viable oxidizing groups of sulfate reducers seem to counting techniques to recover the majority of dominate in marine sediments, primarily members bacteria present. However, MPN techniques yield of the Desulfovibrionaceae, but also Desulfobul- 2 - more realistic estimates of SO . reducer abun- bus species are abundant. However, the distri- 8 _1 dance (10 -10 ml ) when applied to organic- bution of groups varies with depth and among rich habitats such as marine microbial mats habitats. For example, members of the Desulfo- (Visscher et al., 1992; Ramsing et al., 1993) and bacteriaceae and Desulfovibrionaceae were salt marsh sediments (Hines et al, 1999). equally abundant in the upper 2.0 cm Due to the complex nature of the anaerobic of sediments, but the incomplete oxidizers bacterial food web, it is generally held that SO2.- (Desulfovibrionaceae species) dominated at reducers account for only —5% of the total greater depths (Devereux et al, 1996). Complete bacteria present despite their important role at oxidizers were essentially absent from Arctic the end of the food web (Devereux et al., 1996; Ocean sediments (Sahm et al, 1999a), yet they Li et al, 1999; LlobetBrossa et al, 1998). In situ dominated sediments inhabited by salt marsh hybridization techniques that use fluorescent grasses (Hines et al, 1999). Complete oxidizers, oligonucleotide probes to visualize individual i.e., Desulfobacteriaceae species accounted for cells of specific bacteria groups have yielded over 20% of the total recovered RNA in some SO2."-reducing bacteria counts in marine sedi- instances in the marsh, while Desulfovibrionaceae ments as high as 3 X 107 ml-1, which represented species were a small fraction in all cases (Hines up to 6% of the total Bacteria (LlobetBrossa et al, etal, 1999; Rooney-Varga et al, 1997). Members 1998). Hybridizations of bulk sedimentary RNA of the Desulfobacteriaceae are metabolically with probes also demonstrated that —1-6% of the diverse and may be well suited for environments total bacteria present in sediments were SO2.- that undergo widely changing seasonal redox reducing (Devereux et al, 1996), although this changes like those encountered in salt marsh technique demonstrated that 20% of the prokar- sediments and at the sediment surface. The yotes in a subtidal sediment were likely SO2.-- presence of high numbers of Desulfobulbus reducing (Sahm et al, 1999b). Sulfate reducers in species in marsh sediments (Hines et al, 1999) salt marsh sediments can account for >30% of the may reflect the ability of members of the genus to total bacteria (Hines et al, 1999). This high conduct sulfur disproportionation reactions, which percentage is likely due to the fact that marsh would be stimulated by plant-mediated redox grasses exude organic substrates from roots that cycling within the rhizosphere and the production can be used directly by SO2.- reducers, thus of intermediate redox states of sulfur. circumventing the need for fermenting bacteria. A wide variety of SO2, -reducing bacteria are found within sediments. Enumerations of bacteria 8.08.7.5 Factors Regulating Sulfate Reduction using MPN methods supplemented with specific Activity substrates for SO|~-reducing groups have shown 8.08.7.5.1 Sulfate-reducing activity the presence of bacteria able to use many substrates including lactate, ethanol, acetate, malate, and Knowledge of the role of SO2.- reduction in propionate (Laanbroek and Pfennig, 1981). Lactate sediments has increased greatly since the introduc- and acetate utilizers often outnumber other groups tion of the use of 35S as a tracer for determining illustrating the importance of both incomplete and sulfate reduction rates (J0rgensen and Fenchel, complete oxidizing SO2, reducers in sediments. 1974; Sorokin, 1962). At first, rates were estimated Sulfate-reducing marine sediments display an from the incorporation of 35S into dissolved sulfide abundance of even chain bacterial phospholipid (H2S) and acid volatile sulfides (iron monosulfides, fatty acids indicative of the presence of acetate- FeS), which were thought to be the only significant utilizing bacteria of the genus Desulfobacter products (J0rgensen, 1978a). However, it was (Parkes et al, 1993). Using MPN methods, determined that in some habitats, such as salt 380 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes marsh sediments and near the oxic/anoxic bound- competing electron acceptors (Crill and Martens, ary, a significant fraction of the reduced sulfur pro- 1987; Hines et ah, 1982; J0rgensen, 1977). In duced during SO|~ reduction is rapidly converted general, rates of activity can vary by factors of <5 to pyrite (FeS2), a compound that was previously to >30 between winter and summer, and these thought to be produced at rates on the order of changes have profound influence on redox con- several years (Howarth, 1979). Methods now ditions and the accumulation of reduced species routinely include a reduction step using reduced (J0rgensen, 1977). chromium under acidic conditions to incorporate all major reduced inorganic sulfur species that may form during SO4" reduction (Westrich, 1983; Fossing and J0rgensen, 1989; Meier et ah, 2000). 8.08.7.5.3 Carbon In highly organic soils and sediments, it is Sulfate reduction is controlled strongly by the usually necessary to recover the carbon-bonded 35 quantity and quality of organic matter present. In S label (Wieder and Lang, 1988). Sulfate marine sediments, there is a strong relationship reduction rates can also be determined from losses _ between organic sedimentation rate and SO4" of SC>4 in incubated sediments and from math- reduction rate (Goldhaber and Kaplan, 1975) and ematical models that describe the loss ofSO^ with it is generally held that SO4 reduction is limited sediment depth in terms of sedimentation and by organic matter availability except when SO4" diffusion (J0rgensen, 1978a,b,c). Depth increases concentration is quite low (Boudreau and Wes- in reduced sulfur compounds and the abundance trich, 1984; Dornblaser et ah, 1994; Westrich and of SO4"-reducing bacteria are useful as compara- Berner, 1984). Although SO|~ reducers usually tive indicators of the SO|~ reduction process, consume a relatively narrow suite of organic but they are poor indicators of actual rates of compounds, the tight coupling of SO4 reduction activity. with degradative processes prior to SO4" The rate of organic matter input (i.e., sedimen- reduction results in a stoichiometric relationship tation) and the availability of SO4 control the between the degradation of complex organic rate of SO4" reduction in sediments. Sulfate is matter and the reduction of SO|~ (Richards, rarely limiting in marine systems except in 1965). In general, there is a 2 : 1 molar relationship brackish estuarine waters and with depth in between labile carbon deposited into the SO4 sediments where SO4" has been depleted. Sulfate reduction zone and SO|~ reduced (Thamdrup and reduction rates in sediments can span several Canfield, 1996). The quantity of nitrogen and orders of magnitude, but typical near-shore rates phosphorus mineralized during SO4 reduction in the upper 5-10 cm of marine sediments are 3 1 can be predicted using C/N/P ratios of organic often 50-500 nmol cm" d" (Skyring, 1987). matter and SO4" reduction stoichiometry (Hines Rates in sediment with unusually rapid sedimen- -1 _1 and Lyons, 1982; Martens et ah, 1978). Sulfate tation rates can be reach 2,000 nmol ml d reduction activity responds rapidly to increased (Crill and Martens, 1987). Sulfate reduction in salt inputs of organic material with the seasonal marshes and microbial mats can reach rates as high deposition of spring phytoplankton blooms result- as 4,000 nmol ml"1 d"1 and 14,000 nmol ml"1 - _1 ing in significant increases in activity in lake d , respectively (Canfield and Des Marais, 1991; (Hadas and Pinkas, 1995) and ocean sediments Hines et ah, 1999). Anaerobic CH4 oxidation (Boetius et ah, 2000a). Although it is clear that supports a large fraction of the SO4" reduction in low-molecular-weight fatty acids are important some marine sediments (Table 6). substrates for SO4 -reducing bacteria, different groups of SO4 reducers may consume different acids (Boschker et ah, 2001). 