responses of atmospheric methane consumption by soils to global climate change

Global Change Biology (1997) 3, 351–362

Responses of atmospheric methane consumption by soilsto global climate change

G A R Y M . K I N GDarling Marine Centre, University of Maine, Walpole, ME 04573, USA

Abstract

Soils consume about 40 Tg methane from the atmosphere annually. Thus, soils contributesignificantly to the atmospheric methane budget. However, responses of atmosphericmethane consumption to climate change are uncertain. Predicting these responsesrequires an understanding of the effect on methane consumption of specific variables(temperature and soil water content) as well as interactions among parameters (methane,ammonium, water content). Key considerations involve the limitations of diffusivetransport and controls of methane diffusivity; limitation of methanotrophic activity bywater stress; relatively slow growth rates of methane-oxidizing bacteria on atmosphericmethane; ammonium toxicity. Interactions among these parameters may be particularlyimportant, and lead to responses contrary to those predicted from changes in temperatureand water content alone. Results from a number of analyses indicate that atmosphericmethane consumption is especially sensitive to anthropogenic disturbances, whichtypically decrease activity. Continued increases in wet and dry ammonium depositionare likely to exacerbate inhibition resulting from changes in land use. Changes inhydrological regimes could further decrease activity if dry periods increase water stressat soil depths currently colonized by methanotrophs. Future trends in the soil methanesink are likely to lead to enhanced accumulation of atmospheric methane.

Keywords: ammonium, atmosphere, forest, land use, methane, methanotroph, soil, water potential

Introduction

Biological methane oxidation is widely distributed in ability and temperature, all of which can change markedlyterrestrial, marine, lacustrine and wetland ecosystems both among and within systems (King 1992, 1993;(King 1992; Conrad 1996; Hanson & Hanson 1996; Hoehler Mancinelli 1995; Hanson & Hanson 1996; King 1996b).& Alperin 1996). The process occurs under both oxic Other biotic controls may include competition (Grahamand anoxic conditions, and over a range of methane et al. 1993) and bactivory. While the responses of bothconcentrations from sub-nanomolar to millimolar (King cultures and in situ methane oxidation to some of these1996a). The microbiology and controls of methane oxida- parameters are known from manipulative experiments,tion are similarly varied. The relevant microorganisms and changes in some of these parameters can be predictedinclude freshwater and soil populations similar to with modest certainty at regional scales, reliable predic-methanotrophs extant in cultures (Hanson & Hanson tions of either the relative or absolute extent of future1996); uncharacterized, presumably novel methanotrophs methanotrophic activity are not yet feasible.in soils (Bender & Conrad 1992; King 1992); largely Limitations for predicting the response of methaneuncharacterized methanotrophs and perhaps ammonia- oxidation to global-scale changes in climate, atmosphericoxidizing bacteria in the marine water column (Sieburth composition, eutrophication and land use do not simplyet al. 1987; Ward 1987; Lees et al. 1991); uncharacterized derive from a lack of understanding of the controlsand novel sulfate-reducing bacteria or consortia of sulfate of methane oxidation. In a more fundamental sense,reducers and methanogens that are active in anaerobic estimates of the contemporary magnitude of global meth-methane oxidation (Hoehler & Alperin 1996). Controls

ane oxidation remain uncertain (Reeburgh et al. 1993).of the activity of these varied organisms include methane,

The integrated rate of atmospheric methane consumptionoxygen and ammonium concentrations, pH, water avail-

by soils appears constrained to a relatively small rangeof values, as are the rates of methane oxidation in anoxicCorrespondence: fax 11/207–563–3119,

e-mail [email protected] marine sediments and in the water columns of lakes and

© 1997 Blackwell Science Ltd. 351

352 G . M . K I N G

the oceans. However, considerable variability has been reported from the tropics to the tundra, grasslands toforests, agricultural soils to deserts (Keller et al. 1983;reported for methane oxidation associated with rooted

aquatic vegetation and freshwater sediments (Holtzapfel- Whalen & Reeburgh 1990; Whalen et al. 1990; Crill 1991;Mosier et al. 1991; Striegl et al. 1992; Adamsen & KingPschorn et al. 1986; King et al. 1990; Epp & Chanton 1993;

