Using Natural-Gradient Tracer Testst

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0099-2240/91/071997-08$02.00/0

In

Situ Measurement

of Methane Oxidation in Groundwater by

Using Natural-Gradient Tracer

Testst

RICHARD L. SMITH,'* BRIAN L. HOWES,2AND STEPHEN P. GARABEDIAN3

WaterResources Division, U.S. GeologicalSurvey, Lakewood, Colorado802251; BiologyDepartment, Woods Hole Oceanographic Institution, Woods Hole, Massachusetts 025432; and Water Resources Division, U.S. Geological Survey,

Marlborough, Massachusetts

017523

Received 11 February 1991/Accepted 17 April 1991

Methane oxidationwasmeasuredinanunconfinedsand and gravelaquifer (Cape Cod, Mass.) by using in situnatural-gradient tracertests atbothapristine, oxygenated site andananoxic, sewage-contaminated site.

Thetracersiteswereequipped with multilevel sampling devicestocreate targetgrids of sampling points; the

injectatewas prepared withgroundwater from thetracersiteto maintain thesamegeochemical conditions.

Methane oxidation was calculated from breakthrough curves of methane relative to halide and inert gas

(hexafluoroethane) tracers and was confirmed by the appearance of

13C-enriched

carbon dioxide in

experi-mentsinwhich "3C-enrichedmethanewasusedasthetracer.A Vmaxfor methane oxidation could becalculated when themethane concentration was sufficiently highto result in zero-order kinetics throughout the entire transport interval. Methane breakthrough curves could be simulated by modifying a one-dimensional

advection-dispersion transport model to include a Michaelis-Menten-based consumption term for methane oxidation. The Kmvalues for methane oxidation thatgavethe best match for the breakthroughcurve peaks were 6.0 and 9.0 ,uM forthe uncontaminated and contaminated sites, respectively. Natural-gradient tracer

testsare apromising approach for assessingmicrobialprocessesandfor testing in situ bioremediation potential ingroundwatersystems.

Estimating the distribution and rates of microbial pro-cesses in groundwater environments has proven to be a difficulttask. This is because many problems arise when the traditional methods developed for soils or aquatic sediments are applied to subsurface systems (8). These approaches usually include attempting to obtain undisturbed and uncon-taminated subsurface samples and maintaining in situ condi-tions during activity measurements (29). However, many groundwater systems present the opportunity to assess microbial processes with a unique approach, the use of in situnatural-gradienttracer tests.Thesetestswereoriginally designed and have beenused primarily to study the hydro-geology ofaquifers (6, 19), but with appropriate modifica-tions theyhave thepotentialto assessmicrobialprocesses as well (23). The advantage of such tracer tests over other approaches is thatthey usenatural groundwaterflow (i.e., natural gradients) to transport tracers through undisturbed sections ofanaquifer (6, 7, 11, 19, 23, 27).

In this study, we adapted the tracer test technology to studyamicrobiallymediated process, methaneoxidation,in situ inan unconfined sand andgravel aquifer. We choseto

study methane oxidation because of the potential use of methaneinjectionas aninsitutreatmentprocessfor removal of halogenated aliphatic compounds from contaminated groundwater (25,27, 35)andpreliminarydataindicatedthat the potential for methane oxidation existed in the aquifer understudy. Using in situtracertechniques,we wereableto demonstrate that methane oxidation couldoccurwithin both a pristine, well-oxygenated region of the aquifer and a

nitrate-enriched, anoxic section,andby couplingtheresults with a groundwater transport model we were also able to determine the kinetics of methane oxidation under both

* Correspondingauthor.

tWoods Hole OceanographicInstitute contribution no.7633.

conditions. Our results demonstrate the utility of in situ natural-gradient tracer tests to study microbial processes in groundwater systems.

