regulators of heterotrophic microbial potentials in ... · regulators of heterotrophic microbial...

Regulators of heterotrophic microbial potentials in wetland soils

Elisa M. D'Angelo *, K.R. Reddy

University of Florida Wetland Biogeochemistry Laboratory, Soil and Water Science Department, 106 Newell Hall, P.O. Box 110510, Gainesville,

FL 32611-0510, USA

Accepted 10 November 1998

Abstract

Potential rates of aerobic respiration, denitri®cation, sulfate reduction and methanogenesis were investigated in 10 di�erentwetland soils with a wide range of biogeochemical characteristics, with the objective of determining relationships betweenprocess rates and soil properties. Electron acceptor amendments to methanogenic soils caused gradual (1±13 d) to immediate

transitions in electron ¯ow from methanogenesis to alternate electron acceptors. Rates of organic C mineralization rangedbetween 0.2 and 34 mmol C gÿ1 dÿ1 and averaged three times faster with O2 as compared to alternate electron acceptors. Therewas no signi®cant di�erence between rates of organic C mineralization (CO2+CH4 production) under denitrifying, sulfate-

reducing and methanogenic conditions, indicating that soil organic carbon availability was similar under the di�erent anaerobicconditions. Rates of electron acceptor consumption ranged between 1 and 107 mmol gÿ1 dÿ1 for O2, 0.5 and 9.3 mmol gÿ1 dÿ1

for NO3ÿ , 0.1 and 11.1 mmol gÿ1 dÿ1 for SO4

2ÿ and 0.1 and 6.2 mmol gÿ1 dÿ1 for CO2. Heterotrophic potentials in wetland

soils were strongly correlated with inorganic N and several available C indices (total, dissolved and microbial C), but not withpH or dissolved nutrients (P, Ca2+, Mg2+, Fe(II)). Microbial activity±soil property relationships determined in this study maybe useful for predicting the fate of pollutants that are in¯uenced by microbial oxidation±reduction reactions in di�erent types ofwetland soils. # 1999 Elsevier Science Ltd. All rights reserved.

1. Introduction

Natural and constructed wetlands are recognized ascritical components of the landscape because theyoften function as e�cient transformers of agricultural,industrial and domestic point and non-point dis-charges, thereby improving water quality. Moreover,concerns about the adverse e�ects of eutrophication,greenhouse gas emissions, toxic organics and metals onenvironmental quality have drawn attention to theneed for research on the role of wetlands as sourcesand sinks of these pollutants and factors a�ecting pol-lutant transformations.

Many transformations in wetlands are directlymediated by diverse groups of aerobic and anaerobicmicroorganisms in the soil. Microbial activities also in-directly in¯uence redox potential- and pH-sensitiveprocesses, such as trace metal and phosphorus solubi-lity, precipitation, sorption and hence mobility(D'Angelo and Reddy, 1994; Gambrell, 1994; Olila

and Reddy, 1997; Reddy et al., 1998a,b). Depending

on the soil and other site-speci®c properties, one ob-

serves several orders of magnitude di�erences in rates

of microbial activity. This makes it di�cult to predict

pollutant fate without extensive experimentation on a

site-by-site basis. Development of a database of soils

information including rates of microbial processes and

relatively easily measurable properties may produce re-

lationships that could be used to predict potential

rates of microbial activity in many di�erent types of

soil environments.

A number of investigations have found signi®cant

correlations between microbial respiration rates in soils

and environmental properties, including nutrient avail-

ability (McKinley and Vestal, 1992; Amador and

Jones, 1993; Aerts and Toet, 1997), pH (Bergman et

al., 1998), temperature (Westermann and Ahring,

1987; Bridgham and Richardson, 1992; Prieme, 1994)

and electron donors (Burford and Bremner, 1975;

Yavitt and Lang, 1990; Jorgensen and Richter, 1992;

Crozier et al., 1995). However, results from di�erent

studies often lead to con¯icting conclusions and re-

lationships between microbial process rates and soil

Soil Biology and Biochemistry 31 (1999) 815±830

0038-0717/99/$ - see front matter # 1999 Elsevier Science Ltd. All rights reserved.

PII: S0038-0717(98 )00181 -3

PERGAMON

* Corresponding author. Tel.: +1-352-392-1804; fax: +1-352-392-

3399; e-mail: [email protected]¯.edu

properties are often unpublished. Moreover, many ex-periments have limited scope with respect to the rangeof soil characteristics and metabolic processes studied,so that predictions for soils with characteristics outsideof a narrow range require extrapolation.

Wetland soils are di�erent from most upland oraquatic soils and sediments because they often undergointermittent ¯ooding and draining, thus supportingboth aerobic and anaerobic microbial communitiesthat exploit a wide range of electron acceptors duringrespiration, such as O2, NO3

ÿ , Fe(III), SO42ÿ and CO2

(Sorensen et al., 1979; Lovley and Klug, 1986).Wetlands also contain soils ranging from mineral toorganic, eutrophic to oligotrophic and saline to fresh-water (Reddy et al., 1998a,b). Therefore developmentof soil microbial activity±property relationships forwetlands requires evaluation of aerobic and severalmodes of anaerobic respiration for soils with a widerange of characteristics.

The hypothesis we have tested is that potential ratesof aerobic respiration, denitri®cation, sulfate reductionand methanogenesis in wetland soils are signi®cantlycorrelated to soil properties. Our results provideempirical relationships that predict potential rates ofheterotrophic activity from relatively easily measurablewetland soil characteristics. These relationships may beuseful for predicting rates of microbial transformationsof many chemical species and their fate in the wetlandenvironment.