King 1994; Schipper & Reddy 1996; King 1996). Since 1993; Dunfield et al. 1993; Jones & Nedwell 1993;Koschorreck & Conrad 1993; Funk et al. 1994; Dunfieldfreshwater sediments and aquatic plants contribute signi-

ficantly to the total global methane budget (Cicerone & et al. 1995; Poth et al. 1995; Thurlow et al. 1995; Steudleret al. 1996a). Rates in these various systems are compar-Oremland 1988; Reeburgh et al. 1993), uncertainties in

these terms necessarily render budgets imprecise. able, typically falling within a range of about 0.5–2 mgmethane m–2 d–1 (Table 1). Global estimates of methaneWith these limitations as a background, one can exam-

ine the sensitivity of methane oxidation to global change oxidation have been derived from datasets of diversesystems using soil texture as a primary determinant (Dorrby considering oxidation rates in broad categories of

important ecosystems or landscapes. For the purposes et al. 1993). The apparent success of this approach suggeststhat a small number of parameters may contribute toof a comparative ecosystem analysis, aerobic methane

oxidation can be separated conveniently into several much of the observed variability.With few exceptions, the controls of methane oxidationcategories. These include the soils of various grasslands,

tundra, forests and deserts; the oxic surface sediments of appear similar among the various soil types. The depthdistribution of methanotrophs is one of the major proxim-freshwater wetlands and lakes; marine sediments; the

water columns of fresh-water and marine systems. Of ate controls of oxidation rates; in most soils, methano-trophs are localized well below the soil surface (Crillthese, aerobic methane oxidation is quantitatively least

important in lakes and marine systems in general. This 1991; Adamsen & King 1993; Koschorreck & Conrad1993; Bender & Conrad 1994). As a consequence, rates ofis not to suggest that high methane oxidation rates do

not occur in these systems, as they certainly do in some activity are often limited by the diffusive transport ofmethane through the soil gas phase (Born et al. 1990;cases. Rather, the integrated rates for lakes and marine

systems appear relatively small on a global basis Adamsen & King 1993; Castro et al. 1995). Diffusivetransport is, in turn, constrained by soil water content,(Reeburgh et al. 1993). Thus, processes in soils will be

emphasized here. which controls soil gas volume along with soil texture(Whalen et al. 1990; Crill 1991; Adamsen & King 1993;Although soils emit methane infrequently by several

mechanisms (Hao et al. 1988; Sextone & Mains 1990; Castro et al. 1995; Schnell & King 1996). To a first approxi-mation then, the flux of atmospheric methane into soilHackstein & Stumm 1994; Wang & Bettany 1995; Steudler

et al. 1996), they represent the only net biospheric sink for can be described in terms of Fick’s first law, J 5 –DsdC/dx, where dx represents depth from the soil surface to theatmospheric methane with a strength of about 40 Tg y–1

(Reeburgh et al. 1993). This sink accounts for about 7– zone of maximal activity (where methane concentrationsare minimal); dC 5 the concentration gradient, 1.7 ppm10% of the total annual global emission of methane, and

is roughly equal to the annual increase in the atmospheric to about 0.1–0.2 ppm; and Ds is the effective gas phasediffusivity, accounting for the effects of both tortuositymethane burden (Watson et al. 1990). Atmospheric meth-

ane consumption by soils occurs widely, and has been within the soil as well as water content.

Table 1. Selected rates of atmosphericmethane consumption by soils. Methodsinclude FC, flux chambers; IC, intactcores; FM, flux models.

Site Method Rate (mg m–2 d–1) Reference

Tundra, Alaska FC 2.7 Whalen & Reeburgh (1990)Tundra, Canada FC 3.3 King et al. (1991)Forest, Boreal FC 0.3–1.6 King et al. (1991)Savannah, Africa FC 1.2 Seiler et al. (1984)Tallgrass prairie FC 0.6–1 Tate & Striegl (1993)Grassland FC 0.1–0.6 Mosier et al. (1991)Temperate forest IC 0.5 Koschorreck & Conrad (1993)Temperate forest IC 2.0 Yavitt et al. (1990)Temperate forest FC 1.6–8 Castro et al. (1995)Tropical forest FC 0.5 Keller et al. (1986)Tropical forest FC 0.7–1.5 Steudler et al. (1996)Mesophytic forest FC 0.8 Keller et al. (1986)Mixed forest FM 0.3–3.5 Born et al. (1990)Mixed forest FC 0–3.5 Crill (1991)