8.08.7.5.2 Temperature Some electron donors are more readily metab- olized by methanogens than SO4 reducers and Sulfate reduction occurs over a wide range these "noncompetitive" substrates can allow (0-110°C) of temperatures (Castro et ah, 2000; methanogenesis to occur in the presence of active Elsgaard et ah, 1994; J0rgensen et ah, 1992; SOl" reduction (Oremland et ah, 1982). Knoblauch and Jorgensen, 1999; Knoblauch et ah, Examples of these types of substrates include Ci 1999a; Kostka et ah, 1999b; Sievert and Kuever, compounds like methylated amines, methylated 2000; Stetter et ah, 1993). Above -115 °C, SO4" sulfur compounds (e.g., dimethylsulfide and reduction is believed to occur only by thermo- methane thiol), and methanol. Methylated nitro- chemical reactions (Machel, 2001). Like other gen and sulfur compounds are common degra- biological processes, biological SO^- reduction is dation products of osmoregulating compounds affected strongly by temperature. Seasonal found in certain marine algae and salt marsh changes in SO4 reduction activity often follow grasses, and their use by methanogens partially temperature well, except for lags in activity due to explains the occurrence of significant concen- the time required to remove oxygen and other trations of CH4 in SOl~-containing estuarine and Sulfur 381 salt marsh sediments (Dacey et al, 1987; Kiene, Indeed, rates of S04~ reduction measured in oxic 1996b). In fact, the degradation of the osmoregu- regions of microbial mats can equal or exceed lant glycine betaine in marine sediments produces those noted in deeper anoxic (Canfield and acetate and , the former of which is Des Marais, 1991; Frund and Cohen, 1992; consumed by sulfate-reducing bacteria, while the Visscher et al., 1992). The complete oxidizing latter is consumed by methanogens (King, 1984). and gliding SO4" reducer, Desulfonema, is a Methanol is a degradation product of plant common inhabitant of the oxic-anoxic interface structural components such as pectin. in microbial mats (Fukui et al., 1999; Risatti et al., 1994; Teske et al., 1998), and is also abundant within the partially oxygenated rhizosphere of salt marsh plants (Rooney-Varga et al., 1997). Fila- 8.08.7.5.4 Sulfate and molecular oxygen mentous morphology, aggregate formation, and concentrations diurnal migration are all adaptations that allow The SO4 concentration in sediments affects this species to thrive in rapidly changing SC>4_ reduction only when concentrations are conditions. quite low. The reduction of SO|~ in marine sediments appears to be zero-order with respect to S04 to concentrations of ~2 mM (Boudreau and g>08>7>6 Microbial Reduction of Sulfur Westrich, 1984; Goldhaber and Kaplan, 1974). In freshwaters, SO4 concentrations must be much The microbial reduction of elemental sulfur lower before they limit SO4" reduction (Bak and (S°) was not discovered as early as the reduction Pfennig, 1991; Lovley and Klug, 1983b; Sinke of SO4". The first report of the sole use of S° as a et al., 1992). Because freshwater contains very terminal electron acceptor for growth was little SC>4_ compared to seawater, the importance reported well into the twentieth century (Pelsh, of SO4" reduction in sediments increases in an 1936), and the first pure culture (Desulfuromonas estuary as the salinity increases (Capone and acetoxidans) was isolated four decades later Kiene, 1988). Therefore, the vertical extent of the (Pfennig and Biebl, 1976). D. acetoxidans was SC>4_ reduction zone increases substantially as discovered as a partner living syntrophically with more SO4 becomes available, while the metha- an anaerobic phototroph that oxidized sulfide to S° nogenic zone is "pushed" deeper into the sediment and provided organic electrons donors. Hence, a and its contribution to carbon mineralization complete anaerobic sulfur cycle was maintained. decreases in importance. Several other S° reducers were isolated soon Sulfate-reducing bacteria are generally anaero- thereafter, which couple the reduction of S° with bic, but recent studies have shown that some the oxidation of compounds like acetate, formate, varieties are capable of O2 use, are able to lactate, fumarate, aspartate, dimethylsulfoxide, withstand several hours of full aeration, and are and H2,(Widdel, 1988). Some SO^-reducing bac- common inhabitants of oxic regions of microbial teria possess the ability to reduce S° (Biebl and mats (Cypionka, 2000; Krekeler et al, 1997). Pfennig, 1977; He et al, 1986), but this is Sulfate reducers withstand 02 stress better in the relatively rare. Certain Fe(III)-reducing bacteria presence of a co-metabolizing bacterium also possess the ability to reduce S°, and species that presumably consumes 02 (Gottschalk and have been isolated that are able to use a variety of Szwezyk, 1985). However, studies of co-cultures other electrons acceptors such as NO^~, Mn(IV), of a SO4 reducer and a facultative anaerobe thiosulfate, and 02 (Caccavo et al, 1994; Myers suggested the occurrence of 02-dependent growth and Nealson, 1988; Roden and Lovley, 1993). by the SO4" reducer in the absence of SO4" Dissimilatory S° reduction also occurs in bacteria (Sigalevich et al., 2000a,b; Sigalevich and Cohen, that are known to grow aerobically at normal 02 2000). tension (Balashova, 1985; Lovley et al, 1989; Sulfate-reducing populations within active Myers and Nealson, 1988). The ability to use sediments, like microbial mats, can exhibit a organic disulfide molecules, such as cysteine or bimodal distribution with two distinct maxima of glutathione, as terminal electron acceptors appears differing populations, one within surficial oxic to be restricted to certain members of the S°- layers and another in deeper anoxic sediments reducing bacteria (Pfennig and Biebl, 1976). (Minz et al, 1999a; Ramsing et al, 1993; Risatti Among the S°-reducing bacteria, several species et al, 1994; Sass et al, 1997). Sulfate reducers are able to generate ATP during the reduction of within or near oxic regions can benefit from S° (Hedderich et al, 1999). However, many labile organic compounds released by photosyn- members of the Archea are strictly fermentative thetic bacteria (Canfield and Des Marais, 1991; S° reducers that use S° reduction as an electron Teske et al, 1998; Visscher et al, 1992) or sink in lieu of respiratory S° reduction (Stetter, vascular plants (Scheid and Stubner, 2001; 1996). Since S° is highly insoluble, it is likely that Blaabjerg and Finster, 1998; Mines et al, 1999). many S° reducers utilize polysulfide as an electron 382 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes acceptor, or at least as an intermediate in S° Pure cultures have been isolated that are capable reduction (Hedderich et al., 1999). In general, of all these reactions including an isolate that is an the habitat of sulfur reducers is similar to that obligate disproportionator and does not reduce of SO4" reducers, so ecologically they coexist. SO4" (Finster et al, 1998). However, in most instances SO4 reducers Disproportionation reactions are important in produce more sulfide. sediments because they provide a mechanism for Several thermophilic microorganisms are anaerobic sulfur cycling (J0rgensen, 1990a), and capable of sulfur reduction (Bonch-Osmolovskaya SO4- reducers capable of this type of metabolism et al, 1990; Huber et al, 1992; L'Haridon et al., are numerically abundant (Bak and Pfennig, 1987; 1998) and this capability was employed in the J0rgensen and Bak, 1991). Thiosulfate appears to early isolation of thermophilic Archea (Stetter, be a key component of the sulfur cycle in 1996), some of which are capable of aerobic sediments because it can be oxidized, reduced, growth using sulfur as an electron donor and and disproportionated. Sulfate-reducing bacteria anaerobic growth using sulfur as an electron are key components of this cycling since they are acceptor (Segerer et al., 1985). In fact, most of capable of reducing SO4" and thiosulfate, and the sulfur-reducing bacteria known belong to the they are able to disproportionate thiosulfate. A Archeal domain (Hedderich et al., 1999). In the "thiosulfate shunt" seems to exist in sediments presence of sulfur, methanogenic Archea, where thiosulfate couples both reductive and especially thermophilic strains, reduce sulfur to oxidative pathways of the sulfur cycle (J0rgensen, H2S, while CH4 formation is highly curtailed 1990b) (Figure 26). Thiosulfate can be the main (Stetter and Gaag, 1983). Sulfur reduction is product of sulfide oxidation in reducing sedi- common in microorganisms situated within deep ments, and much of this can be disproportionated branches of the phylogenetic tree suggesting that with the remainder either oxidized to SO4" or sulfur reduction is a very ancient process (Stetter, reduced to sulfide. Hence, most of the sulfide 1997; Woese, 1987). produced during SO|~ reduction may be ulti- A Desulfuromonas sp. was the first eubacterium mately oxidized to SO4" after passing through a to be classified using rRNA techniques and it was thiosulfate intermediate. found to affiliate with the S-subclass of Proteo- Although S° disproportionation is important in bacteria and closely with the completely oxidizing the sulfur cycle, it is not an energetically favorable SO|~-reducing bacteria (Fowler et al., 1986; reaction under standard conditions (Bak and Liesack and Finster, 1994). However, sulfur Cypionka, 1987) and some have argued that it is reducing eubacteria are quite diverse and are not a major mode of metabolism in terms of found within other subclasses of the Proteobac- bacterial energetics or growth (J0rgensen and Bak, teria, many of which are not related phylogeneti- 1991). However, bacteria are able to grow via this cally to SO4 reducers (Lau et al., 1987; Schleifer process when metal oxides are present to rapidly and Ludwig, 1989). In general, the sulfur reducers remove sulfide (Lovley and Phillips, 1994a; as a group still require proper classification Thamdrup et al., 1993). Such sulfide scavenging (Widdel and Pfennig, 1999). by sedimentary metal oxides may allow the