2. Materials and methods

2.1. Soil collection and incubation

Three mineral and seven organic soils were collectedfrom various wetlands in the continental USA(Table 1), which included soils from freshwater and

estuarine, eutrophic and oligotrophic, organic andmineral, and natural and constructed wetlands.Selection of soils from these wetlands was arbitrary,except that it was desired to obtain soils with a widerange of chemical characteristics so that derived re-lationships would be widely applicable. Soils were col-lected under drained (CR, LAAF, NCB) or ¯oodedconditions (remaining soils) hence, they initially haddi�erent water contents and redox potentials. Surface15-cm sections of soil were collected with a polyvinylchloride (PVC) corer, which was then transferred to a4-l plastic bottle and returned in an ice chest to thelaboratory by overnight mail. When present, surfacewater samples from each wetland were also collected.Soils were sieved (0.5-cm) to remove large plant debris,shells and stones. Soils and water were stored for amaximum of 3 months at 48C before being used in ex-periments.

Aerobic, denitrifying, sulfate-reducing and methano-genic activities were determined from soil slurries,which were prepared from soil and site water or deio-nized water. Slurries were used to avoid di�usion con-straints and development of anaerobic microsites. Bulkdensities of slurries were 0.056±0.448 g cmÿ3 for or-ganic soils and 0.138±0.669 g cmÿ3 for mineral soils,which were chosen to maximize solids content whilemaintaining soil homogeneity.

Slurries were conditioned under anaerobic con-ditions to consume electron acceptors initially presentin the soil. Five-ml slurry was transferred to 27-ml an-aerobic tubes (Bellco Glass, Vineland, NJ), sealed withbutyl stoppers-aluminum crimps (Wheaton, Millville,NJ) and purged with O2-free N2. Anoxic conditionswere con®rmed by gas chromatography. Soil tubeswere incubated horizontally in the dark at 288C withshaking at 180 rpm with 2.5 cm rotation diameter(New Brunswick Incubator Shaker Model G25,Edison, NJ) in order to avoid di�usion constraints and

Table 1

Description of organic and mineral wetland soils used in the study

State Soil name (symbol) Description (dominant vegetation)

Organic

Michigan Houghton Lake Peat (HLPI) Peat soil, impacted by domestic waste discharge (Typha spp.)

Michigan Houghton Lake Peat (HLPU) Peat soil, not a�ected (Carex spp.)

Florida Everglades (W2) Peat soil, a�ected by P enriched agricultural discharge (Typha spp.)

Florida Everglades (W8) Peat soil, not a�ected (Cladium spp.)

Louisiana Salt marsh (LSM) Organic salt marsh sediments (Spartina spp.)

North Carolina Belhaven muck (NCB) Subsided organic agricultural soil

Florida Lake Apopka muck (LAAF) Subsided organic agricultural soil

Mineral

Alabama Talladega (TAL) Freshwater sediment (Juncus spp.)

North Dakota Parnell (PPP) Prairie pothole-silty clay loam

Louisiana Crowley (CR) Paddy soil-silt loam (Oryza sativa)

E.M. D'Angelo, K.R. Reddy / Soil Biology and Biochemistry 31 (1999) 815±830816

development of microsites (Stark, 1996). Tests showedthat rates of methane production were not signi®cantlya�ected by this shaking procedure. All experimentswere conducted in triplicate from the same soil sample.The duration of the conditioning period was deter-mined from the time to reach the maximum reducingcondition, which was indicated from accumulation ofmethane in the headspace of the tubes.

After maximum reducing conditions were attained,soil slurries were amended with di�erent electronacceptors. Electron acceptor treatments includedaerobic (1 ml distilled water+5 ml O2), denitrifying (1ml deoxygenated 80 mM KNO3 with or without 4 mlC2H2), sulfate reducing (1 ml deoxygenated 20 mMK2SO4), Fe(III)-reducing (2 ml deoxygenated 140 mMbioavailable Fe(III) solution) and methanogenic con-trols (1 ml deoxygenated water). For some soils, ad-ditional electron acceptor amendments were requiredduring the incubation to maintain non-limiting concen-trations for zero-order kinetics (Hallberg et al., 1976;Murray et al., 1989; Bak and Pfenning, 1991; Tiedje,1994). Acetylene was prepared by reaction of CaC2

with water. Fe(III) solution was prepared by neutraliz-ing FeCl3 solution with 10 M NaOH, followed by sev-eral water rinses of precipitate to remove excesschloride (Lovley et al., 1991; Ghiorse, 1994). Soils(LSM, NCB and LAAF) with low amounts of bioa-vailable Fe were not evaluated for Fe(III) reduction.

Oxygen consumption (aerobic treatments) and N2Oproduction (denitrifying treatments) were determinedby gas chromatographic analysis of headspace gasesevery 30 min for the initial 3 h, followed by daily andbiweekly measurements. Carbon dioxide and CH4 inthe headspace were measured by gas chromatographyat daily and biweekly intervals. In the sulfate treat-ments, dissolved SO4

2ÿ in soil was measured every 3±5d by ion chromatographic analysis of extracts obtainedby shaking soils with 10 ml 1.7 mM NaHCO3±1.8 mMNa2CO3 solution for 1 h, centrifuging at 400g and ®l-tering with 0.45-mm membrane ®lter. The alkalineextractant was used to prevent sorption of SO4

2ÿ withanion exchange sites of soil particles. In the Fe(III)treatments, attempts were made to measure biotic re-duction by analyzing production of Fe(II) in 1 M HClsoil extracts (Lovley and Phillips, 1987). Due to the ac-cumulation of reduced chemical species during metha-nogenic conditioning, however, it was not possible todi�erentiate biotic and abiotic Fe(III) reduction.Therefore, Fe(III) reduction rates were not measuredin this study. Electron acceptors, CO2 and CH4 weremonitored for a total of 20 d.