MATERIALSANDMETHODS

Study site. Thestudy site is afreshwater sandandgravel glacial outwash aquifer located on Cape Cod, Mass., which has been contaminated by land disposal oftreated sewage since1936.Theresulting contaminant plume ismorethan 3.5 kmlong, 0.9 kmwide,and 23mthick(17, 30). Ingeneral,the contaminantplumeconsists ofverticalandhorizontal

gradi-ents of specific conductance (50 to 400

iuS),

dissolved oxygen (0 to 8 mg/liter), pH (5 to 7.1), dissolved

organic

carbon(1to4 mg/liter,muchof itrefractoryinnature), and inorganic compounds, such as chloride (0 to 28 mg/liter), sulfate(4to30 mg/liter), nitrate (0to 16 mg of N per

liter),

andammonium (0to10 mgof N perliter) (4, 17, 18, 31, 33). These gradients haveresulted in

corresponding

distribution patterns of bacterial abundance and

heterotrophic

uptake (12, 31). Thechemical makeup oftheplume,

especially

with regard toinorganic nitrogenspecies, is similartothat asso-ciated with more than 17 x 106 residential

septic

systems across the United States (3) and therefore represents in-creasingly common contamination conditions in regional aquifers. The aquifer

receiving

the contaminants is com-posedofstratifiedsand and

gravel

witha mean

particle

size and porosity of 0.5 mm and 0.39, respectively (17), and a horizontalflowrateofabout 0.4mperday (6). The

ground-watertemperature in the contaminantplume is 10 to 14WC

(17).

Tracer testsites. Methaneoxidationwas

assayed

insitu in thecenter ofthecontaminant

plume

0.25 km

downgradient

from the contaminant source and in an uncontaminated portionof the

aquifer by

using natural-gradient

tracertests. Eachsiteconsisted ofanarrayof15-portmultilevel

sampling

1997

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FIG. 1. Diagram ofthe well locations at the two small-scale, natural-gradienttracer test sites.

devices (MLSs) situated inthe same configuration (Fig. 1). Construction and installation ofthe MLSs have been de-scribed previously (19, 30). At the contaminated site (well F347), the verticalspacing between the MLS ports was 0.6 m, coveringthe interval from5.6 to 13.7 m below the land surface. For theuncontaminated site (wellF168)the vertical spacing betweentheports was 0.4 mfortheinterval from5.0 to 10.3 m below the land surface.

WatersampleswereobtainedfromeachportoftheMLSs with a peristaltic pump (30). Specific conductance was determined by using an electrode, oxygen was determined byiodometric titration (1), and nitrate wasdeterminedwith an Autoanalyzer(Lachat Instruments, Menquon, Wis.).

Tracer tests. The methods used to prepare the tracer solutions differed at the two tracer sites; each method was chosen with particular regard to maintaining the in situ oxygen concentration atthe site. In both cases, the tracer solution was composed of groundwater (100 liters) with-drawn from two adjacent ports ofthe upgradient injection wellMLS with aperistalticpump. Within the anoxic zone of thecontaminated site, thegroundwaterwas pumped into a gas-impermeablebag which had beenpreviouslyflushed five times with N2 and contained 1 liter of a concentrated N2-sparged solution of a conservative tracer (NaCl or NaBr). During the first tracer test, the bag was vented and tracer gases were added by simultaneously bubbling the solutionwith C2F6(14.7liters/min)and CH4 (1.6 liters/min) for1h. For thesecond tracer test 1 liter of

13CH4

(99% pure; Cambridge Isotope Laboratories, Woburn, Mass.) was addedtothebag(which was not vented) and wasallowedto

equilibrate

for 12 h. The excess CH4 gas was then vented,

TABLE 1. Injectate compositionfor each of the natural-gradienttracertests

Injectate Injectate

Tracer test constituent concn'

Contaminatedsite Test 1 C1- 41.1 mM C2F6 25.1,uM CH4 85.0p.M Test 2 Br- 14.7 mM 13CH4 180.2

F.M

Uncontaminatedsite Br- 12.8 mM CH4 66.3 ,uM 02 225.0,uM

a The values are averages of the values for subsamples takenduringthe

injectionprocess.

and the contents of the bag were thoroughly mixed. The tracersolutions werepumped back into the ground through the same two adjacent ports (0.6 m apart) of the

injection

well withaperistalticpump,takingcare toavoidinjection of any gas bubbles. Samples for gas and ion analysis of the tracersolution werecollected atthebeginning, middle, and end of the injection. For the uncontaminated, aerobic groundwater system, groundwater was pumped into an open container,MgBr2 was added,and the solutionwas bubbled withCH4 (1.6 liters/min)and air(14.7liters/min)for 45min. The tracer was then injected into the ground in the same manner as described above for the contaminated site. The injectatecomposition for each of the tracer tests is shown in Table 1.