8.08.7.7 Disproportionation It has been demonstrated that SO^-reducing bacteria have the ability to conduct inorganic fermentations of sulfur compounds, or sulfur disproportionation (Bak and Cypionka, 1987). The ability to disproportionate thiosulfate, sulfite Reduction Oxidation and elemental sulfur (S°) has been found in many SO^-reducing genera (Kramer and Cypionka, 1989), and it occurs as follows: Thiosulfate

+ S2O3" + H20 — SO4" + HS H (35) Elemental sulfur Figure 26 Thiosulfate disproportionation (heavy 4S° + 4H?0-»SOl +3HS 5H+ (36) lines). Typical oxidation (aerobic) and reduction (anaerobic) reactions are included (light lines). A Sulfite combination of these reactions leads to a thiosulfate shunt that allows for the anaerobic production of 4SCK H+ — 3S01 HS (37) intermediate redox states of S in anoxic sediments. Sulfur 383 process to act as a common metabolic transform- marshes than from true salt marshes. In most ation (Finster et al., 1998). cases, despite the fact that dissolved sulfide concentrations can be very high (mM) (Hines et al, 1989; King et al, 1982), only a small 8.08.7.8 Sulfur Gases portion of the gross production of sulfide escapes to the atmosphere (J0rgensen and Gaseous sulfur compounds are produced and Okholm-Hansen, 1985; Kristensen et al, 2000). consumed by microorganisms and they play a Bodenbender et al. (1999) reported H2S fluxes significant role in the global sulfur cycle because from intertidal marine sediments that were up to they connect the terrestrial, freshwater, and 2.6 X 104 times less than the rate of SO4" marine environments via the atmosphere, and reduction. Release of H2S is usually highest at they affect atmospheric chemistry and physics. night, because during the day, photosynthetic The most important volatile sulfur compounds microorganisms at the sediment surface either are H2S, dimethylsulfide (DMS), methane thiol directly consume sulfide or increase the pen- (MeSH), carbonyl sulfide (OCS), carbon disulfide etration of the 02 into sediments that enhances (CS2), and dimethyl disulfide (DMDS) (Bates chemical oxidation (Bodenbender et al, 1999; et al, 1992; Hines, 1996; Kiene, 1996a; Lomans Castro and Dierberg, 1987; Hansen et al, 1978; et al, 2002b). Reduced sulfur compounds are J0rgensen and Okholm-Hansen, 1985). Sulfide subject to chemical and photochemical oxidation emissions from intertidal sediments can increase in the atmosphere, which yield acidic products during flooding due to tidal pumping (J0rgensen that contribute to acid precipitation, and aerosol and Okholm-Hansen, 1985). Vegetated sediments particles that directly attenuate incoming solar tend to release less H2S due to the oxidation of radiation or lead to cloud condensation nuclei, sulfides by 02 released from roots (Hines, 1996). both of which affect the global radiative In general, sulfureta-like sediments, i.e., highly balance (Charlson et al., 1987; Kelly and reducing, unvegetated, and sulfide-rich environ- Baker, 1990; Panter and Penzhorn, 1980). ments, allow for a significant loss of H2S to the Early work predicted that H2S produced by - atmosphere since 02 penetration is minimal and dissimilatory SO4 reduction was the primary metals capable of precipitating sulfide are already biogenic sulfur gas emitted into the atmosphere removed. However, these environments are the (Bremner and Steele, 1978; Natusch and Slatt, exception rather than the rule. 1978). However, although the biogenic pro- duction of organosulfur gases has been known since the 1930s (Haas, 1935), it was only since the 1970s that the importance of these gases has 8.08.7.8.2 Methylsulfides been recognized (Graedel, 1979; Lovelock et al., 1972; Rodhe and Isaksen, 1980). It is now clear Numerous studies of DMS formation and that DMS emissions constitute the bulk of the consumption have appeared since the recognition sulfur that enters the atmosphere each year of the importance of this compound in marine (—75%, primarily from oceanic sources). See biogeochemistry and atmospheric chemistry. Chapter 8.14 of this volume for details on the role Indeed, DMS accounts for —90% of the biogenic of sulfur gases in the global sulfur cycle. sulfur emissions from the marine environment (Andreae and Crutzen, 1997). A link has been proposed among DMS production and flux, the 8.08.7.8.1 Hydrogen sulfide atmospheric oxidation of DMS, and the sub- sequent formation of SO4 aerosols and cloud The liberation of H2S is controlled not only by condensation nuclei (Charlson et al, 1987). In this the rate of its production by SO4 -reducing model, increasing global temperatures lead to bacteria, but also by its pH-dependent speciation, enhanced production of DMS, which attenuates its tendency to rapidly precipitate as metal further warming by radiative backscatter from sulfides, and its rapid chemical and biological aerosols and reflection of radiation from increased oxidation. Only the protonated species (H2S) is cloud cover. Although most work to date has been volatile, and at neutral pH, most inorganic sulfide conducted on the cycling of DMS in seawater, is present as bisulfide ion (HS), whereas considerable effort has also been invested in sulfide (S2_) dominates under alkaline conditions. understanding the biogeochemistry of DMS in These three species are known collectively as wetlands and sediments (Kiene, 1996b). SH2S. Hence, the escape of sulfide should be Several compounds can serve as precursors of enhanced at low pH. Sulfate reduction is most DMS including dimethylsulfonium compounds dominant in marine sediments and this is where such as dimethylsulfoniopropionate (DMSP), the highest emissions of gaseous H2S occur S-containing amino acids, MeSH, dimethylsulf- (Hines, 1996). However, DeLaune et al. (2002) oxide (DMSO), and methoxylated aromatic com- reported higher emissions of H2S from brackish pounds, which are capable of methylating 384 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes inorganic sulfide and MeSH (Bak et al., 1992; MeSH (van der Maarel et al, 1995, 1993). Finster et al, 1990; Kadota and Ishida, 1972; Sulfate-reducing bacteria have been isolated that Kiene and Taylor, 1988a,b; Kreft and Schink, are capable of cleaving DMSP to DMS and 1993; Lomans et al, 2002b). Methionine is acrylate (van der Maarel et al, 1996d) as well as ubiquitous in organisms and has been shown to demethylating DMSP (van der Maarel et al, be a precursor of several sulfur gases including 1996c). A SO4 -reducing bacterium has been DMS (Bremner and Steele, 1978; Kiene and isolated that cleaves DMSP and then reduces the Capone, 1988; Kiene and Visscher, 1987b; Zhang liberated acrylic acid (van der Maarel et al., 1996d). et al, 2000; Zinder and Brock, 1978c). Methion- Acrylate and SO4 reduction occurred simul- ine degradation produces MeSH, which can be taneously, but SO4 reduction yielded more subsequently methylated to DMS (Kiene and growth. Capone, 1988; Kiene and Visscher, 1987b). DMS (and MeSH) concentrations in freshwater S-methyl-cysteine can also be degraded to sediments are considerably lower than in marine MeSH, whereas S-methyl-methionine degrades habitats, primarily due to the low concentrations primarily to DMS (Kiene and Capone, 1988). of DMSP in freshwater plants (Bechard and DMSO reduction by bacteria also produces DMS Rayburn, 1979). However, some small ponds (Griebler, 1997; Zinder and Brock, 1978a; Zinder and wetlands exhibit DMS levels similar to and Brock, 1978b) and several SO4 -reducing those in seawater (3-25 nM) (Kiene and Hines, bacteria with this ability have been isolated (Bale 1995; Nriagu et al, 1987), and DMS emissions et al., 1997; Jonkers et al., 1996; van der Maarel from freshwater habitats can be significant (Hines, gfaZ., 1998). 