2.2. Soil chemical analysis

At the end of the methanogenic conditioning, threetubes were removed for chemical analysis. After cen-

trifugation at 400g for 15 min (Sorvall Model RC5,Wilmington, DW), 0.1 ml supernatant was removedfor Fe(II) determination using ferrozine-hepes color re-agent (Lovley and Phillips, 1987). The pH was immedi-ately determined by immersing an electrode in thesupernatant. Soils were sequentially extracted by shak-ing with 10 ml deoxygenated distilled water and 10 mldeoxygenated 1 M KCl for 1 h each. Individual extractsolutions were separated from solids by centrifugationand ®ltration with 0.45-mm membrane before chemicalanalysis. Water extracts were analyzed for electronacceptors (NO3

ÿ and SO42ÿ), inorganic nutrients

(NH4+ , PO4

3ÿ, Ca2+, Mg2+, Fe) and organic C. TheKCl extracts were analyzed for exchangeable NH4

+

and PO43ÿ. Water and KCl exchangeable NH4

+ andPO4

3ÿ were summed to estimate the available forms forbiological utilization. The amount of H2S at the end ofthe methanogenic incubation was estimated from theamount of SO4

2ÿ produced 20 d after addition of O2 inthe aerobic treatment.

Carbon in living, non-resting microbial biomass wasestimated using the substrate-induced respiration tech-nique, in which aerobic CO2 production rate (ml gÿ1

hÿ1) was determined by gas chromatography everyhour for 5 h after amendment with 1 ml YPD Broth(Difco Laboratories, Detroit, MI) to 5 ml soil slurry at228C (Howarth and Paul, 1994). This was equivalentto the addition of 2 mg glucose, 2 mg peptone, and 1mg yeast extract per ml soil solution. A conversionfactor of 50 was used to estimate microbial biomass C(mg gÿ1) from the respiration rate (Sparling et al.,1990).

2.3. Analytical methods

Total C and N in soils were determined with aCarlo-Erba NA-1500 CNS analyzer (Haak-BuchlerInstruments, Saddlebrook, NJ). Total P was deter-mined by digestion of soils with nitric-perchloric acid(Kuo, 1996) followed by elemental analysis by aninductively coupled argon plasma (Thermo Jarrell AshICAP 61E; Franklin, MA). Dissolved Fe(II) and bioa-vailable Fe (0.25 M hydroxylamine+0.25 M HClreducible Fe) were determined by the ferrozine-hepescolorimetric method and measuring absorbance at 562nm using a Shimadzu UV/VIS spectrophotometer(Columbia, MD) (Lovley and Phillips, 1987). Soil pHwas measured with a semi-micro combination glasselectrode (Orion Sure-¯ow; Fisher Scienti®c,Pittsburgh, PA) and meter (Orion Ionalyzer model407A; Cambridge, MA). Dissolved NH4

+ and solublereactive P were analyzed by the salicylate-nitroprussidetechnique and ascorbic acid technique (USEPA, 1993),respectively, using a Technicon Autoanalyzer II(Terrytown, NY). Dissolved SO4

2ÿ and NO3ÿ concen-

trations were determined using a Dionex 4500i ion

E.M. D'Angelo, K.R. Reddy / Soil Biology and Biochemistry 31 (1999) 815±830 817

chromatograph (Sunnyvale, CA). Dissolved organic Cwas determined using a Dohrman DC 190 carbon ana-lyzer (Santa Clara, CA).

Headspace O2 and CO2 were measured using aShimadzu 8AIT GC equipped with thermal conduc-tivity detector (308C) with He as carrier gas and stain-less steel columns (0.3 cm by 2 m) packed withmolecular sieve 5A for O2 and Poropak N for CO2

(Supelco, Bellefonte, PA), maintained isothermally at308C. Headspace CH4 was measured using a Shimadzu8AIF GC equipped with FID (1108C), with N2 as car-rier gas and a stainless steel Carboxen 1000 column(0.3 cm by 2 m) (Supelco) maintained isothermally at1608C. Headspace N2O was measured using aShimadzu 14A GC equipped with 63Ni ECD (3008C)and stainless steel Poropak Q column (0.3 cm by 2 m)(Supelco), 5 kPa methane in argon as carrier gas andmaintained isothermally at 308C. The total amount ofgases in soil tubes was calculated from the partialpressure in the headspace plus the amount dissolved inthe aqueous phase, using Henry's Law constants (M/atm at 258C) of 1.263�10ÿ3 for O2 (CRC, 1984),34.1�10ÿ3 for CO2 (Butler, 1982; Gale et al., 1992)and 1.49�10ÿ3 for CH4 (Thibodeaux, 1979) and usinga Bunsen absorption coe�cient of 0.544 for N2O(Tiedje, 1994). Dissolved carbonate species were alsoincluded in calculations for total CO2 production,using carbonate equilibrium constants and pH asdescribed by Gale et al. (1992).

2.4. Data analysis

Rates of consumption of electron acceptors and pro-duction of CO2 and CH4 by the slurries were bestdescribed by zero-order kinetics (mmol gÿ1 dÿ1) deter-mined from the slopes of the best-®t regression lineobtained during experiments. Since experiments were

conducted at non-limiting electron acceptor reducingconditions, rates may be considered as maximum vel-ocity rates (Vmax) or potentials. Comparisons amongmean values for di�erent electron acceptors and soiltreatments were made using Fishers least signi®cantdi�erence (LSD) tests performed by the StatGraphicssoftware package (Manugistics, Rockville, MD).Relationships between microbial activities and soilproperties were derived from linear regression analysis.

3. Results and discussion

3.1. Chemical properties of soils

Soils from di�erent wetlands varied widely in chemi-cal composition, including pH, total, dissolved and mi-crobial C, N and P, bioavailable Fe and dissolvedCa2++Mg2+ (Tables 2 and 3). Molar C-to-N and C-to-P ratios ranged between 12±36 and 145±3600, re-spectively. Wide ranges in chemical characteristicslikely re¯ected di�erences in various soil forming fac-tors (e.g. parent material, climate, vegetation andmicrobiota and topography) as well as anthropogenicin¯uences (e.g. nutrient loading). In general, organicsoils had higher nutrient availability than mineral soils,which was especially evident in HLPI and W2 soilsthat received external nutrient loading from domesticand agricultural waste waters, respectively.