Water samples were taken daily with a peristaltic pump from each of the 60 ports in each of the sampling grids. The first 120 ml of each sample wasdiscarded,and the next 20 ml wascollected inasyringeandinjected into a stoppered 50-ml serumbottle, which was then inverted. Most samples were analyzed on the day ofcollection; afew were stored for 1to 3 days at 4°C. The dissolvedgases in the bottle headspace after equilibration at room temperature were assayed by gas chromatography (Porapak N; flame ionization

detector),

after which the halides were assayed with an ion-specific electrode (corroborated by using ionchromatography). For the

13C

tracertest, 1-litersamples werecollected, dissolved inorganic carbon was precipitated as SrCO3 (by adding 60 ml ofSrCl2 6H20 in NH4OH [4g/g]per liter of watersample) and filtered, and the

13C/12C

content was determined

by

using mass spectrometry (31).

Model formulation. A modifiedadvection-dispersion equa-tion with a Michaelis-Menten utilization term was used to simulate the movement, spreading, and losses to microbial consumption of methane in the small-scale injection tests. The one-dimensional transport equation, which contains terms for storage, advection, dispersion, and degradation, canbe written as:

dc

a2c

ac Vmaxc

-

=aD

ax-

v -

+

at aX2 ax c+Km (1)

where c is theconcentration ofCH4 (or solute) insolution,t is time, D is the dispersion coefficient (= aLv), aL iS dispersivity, xis the spatial coordinate, vis thefluid veloc-ity, Vmax is the maximum rate ofCH4degradation,and Km is theMichaelis-Menten constant.

The modelingapproach used in this studyto simulatethe transport of methane was to find the physical parameter

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values (aL, v) by calibrating analytical advection-dispersion model solutions to the observed breakthrough curves of the nonreactive solute (Cl- or Br-) and to find the microbial degradation parameters (Vmax, Kin)bycalibrating numerical solutions of equation 1 to observed methane breakthrough curves. In the case of chloride, there was no degradation (Vmax = 0), and the fluid velocity (v) was calculated from the known distance between the injection point and the sampling point(x1) and the time to the peak concentration (tpeak) for the breakthrough curve (v =

xl/tpeak).

The dispersivity parameter (aL) was obtained from the chloride curve by using the following relationship, which was derived from the one-dimensional solution of the advection-dispersion equa-tion with a pulse input (9):

XL(At/tpeak)2

16 In (2)

(2)

where At is the duration of the breakthrough curve whenCl(,) >1/2

Clmax,

Cl

max

is the peak concentration, and tpeak is the timetopeakconcentration.

Equation1 wassolvedfor the reactive solute, methane, by using finite-difference approximations with a Crank-Nichol-sontime-steppingscheme (13). After the approximations are introduced, the equation is reduced to an algebraic form:

-D -v

c(x-lAxt+

At)[)2

+

4(Ax)

+

1 D Vmax 1

C(X,t

+

L+Att

t)

+

_+(AX)2

₊ _{2[Km + c(x,t +}I~

At)]_

+

-D v

c(x+Ax,t+ At)

(A[)2

+ 4(x =

D v

c(x -AX,t)

T(A-x+

4(Ax)

+

c(x9t)[K.

+ + Lm +

LAt

2[Km

+

~~c(x,t)]

D -v

c(x+

Axt)[2(A

)2 + 4(Ax) (3) whereAx istheblocksize (0.1 m) andAtis the time step (0.1 day). Allofthe concentrationvalues ontheleft side [c(x