1996; Kelly and Smith, 1990). It has been known DMSP is a major source of DMS in marine for sometime that freshwater bacteria have the systems (Kiene, 1996a). DMSP is produced by ability to form DMS from DMSP, and an obligate marine micro- and macroalgae and halophytic anaerobe with this ability was isolated early plants where it acts as an osmoregulant or as part (Wagner and Stadtman, 1962). Freshwater sedi- of an antioxidant system (Sunda et al., 2002; ments are capable of producing DMS from DMSP Yoch, 2002). Although many phytoplankton amendments even though these sediments may be contain a DMSP lyase enzyme, bacteria are nearly DMS-free (Yoch et al, 2001). Enrichments thought to be the primary degraders of DMSP in of these sediments with DMSP led to the isolation seawater (Kiene, 1992). Fungi associated with of a suite of gram-positive DMS-producing decaying salt marsh plants also possess a DMSP bacteria related to Rhodococcus spp. lyase (Bacic et al, 1998). Cleavage of DMSP can Most studies indicate that the primary path of lead to the production of DMS and acrylate, but DMS formation in freshwater sediments is via the DMSP can also be demethylated to 3-methylmer- methylation of inorganic sulfide to MeSH, which captopropionate, which can be demethiolated to is subsequently methylated to DMS (Bak et al, MeSH or demethylated to 3-mercaptopropionate 1992; Drotar et al, 1987a; Finster et al, 1990; (Kiene and Taylor, 1988a,b; van der Maarel et al, Kiene and Hines, 1995; Lomans et al, 2001a, 1996c; Visscher and Taylor, 1994). Members of 1997). Thiol methyltransferase enzymes that the Proteobacteria, especially the a-group, appear catalyze this methylation are relatively wide- to be important DMS producers in the sea spread in microorganisms including bacteria and (Gonzalez et al., 1999). However, DMS producers (Drotar et al, 1987a,b; Drotar and Fall, belonging to the /3-, S-, and ^/-subdivisions have 1985). Methoxylated aromatic compounds, which been identified in seawater (de Souza and Yoch, are degradation products of lignin, have been 1995; Jonkers et al, 1998; Ledyard et al, 1993; shown to be important as methyl donors for the van der Maarel et al, 1996b) and members of the methylation of sulfide and MeSH (Finster et al, ^-subdivision dominated DMS-producing isolates 1990; Kiene and Hines, 1995), and several from a salt marsh (Ansede et al, 2001). It has been anaerobic bacteria have been isolated that can hypothesized for aerobic seawater that the path- perform this sulfide-mediated O-demethylation way used for DMSP degradation is controlled by reaction (Bak et al, 1992; Lomans et al, 2001a; DMSP concentrations and the demand for sulfur, Mechichi et al, 1999). The hydroxylated aromatic and that DMS formation is usually favored at high residue that remains after demethylation is DMSP concentrations (Kiene et al, 2000). degraded to acetate or acetate and butyrate Both direct cleavage of DMSP to DMS/acrylate (Kreft and Schink, 1993). This differs from the and DMSP demethylation pathways have been more common pathway for acetate production in observed in anoxic sediments and oceanic waters which homoacetogenic bacteria transfer methyl (Kiene and Taylor, 1988b; Yoch, 2002). However, groups from methoxylated aromatic compounds to unlike oxic environments, the demethylation of CO to produce acetate (Bache and Pfennig, 1981). DMSP in anoxic sediments appears to require DMS and other volatile organo-sulfur com- two organisms, one to demethylate DMSP to pounds rarely accumulate in sediments, because 3-methiolpropionate, followed by another to form they are consumed by microorganisms as rapidly Sulfur 385 as they are produced (Lomans et al., 2002a,b). similarly to other Q compounds such as methanol Several bacteria are able to degrade DMS and and methylated amines (Ferry, 1999), but the MeSH aerobically, but most of those isolated consumption of methylated sulfur compounds are Thiobacillus, Methylophaga or Hypho- appears to be due to separate inducible enzymes microbium spp. (Cho et al., 1991; de Bont et al., (Ni and Boone, 1993). The methyltransferase 1981; deZwart et al., 1996; Sivela and Sundmann, responsible for the use of DMS in Methanosarcina 1975; Smith and Kelly, 1988a; Visscher et al, barkeri is distinct from those used for acetate, 1991). Many bacteria are capable of oxidizing methanol, or methylated amines (Tallant and DMS to DMSO in the presence of an additional Krzycki, 1997; Tallant et al, 2001). The methano- carbon source (Zhang et al., 1991). genic degradation of two moles of DMS yields Zinder and Brock (1978d) were the first to three moles of CH4 and one of C02. One mole of report the anaerobic conversion of DMS and methyl groups is oxidized to C02, which generates MeSH to C02 and CH4 in freshwater lake reducing equivalents for the reduction of three sediments and sewage sludge. Since then, this moles of methyl groups to CH4 (Finster et al, 1992; process has been reported to occur in many anoxic Kiene et al, 1986). Complete degradation of environments including salt marsh and mangrove methylated sulfides to C02 by methanogens may sediments, hypersaline lakes, and a variety of occur when H2 is removed by hydrogenotrophic freshwater sediments (Jonkers et al., 2000; Kiene, bacteria such as SO|~ or NO^ reducers (inter- 1988; Kiene et al, 1986; Lomans et al, 1999a, species H2 transfer) (Lomans et al, 1999c). 2001b; Lyimo et al, 2000, 2002). Degradation of Methanogens in oligotrophic peat in northern DMS and MeSH in anaerobic environments is bogs do not consume DMS, which accumulates catalyzed primarily by methanogenic, SO|~- during incubations (Kiene and Hines, 1995). It was reducing, NO^-reducing, and phototrophic bac- shown that acetate accumulates in these peats as teria (Kiene, 1988, 1991a; Kiene et al, 1986; Liu well (Hines et al, 2001). et al, 1990; Lomans et al, 2002b; Lyimo et al, 2000; Oremland et al, 1989; Tanimoto and Bak, 1994; Visscher et al, 1995; Zeyer et al, 1987). 8.08.7.8.3 Carbonyl sulfide and carbon Despite the ubiquity of CH4 formation from DMS and MeSH degradation, SO|~-reducing bacteria disulfide appear to dominate degradation in marine sedi- Bacteria are also involved in the formation and ments (Kiene, 1996a) and SO|~-reducing bacteria degradation of other volatile sulfur compounds are known to degrade DMS in freshwater such as carbonyl sulfide (OCS) and carbon environments as well. Methanogens and S04- disulfide (CS2). Much less is known about the reducing bacteria appear to compete for DMS, but bacterially mediated transformations of these methanogens out-compete S04 reducers at high compounds compared to the methylated sulfur DMS concentrations even when S04 is abundant species, especially with respect to the role of (Lomans et al, 2002b). The role of SO4" anaerobic bacteria. They both constitute a rather reduction in DMS degradation is based primarily minor portion of the global flux of biogenic sulfur on the observed decrease in DMS consumption gases to the atmosphere (Bates et al, 1992; Hines, when SO4- reduction is inhibited by molybdate or 1996). Carbonyl sulfide is a long-lived species in tungstate (Kiene and Visscher, 1987a). However, the atmosphere and is the most abundant gaseous a pure culture of a DMS-utilizing SO4 reducer form of S, whereas CS2 has a low atmospheric has been isolated from a thermophilic digester abundance. Both species appear to form photo- (Tanimoto and Bak, 1994). The energetic contri- chemically in aquatic systems, but both also can bution of transformation of methylated sulfur be formed biologically from a variety of organo- compounds to anaerobic bacteria remains unclear, sulfur precursors including S-containing amino but the of the use these com- acids (Banwart and Bremner, 1975; Bremner and pounds by methanogens and SO^- reducers has Steele, 1978) and more unusual sulfur compounds recently been evaluated (Scholten et al, 2003). A like djenkolic acid (CH2[SCH2CH(CH2)COOH]2) pure culture of a DMS-degrading denitrifying and lanthionine (S[CH2CH(NH2)COOH]2) (Piluk bacterium has also been isolated (Visscher and et al, 1998). In general, these gases are formed Taylor, 1993b). anaerobically in organic-rich environments, but Several methanogens capable of using DMS the mechanisms of formation are unclear in most have been isolated from marine and hypersaline instances (Bremner and Steele, 1978; Kiene, environments (Finster et al, 1992; Kiene et al, 1996a; Wakeham et al, 1984). Both OCS and 1986; Oremland et al, 1989; Rajagopal and CS2 are toxic to certain bacteria (Borjesson, 2001; Daniels, 1986), and a freshwater methanogen Bremner and Bundy, 