3.2. Methane production in wetland soils

After soil ¯ooding, there was a lag phase beforemethane production that lasted 10, 23 and 20 d for in-itially drained CR, LAAF and NCB and <5 d for in-itially ¯ooded soils (Table 4 and Figs. 1, 2 and 3). Lagtimes likely re¯ected initial numbers of methanogens

Table 2

Selected physical and chemical characteristics and microbial biomass carbon of wetland soil slurries used in experiments. Each value represents

the mean of three replications2one standard deviation

Soil Total C

(mmol gÿ1)

Total N

(mmol gÿ1)

Total P

(mmol gÿ1)

Bioavailable Fe

(mmol gÿ1)

Microbial biomass C

(mmol gÿ1)

Dry bulk density

(slurry) (g cmÿ3)

Organic

HLPI 3220.3 1810249 5329 9123 9722 0.07920.002

HLPU 3620.2 1739237 2624 7822 6123 0.11420.001

W2 3620.3 2128239 2621 2.320.1 5325 0.05920.000

W8 3720.1 2055223 1320.6 2.520.1 4728 0.05620.000

LSM 1420.1 830217 2220.6 5.520.4 1922 0.13220.003

NCB 1822 495249 525 3.620.1 1224 0.44820.005

LAAF 921 513294 1522 1.220.1 2325 0.41820.005

Mineral

TAL 720.1 365218 2021 74270 3126 0.13820.002

PPP 420.1 290212 2121 3921 2121 0.66920.036

CR 0.820.0 6723 5.523 5.520.2 1321 0.62620.007


(Raskin et al., 1996), presence of toxic O intermediates(Fetzer et al., 1993), nitrogen oxides (Balderston andPayne, 1976; Kluber and Conrad, 1998) and microbialcompetition for available resources (Achtnich et al.,1995a). However, survival of methanogens in oxicpaddy, savanna and desert soils has been demonstratedby Peters and Conrad (1995) and has been postulatedto be possible by removal of toxic O radicals by super-oxide dismutase produced by methanogens (Kirby etal., 1981), neighboring microbial communities, and theability of stable pyrite (FeS2) granules to provideanoxic microhabitats for methanogens.

Following the lag period, there was an exponentialincrease in methane production that reached maximumsteady-state rates between 0.1 and 6.2 mmol gÿ1 dÿ1

(Table 4). These rates cover the range measured forother systems including wetland soils (Westermannand Ahring, 1987; Mayer and Conrad, 1990; Prieme,1994) and lake sediments (Strayer and Tiedje, 1978).

3.3. Decomposition of soil organic C under di�erentelectron acceptor reducing conditions

After steady-state methane production was attained,soils were amended with di�erent electron acceptorsresulting in signi®cant shifts in modes and rates of or-ganic C decomposition (Figs. 1±3; Tables 4 and 5).For many electron acceptors and soils, there was a lagtime before maximum inhibition of methanogenesiswas observed and inhibition was often incomplete(<90%). For example, in salt marsh sediments a 1 ddelay before onset of denitri®cation corresponded toincomplete (83%) inhibition of methanogenesis(Table 4 and Fig. 1). After this, initiation of denitri®-cation coincided with complete (100%) inhibition of

methanogenesis. In several soils (HLPU, TAL and

CR), Fe(III) only partially inhibited (49±67%) metha-

nogenesis compared to control, demonstrating that sig-

ni®cant rates of methanogenesis may proceed in some

soils in the presence of this electron acceptor (Fig. 2).

Likewise in NCB soils, a 6 d lag before onset of sulfate

reduction corresponded to a 13-d lag before inhibition

of methane production (Fig. 3).

The in¯uence of electron acceptors on methanogen-

esis was likely attributed to many factors.

Methanogenesis is inhibited by toxic amounts of O

compounds (H2O2 and O2ÿ), nitrogen oxides and H2S.

Fetzer and Conrad (1993) determined that O2 concen-

trations as low as 0.6 mM (2 mg lÿ1) inhibited methane

production in pure culture suspensions of

Methanosarcina barkeri. Concentrations of NO3ÿ

between 700 and 1000 mM (Balderston and Payne,

1976; Winfrey and Zeikus, 1979), SO42ÿ between 200

and 2000 mM (Winfrey and Zeikus, 1977; Yavitt and

Lang, 1990) and H2S concentrations between 100 and

20,000 mM (Cappenburg, 1975; Bryant et al., 1977;

Winfrey and Zeikus, 1977) have been found to inhibit

methanogenesis.

Toxicity by electron acceptors may be attributed to

enzyme poisoning at elevated redox potentials, such as

conversion of methanogenic coenzyme F420 from its

active to inactive form (Kiener et al., 1988). Toxicity

by H2S may be caused through reaction with essential

iron containing compounds in the cell (e.g. ferredoxins

and cytochromes) (Okabe et al., 1995).

Methane inhibition may also occur when alternate

microbial groups outcompete methanogens for avail-

able electron donors such as H2 and acetate, due to

di�erences in thermodynamic energy yields and reac-

Table 3

Dissolved nutrient concentrations and pH of wetland soils slurries after a methanogenic conditioning (preincubation) period. Each value rep-

resents the mean of three replications2one standard deviation

Soil Water soluble Water soluble+KCl exchangeable

pH Fe(II) (mmol gÿ1) Ca+ 2+Mg+2 (mmol gÿ1) DOC (mmol gÿ1) NH4+ (mmol gÿ1) PO4

3ÿ (mmol gÿ1)

Organic

HLPI 7.2920.04 1020.4 4022 55211 11123 2.620.2

HLPU 6.0520.08 3.320.6 6.521 115231 1020.3 0.0620.0

W2 7.5320.02 0.420.0 nda 23424 3520.2 4.220.1

W8 7.4920.02 0.420.0 7828 12524 2923 2.720.3

LSM 7.0220.00 0.120.0 72210 81222 1821 0.7320.0

NCB 5.4920.05 1.720.2 1.820.3 4524 7.420.1 0.0920.0

LAAF 7.1920.03 0.0520.0 2328 4221 5.320.3 1.320.8

Mineral

TAL 6.8920.02 1221 3.120.2 7423 2120.4 0.0220.0

PPP 6.1520.12 0.520.0 2.420.6 1920.2 8.520.7 0.5520.1

CR 7.4320.04 0.220.0 1020.9 720.3 3.420.1 0.0220.0

a nd=not determined.


tion kinetics between groups (Achtnich et al., 1995a,b;Kluber and Conrad, 1998).