© 1997 Blackwell Science Ltd., Global Change Biology, 3, 351–362

C L I M A T E C H A N G E A N D A T M O S P H E R I C M E T H A N E C O N S U M P T I O N 353

An alternative conceptualization of atmospheric meth- other than diffusivity and water content? What are theimportant interactions among parameters? For whichane consumption can be derived from the observation

that uptake at the soil surface is a first-order kinetic parameters might non-linear responses be expected?Definitive answers to these questions are not yet available.process. This is consistent with the diffusion model above,

and is represented as dC/dt 5 –kC, where dC represents However, the results of a number of studies provide aframework for further analysis.an instantaneous change in methane concentration at the

soil surface. The first-order rate constant, k, for a givensoil or site incorporates a number of factors, including

Diffusion-limited methane uptakeDs, dx, the biomass of methane oxidizers and more. Therate constant can also be viewed as a ‘capacity’ factor Because the zonation of methane oxidation in soils

remains enigmatic, predictions of future trends in dx arethat assumes characteristic values for a given soil or siteas a function of time. conjectural. Several explanations for the sub-surface locus

have been offered but not explicitly tested. These involveUsing diffusion or kinetic models as a focus, severalquestions might be asked. What determines dx, dC and exclusion of methane oxidation from surface soils due to

high ammonium concentrations and rates of nitrogenDs, and how might these parameters change in variousregionally important soils? What physical and biological mineralization; bactivory in excess of methanotrophic

growth rates; water stress resulting from episodic orfactors determine the capacity for methane consumptionsustained drying of surface soils (King 1993; King &Schnell 1994a; Schnell & King 1994, 1996). Each of thesefactors may contribute to varying degrees in differentsystems.

Changes in nitrogen mineralization and increases inammonium inhibit atmospheric methane consumption(Fig. 1; Steudler et al. 1989; Conrad & Rothfuβ 1991;Mosier et al. 1991; Nesbit & Breitenbeck 1992; Hutschet al. 1993, 1994, 1996; Castro et al. 1994; King & Schnell1994a; Neff et al. 1994; Steudler et al. 1996b). In general, theshort-term response to ammonium fertilization involvesinhibition of activity throughout the active methano-trophic zone rather than displacement of the zone deeperinto the soil profile (dx does not change). This effect islargely due to the direct toxicity of ammonium in thepresence of low methane concentrations (Bedard &Knowles 1989; King & Schnell 1994b). Chronic, long-termincreases in ammonium deposition on soils, such ashave been attributed to anthropogenic eutrophication atregional to global scales (Matthews 1994), could elicit asomewhat different response, perhaps involving both adecrease in activity at specific depths as well as a down-ward displacement (increase in dx) of the methanotrophiczone. This might be especially evident in soils for whichthe capacity for ammonium oxidation or adsorption atthe surface is low, thereby allowing transport of ammo-nium into the soil profile. Long-term responses mightalso involve shifts in the types of bacteria that oxidizemethane (Castro et al. 1994; Steudler et al. 1996b); suchshifts could have an impact on dx. For instance, anincrease in the importance of ammonia-oxidizing bacteriaaccompanying chronic nitrogen input could result in a

Fig. 1 Effect of methane concentration on ammonium inhibition smaller dx if the ammonia oxidizers were primarilyof methane oxidation. Open symbols represent in situ methane

located at the soil surface (Castro et al. 1994). This woulduptake in unfertilized forest soils; closed symbols represent ratesnot necessarily lead to increased methane oxidationin soil plots fertilized with ammonium chloride. Bars representthough, since ammonia oxidizers have lower intrinsicinhibition of methane uptake in fertilized plots relative to

controls. See King & Schnell (1994a) for details. activities than methanotrophs (Bedard & Knowles 1989).


354 G . M . K I N G

Although bactivory by protozoans appears to be animportant determinant of microbial populations in soils(Clarholm 1984), there are no data relating to methane-oxidizers specifically. However, a role for predation canbe inferred from the fact that growth rates for methane-oxidizing bacteria in soils appear relatively slow; calcula-tions based on assumed cell densities, cell carbon contentand methane uptake rates suggest methanotroph doub-ling times . 10 d. In contrast, results from agriculturalsoils suggest that protozoan predation alone may requirereplacement times of 4–20 d (Clarholm 1984); predationby nematodes, arthropods and various annelids wouldfurther reduce the replacement times necessary for main-taining biomass. Thus, predation in surface soils may bea significant control of dx. Of course, this argumentassumes that rates of bactivory diminish sufficientlywith depth to allow the development of sub-surfacemethanotrophic populations. While dimunition of bacti-vory with depth is plausible, the general relationshipbetween bactivory and methane consumption needs fur-ther analysis. Nonetheless, one can suggest that increases(or decreases) in bactivore activity could alter atmosphericmethane consumption by altering dx. Changes in theactivity of bactivores might occur in response to changesin a number of parameters, including soil water regimes,temperature, and litter input. For example, decreasedwater content could restrict the activity of many bactiv- Fig. 2 Top panel: water potential vs. water content in soilsores, perhaps promoting colonization of surface soils by from the Darling Marine Centre forest. Lower panel: methanemethanotrophs. consumption at atmospheric concentrations (closed symbols)