1874; Seefeldt et al, 1995) with this capability has also been isolated and and their release from roots has been implicated represents a new genus (Lomans et al, 1999b). The as a plant defense mechanism (Kanda and Tsuruta, degradation of DMS and MeSH appears to occur 1995). Both OCS and CS2 are consumed 386 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes aerobically, primarily by bacteria involved in green bacteria) and a few cyanobacteria species. oxidation of inorganic sulfur compounds (Kelly The electrons of HS are donated to PS I (Section and Baker, 1990; Kelly et ai, 1994). They are also 8.08.2.1) and utilized to generate reducing consumed in anaerobic sediments (Zinder and biochemical equivalents (NAD(P)H) and ATP. Brock, 1978a), but the mechanism for this In contrast to oxygenic photolithoautotrophs, the 2 - consumption is unresolved. CS2 can be oxidized oxidation product of photosynthesis is SO . to OCS and H2S anaerobically (Smith and Kelly, instead of 02. In both CSB and phototrophic 1988b) and denitrifying bacteria have been sulfur bacteria, much of the energy generated from isolated that are capable of degrading CS2 sulfur oxidation is used to assimilate C02 into cell anaerobically (Jordan et ah, 1997). material. Other reduced sulfur compounds that play a significant role in oxidation-reduction reactions 8.08.7.9 Microbial Oxidation of Sulfur in the environment include inorganic compounds such as FeS, FeS , S 02 , S2_(polysulfides), Sulfide and a wide variety of additional organic 2 2 SyOl~ (polythionates), and S°; organic com- and inorganic reduced sulfur compounds can be pounds include methylsulfides (CH SH, used as electron donors for microbial redox 3 (CH3)2S). Clearly, in sulfidic environments, reactions. In addition to its importance in "natural" reduced sulfur compounds represent an important depositional and hot spring environments, sulfur source of energy for chemo- and photolithoauto- oxidation plays a significant role in sewage sludge, trophic sulfur oxidizers. However, sulfur oxidizers paper- and wood-mill effluents, coal desulfuriza- must compete with a variety of biotic and abiotic tion, and leaching of ores and minerals. Sulfide- reactions for terminal electron acceptors. oxidizing microbes remove noxious sulfides (organic and inorganic) and interact with minerals through oxidation of metal sulfides (e.g., FeS, 8.08.7.9.1 Colorless sulfur bacteria FeS2), deposition of elemental sulfur (S°), precipitation of gypsum and dissolution of calcium CSB gain energy through chemolithotrophic carbonate (Baas Becking, 1934; Bos and Kuenen, sulfur-oxidation using 02 or NO3" as terminal 1990; Edwards et al., 2000a; Ehrlich, 2002; Lens electron acceptors. CSB include a wide variety of and Kuenen, 2001; Visscher and Van Gemerden, morphologically and phylogenetically distinct 1993). groups (e.g., filamentous and Thioploca The Gibbs free energy yield of carbon oxidation 2 - spp. and single-cell Thiobacillus and Thiomicros- with SO . as a terminal electron acceptor is about pira spp.). The filamentous CSB have a conspic- 4-10 times lower than with oxygen. However, uous morphology, with cell sizes that allow 2 - sulfide, the product of SO . reduction is an observation with the naked eye. For example, excellent electron donor; the reaction of which namibiensis has a cell diameter of with oxygen: up to 0.75 mm (Schulz et ai, 1999; Schulz and Jorgensen, 2001) and filaments of certain Thio- HS" + 20, — SOI + FT (38) ploca species may be as long 70 mm (J0rgensen has a AG0/ of -798kJmol"1. Alternatively, and Gallardo, 1999). Not surprisingly, Wino- anaerobic oxidation using NO^, which couples gradsky (1887) made his early observations of the sulfur and nitrogen cycles, provides ample "chemosynthetic" sulfur oxidation in Beggiatoa energy as well: spp. It has been claimed that the Thioploca- Beggiatoa-Aomm&tcd mats in the upwelling area

5HS + 8N03 — 5SOJ + 4N2 + 30FT H20 off the coast of Chile embody the largest microbial ecosystem, estimated 104 km2 (J0rgensen and (39) Gallardo, 1999), although it is possible that 1 with a AG °' of - 744 kJ mol" H2S. Chemolifho- other such systems exist (Gallardo et al., 1998; trophic microbes that derive their energy in this Namsaraev et al., 1994). However, the single- fashion are collectively referred to as colorless celled CSB are abundant in most sulfur-containing sulfur bacteria (CSB) in contrast to a physiologi- ecosystems and are metabolically diverse (Kelly, cally distinct group of pigmented organisms (e.g., 1982, 1988; Kelly et al, 1997; Kuenen and purple-sulfur bacteria (FSB)). The latter couple Beudeker, 1982). Thus, they can be considered the oxidation of sulfur to photoautotrophic C02 to be at least equally important from a global fixation (Van Niel, 1931, 1941): perspective.

HS 2C02 + 2H20 — SO4 + 2[CH20] + W 8.08.7.9.2 Single-cell colorless sulfur bacteria HS and certain other reduced sulfur-compounds can serve as electron donors for photosystem (PS) Single-celled, rod-shaped microbes that I in anoxygenic photolithotrophs (purple and derive energy from oxidation of reduced Sulfur 387 sulfur-compounds were traditionally classified as has a high growth yield and requires ~2 mol of Thiobacillus spp. (Vishniac and Santer, 1957), HS~ oxidized per mole C02 fixed. Members of whereas smaller spirillum-shaped species were the second group include T. thioparus, Thiomicrospira (Timmer-Ten Hoor, 1975). Based T. aquaesulis and T. denitrificans. Autotrophic on molecular phylogeny, the validity of early thiobacilli deploy the Calvin cycle for C02 classification has recently been challenged and the fixation and some have cytoplasmic inclusions phylogeny of single-celled CSB is currently under that contain crystalline Rubisco (Beudeker et al., revision. Furthermore, physiological studies 1980). Many CSB are not obligate chemolithoau- revealed the need for a careful description of the trophs, but are either facultative autotrophs products of sulfide-oxidation (Gottschall and (e.g., T. pantotropha, T.(or Paracoccus) denitri- Kuenen, 1980; Tuttle and Jannasch, 1977; ficans, T. (or Paracoccus) versutus) or so-called Vishniac and Santer, 1957). Many heterotrophic mixotrophs (chemolithoorganotrophs; e.g., T. microorganisms that seem capable of oxidizing novellus) that use sulfur-compounds as electron HS~ or S20|~ produce polythionates (S^Og) donors but use organic carbon for biomass. rather than SO4 as reaction products (Tuttle and Single-celled CSB are found in a variety of Jannasch, 1977, 1979). Some heterotrophs have environments, both terrestrial and aquatic, ranging been found to oxidize SO3 and other reduced from freshwater to hypersaline, with pH values sulfur compounds to SO4 (Sorokin and Lysenko, from <1 to >10, and displaying psychrophilic, 1993), but since thiosulfate utilization is generally mesophilic, and thermophilic temperature not tested as a metabolic trait, it is possible that responses. Several species are symbionts in many more heterotrophs possess the capacity to hydrothermal vent invertebrates (e.g., the tube oxidize sulfur-compounds. Although these hetero- worm Riftia pachyptilia, the bivalve Calyptogena trophs are not considered to be CSB, their role in magnifica, the mussel Bathymodiolus puteoser- HS~ oxidation in the environment may be pentis; Cavanaugh, 1994; Cavanaugh et al., 1981; significant (Sorokin and Lysenko, 1993; Tuttle Southward et al., 2001) as well as in bivalves and Jannasch, 1979). (e.g., Thyasira sp.) and oligochetes and clams Single-celled CSB oxidize a wide variety of in estuarine and other environments (Dubilier inorganic and some organic sulfur-compounds, et al., 1999; Krueger et al., 1996; Wood and mixtures of which are typically present in the Kelly, 1989). environment (Luther and Church, 1988; Luther Thiobacilli are typically present in high den- et al., 1999; Visscher and Van Gemerden, 1991; sities, especially in marine sediments near the Zopfi et al., 2001). In mixtures of sulfur- oxycline, where densities of 106—109 cells cm-3 compounds, CSB have a strong preference for have been reported (Gonzalez et al., 1999; HS~ (Kuenen and Beudeker, 1982), and under Sorokin, 1972; Teske et al., 1996; Visscher et al., oxygen-limitation, S°, polythionates and polysul- 1991, 1992, 1995). Similar densities could be fides are the oxidation products in species that do expected in sewage outfall and wastewater treat- not utilize NO^ (Van den Ende and Van ment facilities, where high rates of SR or of HS~ Gemerden, 1993). S° deposited outside the cell, loading can be expected (Lens and Kuenen, 2001). consists primarily of long-chained polythionates Some CSB species thrive in extreme