Lag times before methane inhibition may re¯ect thetime required for denitri®ers and sulfate reducers toincrease in numbers su�ciently to outcompete metha-nogens for electron equivalents (Raskin et al., 1996).Partial inhibition of methanogenesis may result whenmethanogens utilize some types of substrates morereadily than Fe(III) or SO4

2ÿ reducers, including acetate(Achtnich et al., 1995b), methylamines, methionine ormethanol (Oremland and Polcin, 1982) or when elec-tron donors are non-limiting (Yavitt and Lang, 1990).Combinations of these processes likely play roles inin¯uencing microbial activity in these complex commu-nities.

Addition of O2 resulted in immediate increase inCO2 production, with rates ranging between 0.2 and33.5 mmol gÿ1 dÿ1 (Figs. 1±3; Table 5). Turnoverrates for total organic C (calculated from the ratio ofmineralization rate and total C content) rangedbetween 0.0002 and 0.00153 dÿ1.

These rates averaged three times faster than anaero-bic rates. To our knowledge, this is the ®rst study todemonstrate no signi®cant di�erence (P>0.05) in soilorganic C mineralization (CO2+CH4 production)using NO3

ÿ , SO42ÿ or CO2 as electron acceptors

(Fig. 4). These results indicate that the bioavailabilityof organic C was higher under aerobic conditions, butwas similar under denitrifying, SO4

2ÿ-reducing and

methanogenic conditions. Zehnder and Colberg (1986)

indicated that more complex soil organic constituents

(e.g. lignin and humic substances) are available to

aerobes through production of mono- and dioxygenase

enzymes that oxidize these chemicals. The activity of

these enzymes was indicated from O2-to-CO2 ratios

>1.0, in which excess O2 consumption may have been

due to incorporation into organic molecules. Under

each of the anaerobic electron acceptor reducing con-

ditions and when complex structural organic com-

pounds predominate, the lack of oxygenase activity

would be expected to curtail the initial hydrolytic and

oxidative steps that are rate-limiting to overall de-

composition (Kristensen et al., 1995). Hence, rates of

anaerobic decomposition were similar under various

electron acceptor reducing conditions. It has yet to be

established whether there are signi®cant di�erences in

mineralization rates of small molecular weight break-

down intermediates under di�erent anaerobic electron

acceptor reducing conditions.

In many of the Fe(III)-treated soils, CO2 evolution

into the headspace was negligible, which was likely due

to precipitation of Fe(II) with carbonate species form-

ing the mineral siderite (FeCO3). Sulfate reducers that

mediate Fe(III) reduction have been shown to promote

siderite formation (Coleman et al., 1993). This process

made it di�cult to determine organic C mineralization

for this treatment.

Table 4

Methane production potentials and percent inhibition by electron acceptors in wetland soils. Each value represents the mean of three replications

2one standard deviation and values in parentheses are number of days that inhibition was observed

Soil Lag time for

methane

production (d)

Methane production

potential

(mmol gÿ1 dÿ1)

Electron acceptor

Fe(III) SO42ÿ

initialb

(% inhibition)

®nal

(% inhibition)

initial

(% inhibition)

®nal

(% inhibition)

Organic

HLPI <1 6.2420.2 9423 (4) 7721 (16) 22242 (2) 9623 (18)

HLPU <1 2.1720.2 4924 (20) 47210 (10) 9524 (10)

W2 2 1.6720.1 9522 (20) 10023 (0)

W8 5 1.6320.2 9721 (20) 64233 (1) 10020 (19)

LSM <1 1.6320.1 ndc 10020 (20)

NCB 27 0.1620.0 nd 226 (13) 79215 (7)

LAAF 23 0.3520.0 nd 8521 (6) 10020 (14)

Mineral

TAL <1 1.1220.3 6721 (20) 10020 (20)

PPP <1 0.3920.0 9521 (3) 9920.2 (17) 7222 (3) 9820 (17)

CR 10 0.1220.0 6029 (20) 9122 (20)

a Methane production was completely inhibited by O2 and NO3ÿ for all wetland soils during the 20 d experiment, except for LSM sediment in

which methanogenesis was inhibited by 83% in the ®rst day followed by complete inhibition. bInitial and ®nal refer to inhibition observed

directly and several days after addition of electron acceptors to methanogenic soils, with the number of days indicated in parenthesis. cnd=not

determined.


Oxygen consumption followed a nonlinear trendwith rapid rates during day 1 (between 11 and 107mmol gÿ1 dÿ1) and declining thereafter (between 1and 42 mmol gÿ1 dÿ1) (Figs. 1±3; Table 6). Therewere signi®cant di�erences (P<0.05) in rates of O2

consumption among soils, with organic soils showingfaster rates than mineral soils. Molar ratios of O2 con-sumption-to-CO2 production during these two phasesaveraged 5.9 and 1.4, respectively. Ratios greater thanunity indicated consumption of O2 by reduced soilconstituents in addition to organic C, including Fe(II),NH4

+ and H2S. Under ®eld conditions, Phase I ratesof O2 consumption would be expected during day 1after exposing anaerobic soils to O2, such as followingdraining of soils, or resuspension of anaerobic sedi-ments into aerobic water column. These rates may also

be expected at aerobic±anaerobic interfaces, wherereduced substrates di�use from anaerobic zones andare rapidly oxidized in adjacent aerobic zones. PhaseII rates would be expected in soils after longer periodsof O2 exposure and would mostly be a result of hetero-trophic oxidation of organic carbon.