and 200 ppm (open symbols) for sieved soils incubated atThe effect of water stress on soil methanotrophy hasvarious water potentials established by drying or wetting freshlikewise received little attention until recently (Schnell &soils as necessary (from Schnell & King 1996).King 1996). Several studies have indicated that extreme

drying and the attendant water stress inhibits methaneconsumption (Whalen et al. 1990; Schnell & King 1996); events with values , – 4 MPa not uncommon. King

(unpublished data) has also shown that the ability ofit has also been suggested that increased gas transportaccompanying soil drying in a future, warmer climate methanotrophs to recover from water stress is very

limited. As a result, periodic drying of surface soils couldwould contribute to greater methanotrophic activity(Whalen & Reeburgh 1990; Whalen et al. 1990). The extent be a major factor constraining methanotrophs to sub-

surface horizons where the temporal excursions in waterto which this is true obviously depends on the waterstatus of a given soil at present: methane consumption potential appear less severe (Schnell & King 1996). This

would mean that average water contents (or potentials)in soils at or near an optimum water content woulddecrease with significant drying in the future. in surface soils might be less important than temporal

extremes for predicting methanotrophic activity. In aThe relationship between water stress and dx deservesparticular attention in the context of future changes. warmer, drier climate one might predict that the zone of

surface soils unavailable for methanotrophic colonizationSchnell & King (1996) have shown that decreasing watercontents stimulate methane consumption until the would increase, even if the average water regime was

not particularly stressful. This would have the net effectlower water contents result in water potentials of about– 0.5 MPa (Fig. 2; see Brown 1990 for a discussion of of increasing dx, possibly leading to net decreases in

methane oxidation in spite of increased gas transport.water potential). As water contents and potentials furtherdecrease, methanotrophic activity is substantially inhib- Rates of atmospheric methane consumption by soils

can obviously be affected by changes in dC as well as dx.ited in a manner comparable to that for ammonia-oxidizing bacteria in soil (Stark & Firestone 1995). Water dC is simply CAtm – CThresh, where the first term represents

the methane concentration at the soil surface, and thepotentials much lower than – 0.5 MPa occur periodicallyin surface soils during intervals between precipitation second represents the threshold concentration at which



methane oxidation ceases. CAtm is presently increasing at coefficent for methane in air is about 0.2 cm–2 s–1, it isonly about 1.5 3 10–5 cm–2 s–1 in aqueous solutiona rate equivalent to about 1% y–1 (Watson et al. 1990),

although changes in atmospheric chemistry related to the (Cussler 1984). Thus, Ds in soils is determined to a largedegree by water content, or more aptly, the percentageMt. Pinatubo eruption (Dlugokencky et al. 1996) or coal

and gas production (Law & Nesbit 1996) temporarilly of water-filled pore space. Soil texture also contributesby determining soil tortuosity, that is the path length ofslowed the rate of increase. Nevertheless, in the absence

of any other changes (in land use), ∆J should equal ∆CAtm, a diffusing molecule per unit length of the soil profile(Born et al. 1990). However, in the context of globaland the relative significance of the soil methane sink

should remain approximately constant (but see below). climate change, soil texture is likely much less significantthan are soil water regimes, since the former shouldThe direction of change in CThresh, if any, is essentially

unknown. Several studies have shown that Cthresh is change relatively little on decadal to century time scales.Changes in soil water regimes are expected as a con-™ 0.1–0.2 ppm (Bender & Conrad 1992, 1995; Adamsen