environ- that form micelle-like structures (Steudel et al., ments. For example, T. ferrooxidans is commonly 1987). This "zero-valent" sulfur contains hydro- found in acid mine drainage where it oxidizes philic, charged sulfonate groups on the outside, Fe(II) to Fe(III), Cu(I) to Cu(II), and HS" to which explains how growth on highly insoluble SO4 . This organism is capable of pyrite oxi- "elemental" sulfur is possible. The enzymes dation and plays an important role in leaching of involved in the sulfide oxidation pathway, ores (e.g., copper, nickel, and uranium) and which are well characterized (Kelly, 1982; desulfurization of coal (Bos and Kuenen, 1990). Kelly et al., 1997), include two that oxidize T. thioparus is versatile in organic carbon sulfite to sulfate: and reverse utilization and oxidizes MeSH and DMS adenine phosphosulfate reductase (Kappler et al., (Kanagawa and Kelly, 1986; Smith and Kelly, 2001). The latter is highly efficient in energy 1988a; Visscher and Taylor, 1993b; Visscher and conservation and couples oxidation of sulfite Van Gemerden, 1991). Another Thiobacillus sp., directly to ATP generation. This could in part resembling T. thioparus, is able to degrade DMS explain the observation that two groups of with NO^~ instead of 02 and can also oxidize the thiobacilli can be distinguished based on the thiol group of a range of alkylthiols to sulfate amount of energy required for autotrophic growth (Visscher and Taylor, 1993a). (Kelly, 1982, 1988). The first group has a low The genus Thiomicrospira contains three growth yield and oxidizes ~4 mol HS ~ per mole known species, all of which are chemolithoauto- of C02 fixed into structural cell material. trophs. These organisms thrive under microaer- Members include Thiobacillus neapolitanus, ophilic conditions, and one strain can couple T. thiooxidans, T. versutus. The second group sulfide and thiosulfate oxidation to N07 388 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes reduction (Timmer-Ten Hoor, 1975). Another Thioploca spp. are sheathed filamentous CSB, exceptional genus of sulfide-oxidizing Eubacteria that resemble Beggiatoa spp. in many charac- is Achromatium, a facultative that teristics. They thrive in the 02-minimum zone stores S° and calcite intracellularly (Head et al., in marine and freshwater mud and seem to 2000) and is found in freshwater, brackish, and accumulate NOJ in their vacuoles (as do some marine environments. The three known species Beggiatoa spp.), which enables these organisms that belong to this genus all have mixotrophic to thrive under conditions where HS~ and 02 growth capabilities. The calcite deposits make or NO^ are not present simultaneously. The this organism one of the largest free-living NO^~ concentration in the vacuole can reach single-celled prokaryotic cells known (estimated 0.5 M, or more than 100 times that outside the biovolume of up to 8 X 104 |xm3 cell-1) (Schulz cell (J0rgensen and Gallardo, 1999). However, and Jorgensen, 2001). The role of the calcite is under anoxic conditions, NO^~ is reduced to obscure; it could perhaps act as a buffer in the NH| and not N2 (Otte et al, 1999). In addition when SO|~ is produced, or provide to a positive tactic response to 02, Thioploca C02 for autotrophic growth. The latter scenario is spp. also move towards NO^ and low concen- somewhat unlikely, since precipitation of CaC03 trations of sulfide; microaerophilic conditions yields H+, and CSB are actually believed to are preferred. Thioploca sheaths may dissolve this mineral when growing autotrophi- contain SO4"-reducing bacteria (Desulfonema cally (Visscher et al, 1998). spp.; (J0rgensen and Gallardo, 1999)) that may The Archeabacteria contain two genera of provide sufficient HS . Filaments that have sulfur-oxidizing organisms, both of which are characteristic tapered ends can live outside the hyperthermophiles. is an aerobic sheath, and S° is stored intracellularly. Thioploca sulfur-oxidizer, which is also capable of using spp. have a mixotrophic lifestyle and, based on Fe2+ and organic carbon (facultative chemo- cell diameter, four size categories can be distin- that grows chemoorganoheterotro- guished: two small groups (2.5-5 and 12- phically). Acidianus resembles Sulfolobus, but can 20 |jum), a larger group (30-43 |xm), and a group grow both in the presence and absence of 02. with gigantic cells (125 |xm). Although the mats Under oxic conditions it oxidizes S° to SO4", off the Chilean and Peruvian coast are quite while under anoxic conditions, S° is used as extensive (see above), Thioploca spp. have been electron acceptor and sulfide is produced. found at other offshore locations including the Arabian Sea, the Benguela Current, and the Mediterranean Sea, as well as in lakes (Lake Erie, Lake Baikal, Lake Biwa) (Gallardo et al., 8.08.7.9.3 Filamentous colorless sulfur 1998; J0rgensen and Gallardo, 1999; Namsaraev bacteria et al, 1994; Nishino et al, 1998). The filamentous CSB include the genera Thioploca, Thiothrix, and Beggiatoa (Larkin and Strohl, 1983; Schulz et al, 1999), which are 8.08.7.9.4 Anoxyphototrophic bacteria relatively closely grouped phylogenetically. Beg- giatoa spp. chemolithoautotrophs were the first to Purple and green sulfur bacteria use reduced - be described (Winogradsky, 1887), but later sulfur compounds (e.g., HS , S°, S* , S20i~) as investigations demonstrated that they might have electron donors for photosynthesis. Purple and a chemolithoorganotrophic (or mixotrophic) life- green nonsulfur bacteria seem to have limited style. These organisms are found in freshwater capabilities to use these donors as well. Interest- and at the interface of 02 and HS~ ingly, some PSB have the ability to oxidize (Schulz and Jorgensen, 2001). sulfide chemolithotrophically in the presence of Beggiatoa stores "elemental" sulfur inside the 02 (De Wit and Van Gemerden, 1987). For cell, which gives the filaments a white appearance. example, although Thiocapsa roseopersicina can Massive blooms of this organism in sulfidic hot use HS ~ phototrophically, in the presence of 02, springs, salt marshes, and marine sediments make it requires anoxia for synthesis of photopigments this gliding organism conspicuous (Larkin and (De Wit and Van Gemerden, 1987). If these Strohl, 1983) when it forms veils or mats. conditions are not met, T. roseopersicina will Hydrothermal vents at the deep seafioor support ultimately be nonpigmented and resort to chemo- sulfide-based ecosystems with unusually large trophic sulfide oxidation for energy requirements. Beggiatoa spp. (Jannasch et al., 1989) in addition Immediately following even a brief period of to unicellular CSB in trophosomes of Riftia spp. anoxia, the reverse shift from chemolithotrophy (Cavanaugh, 1994). In some Beggiatoa spp., the to photolithotrophy occurs. sulfide can be oxidized with 02 or NO^ as an A phylogenetic link between PSB and thioba- electron acceptor, and N2 is the product of nitrate cilli has been suggested, in which the PSB reduction (Sweerts et al., 1990). (and/or other phototrophic sulfur bacteria) Coupled Anaerobic Element Cycles 389 are considered to be the ancestors of certain were only able to coexist with the CSB when the thiobacilli. Blooms of phototrophic bacteria are ratio was less than 1.6 (indicating 02-limitation), observed in lakes (e.g., Guerrero et al., in which case the PSB used partially oxidized 1985; Overmann, 1997)), microbial mats sulfur compounds excreted by the CSB (Van den (Pierson et al., 1987; Van Gemerden, 1993) and Ende et al, 1996). in salt marshes (Baas Becking, 1934). Purple Although several physiological and ecological nonsulfur bacteria (PNSB) can use a wide variety aspects of sulfur-oxidation have been outlined of organic carbon compounds (Madigan, 1988) above, others exist, including the fermentation of and typically thrive in habitats where sulfur and sulfur compounds (Bak and Cypionka, 1987), the organic carbon loading are high (e.g., sewage use of HS~ or S2Of~ as electron donor in certain effluents). Some PSB and PNSB can use organic cyanobacteria (e.g., De Wit and Van Gemerden, sulfur compounds, either as electron donors 1987; Pringault and GarciaPichel, 2000), and the (Visscher and Taylor, 1993c; Visscher and Van coexistence of CSB and SRB under atmospheric Gemerden, 1991) or as electron donors and/or 02 concentrations (Van den Ende and Van carbon sources (Chien et al., 1999). Gemerden, 1993). The need for an improved understanding of ecophysiological functioning of microbes and their distribution as well as biogeo- 8.08.7.9.5 Ecological aspects ofsulfide chemical transformations in the sulfur cycle are oxidation still unambiguous despite 125 years