Denitri®cation occurred in three main phases: a slowphase in the ®rst 2 h (0 to 5.3 mmol N gÿ1 dÿ1, me-dian 0.23 mmol N gÿ1 dÿ1), a rapid phase from 1 to10 d (0.5 to 9.3 mmol N gÿ1 dÿ1) and a slower phaseduring the remainder of the experiment (0.2 to 4.7mmol N gÿ1 dÿ1) (Figs. 1±3 and Table 6).Denitri®cation in the ®rst phase, referred to denitri®ca-tion enzyme activity (DEA) (Smith and Tiedje, 1979),likely re¯ected rates in soils that receive very lowinputs of NO3

ÿ , such as where the only inputs to an-

Fig. 1. Methane production (a), CO2 production (b) and electron acceptor consumption (c) in Louisiana saltmarsh sediment. Anaerobic preincu-

bation (conditioning) denotes time before electron acceptor amendments. A value of unity was added to N2O-N concentrations to allow plots on

a logarithmic scale. Error bars represent2one standard deviation.


aerobic zones are from nitri®cation and di�usion fromaerobic zones.

Although the soils and sediments we studied wereconditioned under a methanogenic regime, most soilsstill maintained low rates of denitri®cation activitysuggesting that denitri®ers were surviving by alternatemetabolic processes (e.g. fermentation) and werepoised to exploit NO3

ÿ (Jorgensen and Tiedje, 1993).One exception was the salt marsh sediment whichshowed no detectable DEA; however N2O wasdetected after 1 d. There was no signi®cant di�erence(P>0.05) in DEA between organic and mineral soils.Denitri®cation in the second phase is equal to denitri®-cation potential (DP) of the soil and was signi®cantlygreater for organic compared to mineral soils. DEA

and DP measured in this study were in the rangemeasured in agricultural soils (Tiedje et al., 1982;Myrold and Tiedje, 1985; Pell et al., 1996), forest soils(Tiedje et al., 1982), wetland soils (Gro�man et al.,1992; Gale et al., 1993; Schipper et al., 1993) and lakesediments (D'Angelo and Reddy, 1993), but were smal-ler than measured in anaerobic digester sludge (Kasparet al., 1981).

Molar ratios of N2O+N2-to-CO2 production duringthree phases of denitri®cation averaged 0.3, 1.1 and0.65, respectively. Values below the theoretical value of0.8 suggest that alternate pathways besides denitri®ca-tion (e.g. fermentation) probably played a role in theterminal step of organic C mineralization. Highervalues suggested that alternate electron donors contrib-

Fig. 2. Methane production (a), CO2 production (b) and electron acceptor consumption (c) in Alabama Talladega sediment. Anaerobic preincu-




uted to N2O production, which may include reduced Scompounds mediated by sulfur oxidizing bacteria suchas Thiobacillus denitri®cans (Dannenberg et al., 1992).

All wetland soils showed potential for sulfate re-duction after exposure to SO4

2ÿ with rates rangingbetween 0.1 and 11 mmol gÿ1 dÿ1 (Figs. 1±3; Table 6).Under ¯ooded wetland conditions, most soils probablysustain some degree of SO4

2ÿ reduction since (i) theycontained signi®cant SO4

2ÿ concentrations (>200 mmollÿ1; D'Angelo and Reddy, unpublished results) thatare greater than the Km of 50 to 100 mM reported forthis process (Ingvorsen et al., 1981; Smith and Klug,1981) and (ii) SO4

2ÿ reducers are ubiquitous in the en-vironment, gaining energy from a wide selection ofelectron acceptors including O2, Fe(III) and nitrogen

oxides as well as fermentation (Dannenberg et al.,1992; Coleman et al., 1993). Belhaven muck agricul-tural soil showed a 6 day lag phase before sulfate re-duction was observed, which may have re¯ected thetime required for communities to switch from fermen-tation to SO4

2ÿ reduction (Fig. 3).Sulfate reduction proceeded signi®cantly faster

(P<0.05) in organic soils compared to mineral soils.Rates determined in this experiment were in the samerange reported for upland soils (Peters and Conrad,1996), wetland soils (Howes et al., 1984; Westermannand Ahring, 1987; Yavitt and Lang, 1990; Achtnich etal., 1995a,b) and lake sediments (Smith and Klug,1981; Bak and Pfenning, 1991). After lag periods insulfate reduction, the average molar ratio of SO4

2ÿ re-

Fig. 3. Methane production (a), CO2 production (b) and electron acceptor consumption (c) in North Carolina Belhaven soil. Anaerobic preincu-




duction-to-CO2 production equaled the theoreticalvalue 0.5, indicating that SO4

2ÿ reduction dominatedthe terminal step in organic matter decomposition.

Addition of Fe(III) to methanogenic soils resulted inan immediate accumulation of high amounts of Fe(II),which did not signi®cantly increase during the courseof the experiment (data not shown). High backgroundFe(II) was likely caused by chemical reduction withaccumulated H2S, which made it di�cult to determinebiotic Fe(III) reduction rates for this treatment.Abiotic reduction could have probably been avoided

by ®rst aerating the soils to oxidize the reduced con-stituents, however, this was not done because of O2

e�ects on soil chemistry and microbial activities.

3.4. Relationships between microbial activities andbiological and chemical properties

Combining biological and chemical properties for allsoils, except those from the hypereutrophic marsh(HLPI), microbial biomass C, inorganic N, DOC, totalC and total N were signi®cantly correlated(0.57< r 2<0.97; P<0.05) to potential rates of O2

consumption, denitri®cation, SO42ÿ-reduction and

methanogenesis in wetland soils (Table 7 and Figs. 5and 6). Inorganic N and microbial C were the most re-liable and best correlated to heterotrophic activity inall wetland soils (including HLPI). These were likelygood predictors because they are indicators of theamount of soil organic C that is readily utilizable byheterotrophic microorganisms. It is suggested thatmeasurement of these soil chemical properties mayprovide a good estimation of potential rates of mi-crobial activity in di�erent wetland systems, whichmay be helpful when more labor and time-consumingexperimentation is not possible.