& King 1993; Koschorreck & Conrad 1993), but it is not sequence of regional-scale changes in climate (Mitchellet al. 1990). However, it is not just the departure fromclear to what extent CThresh is sensitive to changes in

parameters such as temperature or water stress. Adamsen contemporary mean conditions that is important. Thetiming of changes, particularly on a seasonal basis, must& King (1993) have shown that CThresh may increase as

a consequence of ammonium inhibition, a result which be considered. For example, substantially decreased soilwater contents in mid-summer could increase Ds several-suggests that Tthresh should be sensitive to changes in

methanotroph physiological status in general. However, fold over current values, but have little impact on methaneconsumption if such increases are accompanied by watersince CThresh is basically a manifestation of methanotroph

physiology, ∆CThresh is a function of changes in other stress as described above. On the other hand, extendedspring and fall conditions due to a warmer climate withparameters rather than a specific source of changes in

methane consumption. less extreme decreases in soil water content could resultin a much greater stimulation of methane consumption.Changes in the diffusion coefficient of methane in soils,

Ds, can also contribute to changes in methane oxidation. Thus, predictions of the response of atmospheric methaneconsumption by soils to changes in the hydrologic cycleOf course, the heterogenous nature of soils means that

Ds varies among sites (soil types) and that it is not require a high degree of resolution for both spatial andtemporal trends in precipitation.constant over time or with soil depth. However, site-

specific diffusion models have proven successful usingdiffusion coefficients that integrate over vertical variabil-

Controls of methane uptake capacityity (Born et al. 1990; Koschorreck & Conrad 1993). Thus,it is reasonable to focus here on temporal changes in Ds. While it is possible to draw various inferences about

changes in atmospheric methane consumption basedThe magnitude of changes in Ds depends on physicalparameters, specifically temperature and water content. on the preceding analysis, interactions among certain

parameters and non-linearities in physiological responsesDiffusion coefficients respond to temperature accordingto T1.5 to T2, depending on the specific model used to them substantially confound one’s ability to offer

reliable predictions. These interactions can be considered(Cussler 1984). This means that a 5 °C change in temper-ature from 288 to 293 °K would result in only about a in terms of the ‘capacity’ factor, k. For example, if the

supply of methane is not diffusion-limited, the effect of3% change in Ds. Even larger temperature changes thatmight occur during a fraction of the year (summer, for temperature on methane oxidation might take the form

of the Arrhenius model: A 5 A0e–∆H‡/RT, where A 5 ainstance) would have relatively small effects on Ds andthus rates. Accordingly, Q10 values (change in activity rate constant or rate; R 5 the gas law constant; T 5

temperature (°K) and ∆H‡ 5 an activation energy.per 10 °C) for diffusion-limited methane consumption insitu and in intact soil cores in vitro have been relatively Responses to temperature that accord with this model

often result in Q10 values ù 2, which are much largerlow, typically about 1.1–1.5 (Born et al. 1990; Crill 1991;King & Adamsen 1992; Castro et al. 1995; Czepiel et al. than those for diffusion-limited systems. For many soils,

the sub-surface localization of methanotrophic activity1995). Thus, it is unlikely that direct effects of the predicted2–5 °C temperature changes over the next 50–100 year precludes such a response. However, should the depth of

activity decrease and CAtm continue to increase, diffusion-would have a significant impact on methane oxidationas manifest through ∆Ds. In contrast, indirect effects of limitation may become less important, resulting in a shift

in the nature of temperature sensitivity. Certain tundratemperature (resulting in changes in water regimes forinstance) could have much greater impact. soils may already respond according to the Arrhenius

model since maximal atmospheric methane consumptionChanges in water regimes can effect Ds more pro-foundly than temperature changes. While the diffusion has been reported at the soil surface, not at depth (Whalen


356 G . M . K I N G

& Reeburgh 1990). In this case, elevated temperatures to a positive feedback on atmospheric methane accumu-lation.would effectively increase methane consumption by

increasing k (kfuture . kpresent). The response of such soils It is also important to note that the response of methaneconsumption to changes in water content might differto temperature change could thus differ markedly from

that of temperate soils, although future changes in precip- from that based on a consideration of Ds alone. A decreasein water content (increase in gas-filled pore space) coulditation and ammonium deposition could alter the depth

distribution of tundra methanotrophy. have the net effect of increasing methane concentrationsin the soil gas phase. Based on the preceding analysis,A more general illustration of changes in k can be

derived from considering the responses of soil methano- this could decrease k due to increased toxicity of theambient ammonium. Again, while absolute rates of meth-trophy to changes in both ammonium and methane