of research. As the phylogeny, physiology, and biochem- istry of sulfide-oxidizing microbes is under revision, the ecological aspects and biogeochem- ical role of these organisms in carbon, sulfur, and 8.08.8 COUPLED ANAEROBIC ELEMENT nitrogen cycles is anticipated to change as well. CYCLES For example, as outlined above, the capability to oxidize sulfide or thiosulfate is not one of the Although it is pragmatic to consider the element "standard tests" that is performed when character- cycles independently, an understanding of how izing organisms. This may well explain the elements and organisms interact is required to scattered distribution of single-celled CSB in the interpret the relative contributions of the various phylogenetic trees (although the sequence of 16S pathways to anaerobic metabolism in nature. This rRNA and metabolic traits are not synonymous). concluding section elaborates on these inter- In addition to the abundance of sulfur-oxidizers, it actions and briefly reviews the relative contri- is important to acknowledge the environmental bution of the major anaerobic pathways to carbon characteristics, especially those related to 02 and metabolism in ecosystems. HS (and other sulfur-compounds; J0rgensen et al, 1979; Luther et al, 1991 ; Visscher and Van den Ende, 1994; Visscher and Van Gemer- 8.08.8.1 Evidence of Competitive Interactions den, 1993; Zopfi et al, 2001). The expanding role of cultured organisms that The outcome of competition between microor- were not considered to be involved in sulfur ganisms for electron donors can be predicted from transformations and some novel insights in thermodynamic theory (Section 8.08.1.2; Table 1; associations of various sulfur bacteria were Zehnder and Stumm, 1988), and these predictions recently reviewed (Overmann and van Gemerden, are generally consistent with empirical data. Tem- 2000). One example of interactions among poral succession of the microbial metabolic path- sulfide-oxidizers is that of CSB and PSB. When ways that dominate respiration occurs upon the PSB (Thiocapsa roseopersina) and CSB (Thio- flooding of an oxidized soil or sediment (Figure 1). bacillus thioparus) compete for sulfide under Not surprisingly, most examples of temporal alternating oxic-anoxic conditions, PSB out- succession in anaerobic respiration processes compete CSB despite a better affinity for sulfides have come from wetland soils, which are subject in the latter group of organisms (Visscher et al, to cycles of flooding and exposure (Turner and 1992). The storage of S° intracellularly and the Patrick, 1968; Ponnamperuma, 1972; Achtnich capability to use light provide a competitive et al, 1995a; Yao et al, 1999). However, the same advantage. Similarly, other investigators hinted pattern is observed in sediments and even upland that a combination of environmental factors, soils (Peters and Conrad, 1996). including light quality and quantity and sulfide Yao et al. (1999) recognized three phases of and oxygen concentrations, need to be considered reduction following the flooding of rice paddy soil. when evaluating the competitive position In the reduction phase (phase I), the inorganic between CSB and PSB (J0rgensen and Des electron acceptors were sequentially consumed Marais, 1986). In further investigations, where and C02 emissions were highest; in phase II, CH4 the oxygen/sulfide ratios were varied, the PSB production became dominant and peaked, and in 390 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes phase III there was a steady-state ratio of CH4 and C02. Peak CH4 production during phase II was highly correlated (r2 = 0.79) with the ratio of • T nitrogen to inorganic electron acceptors, reflecting the dual influence of carbon quality and compe- tition on methanogenesis. Methane production is favored by a high nitrogen content, because it is indicative of a large labile carbon pool (Section 8.08.4.3.4), which is consumed preferentially by I nonmethanogenic bacteria until the inorganic electron acceptor pools are exhausted. Thus, electron flow through methanogenesis is highest Methane Sulfate Fe(III) Nitrate or Mn(IV) production reduction reduction reduction when the labile carbon pool is large and the inorganic electron acceptor pools are small. The Figure 27 Steady-state H2 concentrations in sedi- length of phase I increases with the inorganic ments with different dominant terminal electron electron acceptor pool size, and decreases with a accepting processes (Lovely and Goodwin, 1988) decrease in the labile pool size or an increase in (reproduced by permission of Elsevier from Geochim. temperature (van Hulzen et al., 1999). Cosmochim. Acta. 1988, 52, 2993-3003). Spatial zonation develops due the progressive consumption of terminal electron acceptors from their source to points downgradient. Such zonation The link between and H offers occurs as a function of depth in systems that are 2 continually inundated such as aquatic sediments. It a nondestructive method of determining the also occurs downstream of an organic carbon point dominant terminal electron accepting process source in rivers and aquifers (Lovley et al., 1994a). in situ (Hoehler et al., 1998). Such an approach A final source of observations that indicate is most effective when the influence of tempera- competition for electron donors comes from direct ture, pH, mineral reactivity and other factors are manipulation of soils or sediments. Addition of taken into account by calculating the free energy of the system. Hoehler et al. (1998) observed that 30 g ferrihydrite per kg soil reduced CH4 emission from a paddy soil by 84% (Jackel and Schnell, the AG value for a given terminal respiration 2000), and similar responses have been reported process remained constant when temperature and following a fall in the water table (Kriiger et al., pH were manipulated, and suggested that it 2001). Freeman et al. (1994) attributed a decline in reflected the need for microorganisms to operate CH4 emissions following drought to inhibition at their energetic limit in order to successfully by S04 reduction, which is consistent with the compete for H2. The theoretical minimum results of direct S04~ addition studies (Gauci et al., energy yield that can support life (i.e., the 2002). biological energy quantum) is one-third ATP per round of metabolism or about -20kJmol_1 (Schink, 1997). This limit agrees at least roughly with in situ observations. The low energy yields 8.08.8.2 Mechanisms of Competition of fermentation, methanogenesis, and most other forms of anaerobic metabolism suggest that Molecular hydrogen (H2) is the most abundant product of fermentation and the most common anaerobic organisms function at a level of near- electron donor in terminal anaerobic metabolism. starvation (Valentine, 2001). Thermodynamic considerations suggest that Extremely low (nanamolar level) H2 concen- trations in anaerobic sediments are evidence of competition for acetate should also favor the metabolic pathway with the highest free-energy keen competition for H2. Metabolic pathways that yield relatively large amounts of free energy (i.e., yield. However, acetate concentrations have been those with more negative AG values) tend to be reported to change in response to a shift in the dominant metabolic pathway in some cases but associated with low pore-water H2 concentrations (Table 1; Figure 27). Thus, S04"-reducers inhibit not others (Achtnich et al., 1995b; Chidthaisong H2-dependent methanogens by reducing the H2 and Conrad, 2000; Sigren et al, 1997). concentration to a value below the threshold at Thermodynamic control of acetate concentrations which CH4 production is thermodynamically may be superseded by kinetic effects due to its feasible (Lovley et al., 1982); Fe(III) reducers slow diffusion rate. have the same influence on S04~ reducers, and so A more recent addition to the list of terminal forth according to the free-energy yield of the electron acceptors is humic substances (Section processes when operating near chemical equili- 8.08.6.4.1). Cervantes et al. (2000) demonstrated brium (Lovley and Goodwin, 1988; Postma and that a humic acid analogue (AQDS) inhibited Jakobsen, 1996). methanogenesis due to a combination of toxic and Coupled Anaerobic Element Cycles 391 competitive effects. The thermodynamic yield of accepting processes. It is well established that AQDS reduction was estimated to be more denitrifying bacteria are active in anaerobic favorable than SO|~ reduction or methanogen- microsites imbedded in upland soils (Section esis, but less favorable than Fe(III) reduction. 