One soil (HLPI) did not ®t several of the microbialactivity±soil property relationships observed for theother soils. For example, aerobic, denitrifying, sulfatereducing and methanogenic activities in this soil weremuch higher than expected based on its DOC concen-tration (Fig. 5). Therefore, this soil was not includedin establishing the relationships. Because this soil wasfrom a hypereutrophic marsh that received domestic

Fig. 4. Relationship between aerobic and anaerobic CO2+CH4 pro-

duction rates under NO3ÿ-reducing, SO4

2ÿ-reducing and methano-

genic conditions in wetland soils. Organic carbon mineralization with

NO3ÿ and SO4

2ÿ as electron acceptors was not determined for HLPI

soil.

Table 5

E�ect of electron acceptors on CO2 production rates of wetland soils

Soil Electron acceptor

O2 (mmol gÿ1 dÿ1) NO3ÿ (mmol gÿ1 dÿ1) SO4

2ÿ (mmol gÿ1 dÿ1) CO2 (mmol gÿ1 dÿ1) LSDa

Organic

HLPI 33.5 ndb nd 8.6 1.25

HLPU 11.3 6.0 3.7 2.1 0.67

W2 13.3 5.6 5.4 3.8 0.62

W8 10.8 3.7 4.5 2.6 0.64

LSM 8.9 3.7 4.1 2.3 0.69

NCB 3.6 1.1 0.5 0.4 0.19

LAAF 4.7 1.7 2.1 2.1 0.61

Mineral

TAL 10.7 2.9 2.6 1.7 0.75

PPP 2.5 0.9 0.7 0.5 0.20

CR 1.0 0.3 0.3 0.2 0.08

LSDc 0.59 0.56 0.54 0.44

a Least signi®cant di�erence (a=0.05) between electron acceptor treatments. bnd=not determined. cLeast signi®cant di�erence (a=0.05)

between soil treatments.


Table 6

Electron acceptor consumption rates in wetland soils

Soil Electron acceptor

O2 NO3ÿ SO4

2ÿ

phase I

(mmol gÿ1 dÿ1)

phase II

(mmol gÿ1 dÿ1)

LSDa phase I

(mmol gÿ1 dÿ1)

phase II

(mmol gÿ1 dÿ1)

phase III

(mmol gÿ1 dÿ1)

LSDa

(mmol gÿ1 dÿ1)

Organic

HLPI 107.0 42.3 12.9 5.29 9.27 4.70 1.72 11.10

HLPU 42.6 16.7 14.9 0.82 2.88 1.69 0.21 1.81

W2 53.7 20.4 6.7 0.39 5.24 2.66 0.25 2.56

W8 43.3 19.8 10.0 0.19 4.26 2.41 0.55 2.59

LSM 71.7 17.1 1.8 0.00 2.49 0.59 0.45 1.66

NCB 30.3 4.9 3.3 0.19 0.93 0.34 0.11 0.35

LAAF 18.6 3.2 2.6 1.10 1.03 0.62 0.09 0.14

Mineral

TAL 45.7 12.4 3.8 0.20 2.88 2.52 0.10 0.93

PPP 19.3 3.4 1.7 0.43 1.12 0.80 0.24 0.43

CR 11.2 1.0 1.1 0.22 0.45 0.17 0.07 0.12

LSDb 7.9 1.1 0.46 0.59 0.49 0.39

a Least signi®cant di�erence (a=0.05) between di�erent phases of electron acceptor consumption.bLeast signi®cant di�erence (a=0.05)

between soil treatments.

Table 7

Regression equations relating wetland soil properties (independent variables) and rates of O2 consumption, denitri®cation, sulfate reduction and

methanogenesis (dependant variables). Regression analysis includes all soils unless otherwise indicated. Microbial activities were not signi®cantly

correlated (P>0.05) with pH, bioavailable Fe, soluble Ca+Mg, or available P

Soil property Electron acceptor

O2 NO3ÿ SO4

2ÿ CO2

phase I phase II phase I phase II

Aerobic C mineralization (mmol gÿ1 dÿ1) slope 2.81 1.29 0.15 0.28 0.34 0.19

intercept 16.3 1.22 ÿ0.55 0.20 ÿ1.25 ÿ0.35r 2 0.83** 0.94** 0.73** 0.94** 0.94** 0.96**

Anaerobic C mineralization (mmol gÿ1 dÿ1) slope 6.17 2.83 0.33 0.62 0.77 0.42

intercept 19.8 2.87 ÿ0.42 0.57 ÿ0.87 ÿ0.13r 2 0.82** 0.92** 0.79** 0.92** 0.96** 0.97**

Microbial biomass C (mmol gÿ1) slope 0.81 0.42 0.047 0.094 0.11 0.062

intercept 13.5 ÿ1.93 ÿ0.89 ÿ0.55 ÿ1.99 ÿ0.83r 2 0.57* 0.82** 0.63* 0.85** 0.78** 0.85**

Inorganic N (mmol gÿ1) slope 0.84 0.38 0.042 0.084 0.10 0.055

intercept 19.3 2.81 ÿ0.35 0.53 ÿ0.87 ÿ0.091r 2 0.85** 0.93** 0.71** 0.95** 0.96** 0.92**