concentrations. In soils and cultures of methanotrophic ane consumption might increase, the relative changecould be less than predicted from ∆Ds.bacteria, rates of methane consumption increase as meth-

ane concentrations rise from atmospheric levels to about Changes in k can also be anticipated for other anthropo-genic disturbances. Based on in vitro and in situ analyses, it1000 ppm (or greater), where the changes in rate can

be described by a Michaelis–Menten function (v 5 appears that acidification of soils resulting from sulphuricand nitric acids in rainfall and snow can decrease methaneVmax[CH4]/(Km 1 [CH4]), or a variant thereof. Assuming

no growth over the preceding concentration range during consumption (Dunfield et al. 1993; Hutsch et al. 1994;Sitaula et al. 1995; Benstead & King, in press) The mechan-a rate assay, and that no other parameter determines

activity, the methane consumption capacity of the soils ism likely involves both ammonium desorption by H1

input as well as physiological effects of lowered pH.or cultures in question remains constant; net growthduring the assay obviously increases k. Incubation of Acidification of soils due to elevated pCO2 in the soil

atmosphere could have a similar effect. Elevated soilsoils or methanotrophic cultures with increasing levelsof ammonium at constant but low concentrations of pCO2 might occur as a consequence of the change in

the soil–atmosphere CO2 gradient due to rising CO2methane (, 1000 ppm) results in a decrease in k due tothe several toxic effects of ammonium. More importantly, concentrations. The extent and significance of such an

effect is unknown at present.k decreases in soils and methanotrophic cultures withincreasing methane concentrations (over a range from More generally, changes in soil cations alter the distri-

bution of ammonium between adsorbed and soil wateratmospheric to a few hundred ppm) at a given ammoniumconcentration. Stated formally: phases. Cation inputs in particular desorb ammonium

and decrease the capacity for methane consumption(δk/δCH4)NH41

, CH4 , 200 ppm , Ø (King 1996a). The extent of desorption and decreasedconsumption increases from Li1 to Cs1 as expected from

(King & Schnell 1994a,b; Schnell & King 1994). This the behaviour of the lyotropic ions; among the Group IIametals, Mg21 is more potent in desorption than Ca21phenomenon contradicts the expectations of simple

models of methane and ammonium competition (as noted (King 1996a). Interestingly, the common counterionsof these alkaline metals and earths, that is Cl–, NO3

–,in Bedard & Knowles 1989), and is explained instead bythe characteristics of one of the key enzymes in methane SO4

2–, and PO43–, also affect ammonium distribution and

the capacity for methane consumption. The mechanismoxidation, methane monoxygenase, and by the multiplesites for the toxicity of ammonium and the by-products likely involves the formation of ion pairs that alter

desorption (KCl desorbs more ammonium than KNO3)of ammonium metabolism (King & Schnell 1994a,b).Since atmospheric methane concentrations and ammo- and adsorption (a greater fraction of NH4NO3 is adsorbed

than NH4Cl).nium deposition on soils are both increasing (Matthews1994), one can predict that the future capacity for methane These phenomena account for observations that Group

Ia and IIa metal salts inhibit methane consumption whenconsumption (k) by soils will be less than the presentcapacity (kfuture , kpresent). Note however, that this predic- added to soils, but that different salts of the same metal

(NaCl vs. NaNO3) yield different levels of inhibitiontion does not require a decrease in absolute rates ofmethane consumption, since rates are calculated as the (Adamsen & King 1993; Crill et al. 1994; Kightley et al.

1995). There are two consequences. The first is that saltproduct of k and [CH4]; kfuture[CH4future] . kpresent[CH4pr-

esent] if kfuture/kpresent . [CH4present]/[CH4future]. Assum- additions do not constitute a true experimental controlfor ammonium addition, nor can salt solutions be recom-ing a doubling of atmospheric methane in 50–100 year,

absolute rates of methane uptake would be greater than mended in lieu of a deionized water control as suggestedby Kightley et al. (1995). More apt controls for soluteat present as long as kfuture . 0.5 kpresent; clearly though,

for any kfuture , kpresent the relative significance of soils addition require the addition of osmotically active non-electrolytes. The second consequence is that the agricul-as an atmospheric methane sink will decrease, leading



tural use of mineral fertilizers containing various cations, not exhibit a starvation response per se (Benstead & King,in press), obviating earlier inferences about alternativesuch as K1, Mg21, and Ca21, can contribute to large-

scale decreases in the capacity for methane consumption substrates. However, a variety of exogenous substratesincluding methanol, acetate, formate or glucose have noas does the use of ammonium. At present, the relative

significance of ammonium vs. other ions is unknown. stimulatory effect on soil methane consumption (Schnell& King 1995; Benstead & King, in press), a result thatIn addition to affecting soil acidity and possibly ammo-

nium exchange, elevated pCO2 may have other, as yet suggests an uncertain role for these compounds in soilmethanotrophy. Thus, it seems unlikely that changes inundocumented indirect effects on methane consumption.