8.08.5.3.2; Figure 14), and there is evidence of Because humic substances can quickly transfer methanogenic activity, the most 02 sensitive of electrons to other terminal acceptors such as anaerobic metabolisms, in upland soils as well Fe(III), they are rapidly recycled and could be (von Fischer and Hedin, 2002). Redox potentials important in anaerobic metabolism despite typi- are notoriously variable in anaerobic sediments, cally low concentrations. suggesting that redox microsites occur in the total absence of 02. Spatial variability has been proposed to explain deviations from the expected 8.08.8.3 Exceptions suppression of SO4" reduction by Fe(III) There are many instances when the succession reduction (Hoehler et al., 1998). It is now possible of anaerobic microbial pathways or segregation in to observe such small-scale variability in space varies from the classical pattern (Section microbial populations using molecular techniques 8.08.8.1). There is a tendency for overlap between (e.g., Boetius et al, 2000b; Orphan et al, 2001b) pairs of terminal electron acceptors with similar and element-specific microelectrodes, including a free-energy yields (e.g., NO^~ and Mn; Fe and recently developed system for measuring iron SC>4_). These exceptions can often be explained speciation (e.g., Luther et al., 2001). by the absence of competition for electron donors (Oremland et al., 1982; Crill and Martens, 1986; Holmer and Kristensen, 1994; Chidthaisong and 8.08.8.4 Noncompetitive Interactions Conrad, 2000; Mitterer et al, 2001; McGuire In many cases, the segregation of terminal et al., 2002). Small amounts of CH4 are produced electron accepting processes is due to factors in rice paddy soils immediately upon the onset of other than competition. The ability of NO^ to anoxia, despite the presence of NO4 , and a high redox potential (Roy et al., methanogenesis has been shown to be due in part to 1997;Yao and Conrad, 1999). The initial burst of inhibition by denitrification intermediates (Kliiber methanogenic activity coincides with a peak in H2 and Conrad, 1998, p. 331; Roy and Conrad, 1999). concentrations, suggesting that H2 production Concentrations of < 100 pJVI NO^, 1-2 pJVI NO from fermentation can exceed demand for a and < 1 mM N20 have been reported to completely period of time (Yao and Conrad, 1999). Once inhibit hydrogenotrophic methanogenesis (Balder- Fe(III) reduction and SO4" reduction draw down ston and Payne, 1976; Kliiber and Conrad, 1998), H2 to a level below the threshold required for while somewhat higher concentrations are necess- methanogenesis, CH4 production ceases. A see- ary to inhibit acetogenic methanogenesis (Clarens mingly uncommon reason for the absence of et al, 1998). Chidthaisong and Conrad (2000) competition between terminal electron accepting reported that NO^ amendments inhibited glucose processes is that the organisms require different turnover in a paddy soil, which indicates that electron donor substrates (Section 8.08.7.5.3). nitrogen oxide toxicity may have affected The order of competing terminal electron the methanogens indirectly by inhibiting accepting processes can vary with any number fermentation. of factors that influence the thermodynamics of the system. One factor that must be considered in ecosystems with mineral sediments or soils is the 8.08.8.5 Contributions to Carbon Metabolism composition of Fe(III) and Mn(IV) minerals. The typical sequence of Fe(III) reduction before SO4 The factors that influence the flux of energy reduction can be reversed with a change in the through aerobic-anaerobic interface ecosystems abundance of labile Fe(III) minerals such as have been the focus of this review chapter. Here ferrihydrite (Postma and Jakobsen, 1996). This is we consider the net effect of these influences on one explanation for the common observation that carbon metabolism in marine and freshwater the zones of Fe(III) reduction and SO|~ reduction ecosystems. Thamdrup (2000) recently compiled overlap in marine sediments (Boesen and Postma, studies that reported the relative contributions of 1988; Canheld, 1989; Canfield et al, 1993b; 02 reduction, Fe(III) reduction, and SO4" Goldhaber et al., 1977; Jakobsen and Postma, reduction to carbon metabolism in marine eco- 1994). Postma and Jakobsen (1996) predicted that systems {n = 16). On average, the dominant the overlap between Fe(III) reduction and SO^- pathway was SO4" reduction (62 ± 17%, reduction should increase as Fe(III) oxide stability x ± SO). Aerobic respiration and Fe(III) respir- (or surface area) increases. ation contributed equally to carbon metabolism Spatial heterogeneity is a likely explanation for (18 ± 10% and 17 ± 15%, respectively). Compared the coexistence of competing terminal electron to previous compilations, —50% of the amount 392 Anaerobic Metabolism: Linkages to Trace Gases and Aerobic Processes attributed to 02 reduction is now credited to unexpected types of metabolism, such as nitrify- Fe(III) reduction, while the contribution of S04 ing bacteria that are capable of denitrification and reduction is unchanged (Thamdrup, 2000). Fe(III) Fe(III)-reducing bacteria that produce CH4. The reduction can also be the dominant carbon reduction of Fe(III) was thought to be primarily an oxidation pathway in salt marsh sediments abiotic process, but it is now understood to (Kostka et al, 2002c; Gribsholt et al., 2003). account for much of the anaerobic carbon There have been remarkably few attempts to metabolism in many freshwater ecosystems. determine the contribution of Fe(III) reduction to Microorganisms appear to play a larger role in anaerobic carbon metabolism in freshwater eco- Fe(II) oxidation at circumneutral pH that pre- systems where it may be the dominant pathway. viously believed. Progress has been aided by the The most extensive such study examined 16 rice development of molecular and stable isotope paddy soils collected from China, Italy, and the techniques that yield detailed descriptions of Philippines (Yao et al., 1999). Fe(III) reduction microbial communities. Such techniques have was 58-79% of carbon metabolism during the recently provided strong support for the hypoth- reduction phase (Section 8.08.8.1), with most of esis that anaerobic CH4 oxidation is achieved by a the remainder attributed to methanogenesis. syntrophic relationship between Archea and Fe(III) reduction contributed —70% of the anaero- Bacteria. The significance of these new microor- bic metabolism in a Juncus effusus marsh (Roden ganisms and metabolic pathways for element and Wetzel, 1996). cycling in situ remains to be determined, but it There is growing evidence that plant activity is likely to lead to important revisions of the causes carbon metabolism to shift away from element budgets for carbon, nitrogen, manganese, methanogenesis to Fe(III) reduction. Roden and iron, and sulfur, and the mechanisms that regulate Wetzel (1996) reported that the contribution of their cycling. The recent downward revision of methanogenesis to anaerobic carbon metabolism aerobic carbon mineralization rates in marine shifted from 69% in the absence of plants to <30% sediments that followed the discovery of respirat- in their presence. Similar observations have been ory Fe(III) reducing bacteria (Section 8.08.8.4) is reported for freshwater marshes dominated by a poignant example of how important the field of Scirpus lacustris and Phragmites australis (van environmental microbiology has become for der Nat and Middelburg, 1998) and a rice paddy advancing our knowledge of biogeochemistry. soil (Frenzel et al., 1999). The mechanism for this Likewise, geochemistry has made significant effect is an abundance of poorly crystalline Fe(III) contributions to environmental microbiology by in the rhizosphere compared to the bulk soil discovering the activity of anaerobic CH4 oxidiz- (Kostka and Luther 1995; Weiss, 2002), and ing bacteria and the anammox process. More perhaps a larger labile organic carbon supply. exciting discoveries can be expected as these Fe(III) and Mn(IV) respiration are subordinate fields continue to become integrated. terminal electron-accepting processes in saturated soils and sediments in the absence of a mechanism for their regeneration to oxides. Plants, bioturba- tion, or physical mixing have this effect. In salt ACKNOWLEDGMENTS marshes, bioturbation by crabs may be more The authors gratefully acknowledge the people important to Fe(III) regeneration than radial 0 2 who reviewed portions of the manuscript: Dave loss from roots in some cases (Kostka et al., Emerson, Joel Kostka, Scott Neubauer, Robin 2002c), but not in others (Gribsholt et al, 2003). Sutka, Bo Thamdrup, and Johanna Weiss. 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