Total C (mmol gÿ1)a slope 0.43 0.092 0.059 0.04

intercept nsc 3.23 ns 0.68 0.12 0.29

r 2 0.66** 0.65* 0.75** 0.59*

ln dissolved organic C (ln mmol gÿ1)a, b slope 14.3 6.71 1.44 0.81 0.61

intercept ÿ20.1 ÿ16.0 ns ÿ3.46 ÿ2.07 ÿ1.43r 2 0.50* 0.81** 0.81** 0.71** 0.67*

Total N (mmol gÿ1)a slope 8.49 1.89 1.18 0.79

intercept ns 2.98 ns 0.57 0.065 0.28

r 2 0.74** 0.78** 0.87** 0.65*

a Correlation analysis does not include soil data from the hypereutrophic HLPI marsh.bRelationship was best described by logarithmic equa-

tion.cns=not signi®cant.*P<0.05.**P<0.01.


wastewater, it is likely that its fraction of labile C inthe DOC pool was higher than for the other wetlands.Sulfate reduction and DOC soils data from anotherhypereutrophic marsh in central Florida (LAMP)(D'Angelo and Reddy, 1995) also fell well beyond therange established for the relatively unimpacted wet-lands (Fig. 5). These results suggest that microbial ac-tivity±soil property relationships may be potentiallyuseful as tools to evaluate pollution e�ects upon wet-lands, with measurements above or below predictedvalues indicating the zone of exposure. However, veri-®cation of this hypothesis will require additionalmeasurements from a�ected wetlands.

Soil properties such as available P, dissolvedCa2++Mg2+, pH and bioavailable Fe were not sig-ni®cantly related with most heterotrophic activities(P>0.05), suggesting that nutrient availability metrespiration requirements in these wetland soils. Thisresult may be attributed to e�cient nutrient uptakesystems of microorganisms (half-saturation coe�cientin the micro- to low millimolar range; Button, 1985)

and homeostatic control mechanisms that control pHdespite changes in external pH (Padan, 1984). The lackof negative correlations with H2S suggested that activi-ties were not inhibited by H2S toxicity. Together theseresults demonstrate that (i) availability of electronacceptors and organic substrates are the dominantchemical regulators of heterotrophic potentials in mostorganic and mineral wetland soils and (ii) measure-ments of these characteristics may provide estimates ofheterotrophic potentials of wetland soils.

4. Conclusions

Our study demonstrated the importance of severalsoil factors in regulating potential rates and modes oforganic carbon mineralization in wetland soils, withelectron acceptor and donor availability found to bedominant. Electron acceptors including O2 and NO3

ÿ

generally resulted in complete and immediate inhi-bition of methanogenesis; Fe(III) and SO4

2ÿ often

Fig. 5. Relationships between dissolved organic C and potential rates of oxygen consumption, denitri®cation, sulfate reduction and methanogen-

esis in wetland soils. Symbols for di�erent soils are de®ned in Table 1. Data from hypereutrophic marshes Houghton Lake (HLPI) and Lake

Apopka (LAMP) were not included in regression analysis.


resulted in less e�ective inhibition. While aerobic or-ganic C mineralization rates were about three timesfaster than under anaerobic conditions, there was nosigni®cant di�erence in rates with NO3

ÿ , SO42ÿ or CO2

as electron acceptors. Including data from all wetlandsoils except the hypereutrophic HLPI marsh, aerobic,denitrifying, sulfate reducing and methanogenic poten-tials were strongly correlated to microbial biomass C,inorganic N, DOC, total C and total N. Relationshipsfound between these variables and microbial activitiesin our study may have dual signi®cance: (i) they pre-dict potential rates of microbial processes across di�er-ent wetland systems and (ii) they assess zones ofadverse pollutant e�ects on wetlands.

Utilization of these relationships in the ®rst case isgiven in the following example. Calculated maximumrates of SO4

2ÿ reduction and methanogenesis are 0.42and 0.35 mmol cmÿ3 dÿ1 for a uniform ¯ooded peatsoil with the following characteristics: 40 cm totaldepth, bulk density=0.3 g cmÿ3, total C=22 mmolgÿ1, O2 penetration depth=2 cm, and SO4

2ÿ

input=10 mmol cm2ÿ dÿ1. Depths of sulfate reduction

and methanogenesis are calculated to be 24 and 14 cm,respectively, with a potential methane release rate of 5mmol cmÿ2 dÿ1. Rates of organic C mineralization onan areal basis are calculated to be 6, 13 and 12 mmolcm2ÿ dÿ1 in the aerobic, SO4

2ÿ-reducing and methano-genic zones, respectively. While calculated rates in thisexample are within ranges measured in wetland sys-tems (Schutz et al., 1989; Boon and Mitchell, 1995;Crozier et al., 1995), the database from which these re-lationships were derived needs to be increased toinclude soils from a�ected and pristine systems. Sincethe fate of many pollutants (e.g. nutrients, heavymetals and toxic organics) are also regulated by redoxconditions, these types of relationships may provideuseful tools for modeling their fate in the soil pro®le.While these calculations provide useful estimates, insitu ¯ux rates may be in¯uenced by other processes(e.g. di�usion and convective transport, removal ofmethane by methanotrophs, nutrient uptake by plants)and environmental conditions (limiting electron accep-tor supply, water content and temperature). Futureresearch should be conducted to incorporate these fac-

Fig. 6. Relationships between microbial biomass C and potential rates of oxygen consumption, denitri®cation, sulfate reduction and methanogen-

esis in wetland soils. Symbols for di�erent soils are de®ned in Table 1.


tors into the relationships developed in this study toimprove predictions under a variety of environmentalscenarios.

Acknowledgements

This research was supported by funding from theUS Department of Agriculture National ResearchInitiative Competitive Grant Program. We gratefullyacknowledge the cooperation of several researcherswho provided soils used in the study: Dr. E. Roden(University of Alabama), Dr. C. Lindau (LouisianaState University), Dr. C. Crozier (North CarolinaState University), Dr. J. Richardson (North DakotaState University), Dr. R. Kadlec (WetlandManagement Services, MI), Dr. R. DeLaune(Louisiana St. University) and J.R. White and M.M.Fisher (University of Florida) and the statistical analy-sis advice of J.M. Harrison (Senior Statistician,University of Florida).

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