Elevated pCO2 enhances the growth of at least some C3 the pool of soil organics will have much of an effect onmethane consumption, and that any positive feedbackplant species (Bazzaz 1990; Rastetter et al. 1991; Keeling

et al. 1996; Luo et al. 1996), although the extent of the involving changing CO2 might be more likely the resultof interactions with ammonium.effect is uncertain (e.g. Melillo et al. 1990; Diaz et al. 1993).

Enhanced plant production might result in depletion of More profound effects on the capacity of soils toconsume atmospheric methane have occurred and willsoil ammonium pools to meet the demand of higher

photosynthetic rates. Ammonium depletion could continue to occur at regional to global scales due tochanges in land use. Currently, about 10% of the terrestrialincrease methane consumption by relieving ammonium

inhibition. The extent to which ammonium depletion surface is used in some form of agriculture (Melillo et al.1990), and this is expected to increase in the futuremight occur in response to increasing pCO2 is uncertain,

as is the extent to which increased nitrogen fixation (Adams et al. 1990). Since forests, woodland/shrublandand savannah tend to have much higher rates of methane(possibly even due to methanotrophic bacteria) would

offset depletion. consumption per unit area than cultivated land (Bornet al. 1990; Hutsch et al. 1994; Ambus & Christensen 1995;Increased primary production accompanying elevated

pCO2 may have yet another indirect effect on the capacity Goulding et al. 1995), conversion of these habitats resultsin a net loss in the capacity for methane consumption.of soils to consume methane. Attempts to account for

the so-called ‘missing’ atmospheric CO2 have led to An approximate upper limit for this loss to date can beestimated by assuming that agricultural lands consumespeculation that pools of terrestrial carbon have increased,

especially in temperate forests and peatlands no methane and that the current sink of 40 Tg y–1 in non-agricultural lands has not been affected by anthropogenic(Kirschbaum 1993; Schindler & Bayley 1993; Keeling

1996). These pools likely include pectins as well as other disturbance. Based on these assumptions the capacity forterrestrial methane consumption has been reduced byplant polymers. Pectins contain methoxylated uronic

acids that can serve as precursors for methanol. Methanol about 4.4 Tg y–1, which is equivalent to 9–11% of therecent rate of increase in the atmospheric methane burdenand other simple organics have been proposed as co-

substrates for soil methane oxidizers. The rationale for (Watson et al. 1990).A number of studies have documented decreases ininvoking non-methane substrates is that atmospheric

methane provides only a marginal source of carbon and methane consumption associated with the conversion offorests and grasslands to various agricultural uses (Mosierenergy. Non-methane substrates may thus be required in

order for soil methane oxidizers to meet the demands of et al. 1991; Nesbit & Breitenbeck 1992 Hutsch et al. 1993;Tate & Striegl 1993; Mosier et al. 1996). Steudler et al.maintenance or growth. To the extent that this argument

proves true, increased plant production and elevated (1996a) not only observed a decrease in methane con-sumption associated with forest-to-pasture conversion inpools of soil organic matter could increase the capacity

for methane consumption. It is also conceiveable that the Amazon Basin, but a change to net methane emissionof a magnitude sufficient to account for 12–14% of theincreased pools of soil organic matter might stimulate

methane production within soils and thus methane avail- change in the total global atmospheric methane contentfrom 1975 to 1988. Clearly a shift from net methanotrophyability, again with the net result of an enhanced capacity

for methane consumption. to net methanogenesis can have profound implications.Losses of methane consumption capacity also appearAlthough the possible secondary effects of CO2 on

methane consumption merit further analysis, existing to occur as a consequence of silviculture (Castro et al.1994; King, unpublished results), even though the areadata raise questions about the role of non-methane sub-

strates (Schnell & King 1995; Benstead & King, in press). in cultivation remains a ‘forest’ by some definitions.Decreases in methane consumption due to land use areKing (1996a,b) has argued that a starvation response

observed in forest soils (Schnell & King 1995; see also typically substantial and persistent (Mosier et al. 1991;Hutsch et al. 1993, 1994), with slow recovery times afterRoslev & King 1994, 1995) precludes an important role

for non-methane substrates. A more detailed study has agricultural use ceases in the tropics (Keller et al. 1993,1994). Similarly slow recoveries have been noted forsubsequently indicated that soil methane oxidizers may


358 G . M . K I N G

Fig. 3 A simple conceptual model ofatmospheric methane consumption bysoils. Model symbols are as follows:u, reservoirs or pools; s, regulatorparameters or flow controls; s, flowsbetween reservoirs. Cross-hatchedsymbols refer to biological processes,pools or controls; open symbols referto primarily physical or chemicalparameters. A dashed arrow denotesan indirect effect. Large boxes indicateparameters or processes external to themethane consumption model thatserve as forcing functions. Parameterssuch as temperature (which affectsdiffusivity and metabolic rates) andcompetition and predation (affectingmethanotroph biomass) are notincorporated into the model tomaintain simplicity.

temperate grasslands (Hutsch et al. 1994; Mosier et al. or zero effects of land use change on methane oxidationappear to be the exception.1996).

However, certain land uses may have little or no One of the primary mechanisms for the observedchanges in methane consumption involves ammoniumnegative effect, or the effects may depend on the specific

details of land use. For example, Sanhueza et al. (1994) fertilization or increases in the abundance of ammonium,even if only temporarily, due to increased turnover ofhave reported no change in methane consumption follow-

ing plowing or fertilization of Venezuelan savannah, soil organic matter. Specific farming practices can also beinhibitory (Hansen et al. 1993), although it is important towhile Poth et al. (1995) have noted that recently burned

Brazilian cerrado consumed atmospheric methane but note that some trace gas exchanges, e.g. carbon monoxideconsumption, are enhanced by physical disturbances ofthat unburned cerrado did not. Tate & Striegl (1993) have

found that burning increased methane consumption by soil (Sanhueza et al. 1994). The response of methaneconsumption to exogenous ammonium has been reason-temperate grasslands, but they also have observed

decreased methane consumption by grassland planted ably predictable in the recent past, but it is evident thatland use and land use history can temper or alter thewith sorghum; wheat culture had no effect. Mosier et al.

(1996) have reported persistent inhibition by ammonium response radically (Hutsch et al. 1994; Flessa et al. 1995;Goldman et al. 1995).fertilization of a sandy loam soil, but no long-term effect

in a finer textured soil. Dunfield et al. (1995) have also Several studies have indicated that ammonium min-eralization provides a better index of inhibition thanobserved no effect of nitrogen additions to an organic rich

humisol; in their soil, ammonium was rapidly nitrified, ammonium concentrations (Mosier et al. 1991; Hutschet al. 1993; Mosier et al. 1996). Aside from the directwhich may have prevented toxicity for methanotrophs.

In addition, their humisol was capable of significant inhibition of methanotrophs as described earlier, or amore general toxic effect (Soderstrom et al. 1983), changesmethane production. As a result the dynamics and con-

trols of methane oxidation in this system may differ from in the capacity for methane consumption may accompanyammonium fertilization as a result of the increased signi-grassland and forest soils. Overall though, these positive



ficance of ammonia-oxidizing bacteria; such a shift in of the areal extent of tundra ecosystems, enhancedmethanotrophy in them could have a significant impactrelative importance has been noted for a number of

fertilized soils (Steudler et al. 1996). Ammonia-oxidizers on the atmospheric methane budget.consume methane, but generally have a lower affinityand lower cell-specific oxidation rates than methane-

Acknowledgementsoxidizing bacteria (Bedard & Knowles 1989). Thus, dis-placement of methanotrophs by ammonia oxidizers may The author acknowledges support from USDA CSRS-CRP 94–

37107–0488 and NSF DEB 9107315. Numerous discussions withexacerbate direct inhibition by ammonium and contributeS. Schnell, P. Roslev and J. Benstead have been most valuable.to the persistence of inhibition.Contribution 304 from the Darling Marine Center.

In contrast, increases in the capacity for atmosphericmethane consumption might accompany some changesin the use or management of wetlands, which are typically Referencesnet sources of methane. For example, decreases in meth-

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responses of atmospheric methane consumption by soils to global climate change

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