-Ax,

t + At), c(x, t + At), c(x +

Ax,

t + At)] are at thenew

timestep(t+At)and areunknown;theconcentration values ontheright side [c(x- Ax, t), c(x, t), c(x + Ax, t)]are atthe present-time level and are known. Equation 3 is nonlinear because concentration appears in the denominator of the Michaelis-Menten termontheleft side.Therefore, an itera-tive scheme is usedtosolve theequation.Theconcentration valueusedin thedenominator oftheMichaelis-Mententerm is held atavaluefrom the previous iteration level. The fixed value is then updated with every iterative step until the solution converges and is stable. After convergence the newlycalculated concentration values arethen used as the present-timevalues and anothersetis calculated for thenext time step, marching onwardin time. Because of the form of equation 3, a tridiagonal matrix solution is used, which is both efficient andaccurate. Restrictions onthe Peclet num-ber and the Courant number, required for model stability,

werecalculatedfor thesesimulations,and the blocksize and

TABLE 2. Groundwater chemistry at the tracer test sites'

Specific Oxygen Nitrate Methane Testsite conductance concn concn concn

(,uS) ($LM) (>LM) (>M)

Contaminated 281 <0.3 650 <0.1

Uncontaminated 58 163 0 <0.1

aValuesatthedepth of the injection ports.

thetime step were limited to satisfy these restrictions (13). The root mean square error (RMSE), which was used to compare differences between the observed and predicted breakthrough curveconcentrations, was calculated byusing thefollowing equation:

RMSE = [- (Cpredicted - Cobserved) (4)

where nis the numberof observed concentration values (c). RESULTS

The two tracer sites were chosen to reflect the general differences in the geochemical environment between the contaminant plume and uncontaminated groundwater. The uncontaminated groundwater is well oxygenated, has a

specific conductance of 60 to 80 ,uS, and does not contain any of the characteristic constituents of the contaminant plume, suchasnitrateorammonium (Table2). However, the contaminant plume is anoxic or suboxic (dependingon the particular location and depth) and is characterized by re-gions of nitrate-containing groundwater. At the contam-inated tracer site there was no detectable oxygen and 650 ,uM nitrate (Table 2). The contaminant plume at this site extendsfrom 6to35 mbelow the land surface. Methane is not a constituent of the plume and was not present in detectable amounts ateithertracer testsite (Table 2).

Even though groundwater methane concentrations were below atmospheric equilibrium levels, natural-gradient

tracer tests demonstrated that the indigenous microbial populations within both the contaminated and uncontami-nated portions of the aquifer possessed the capability to

oxidize methane when itwas added to the

aquifer.

Tracer test breakthrough curves at the contaminated site demon-strated that methane, hexafluoroethane, and chloride were transportedatequivalentrateswithin theaquifer. The

peak

concentration ofeachtracerarrivedatthe

sampling

wellon the same day (Fig. 2). However, when normalized to the injectate concentration, the two conservative components, hexafluoroethane and chloride, had identical

breakthrough

curveswhile the methanecurve was

significantly

attenuated.

Asimilar resultwasevident fora tracertestconductedatthe uncontaminated site(Fig. 3A), where underoxic conditions methane was significantly attenuated

during

transport through the aquifer. Because of differences in groundwater velocity, the arrival of the peaktracer concentration atthe uncontaminated site was 8 days later. In

addition,

the

highest relative concentration

(C/CO)

of the conservative tracer at anysamplingportwas only 60%of the concentra-tionof the contaminated site.

Foranygiven sample, the amountof methane consumed could be calculatedasthedifference between the dimension-less concentration

(C/CO)

for unattenuated transport

(from

hexafluoroethane,

bromide,

or

chloride)

and the

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0.14

_- C06

14 18 22 26 30 34 38

DAYS AFTER INJECTION

FIG. 2. Timecourse of the relative concentrations ofchloride, hexafluoroethane, and methane at the tracer site within the contam-inant plume. The data are from a sampling port that was 9.7 m beneath the land surface.

less concentration ofmethane, times the injectate methane

concentration. Figure 3B shows the result of this calculation

plotted as a function of time for the sampling port exhibiting

the highest tracer concentrations for the tracer test at the

uncontaminated site. The curve is shaped like a

break-through curve because it shows the amount of methane

consumed within various portions of the tracer cloud as the cloud passed by a fixed point (i.e., the fence of target wells).

To test the hypothesis that theattenuation of methane was

the result ofmethane oxidation, a second tracer test was

conducted at the contaminated site by using '3C-enriched

methane. Methane consumption was again evident at this

site; up to 7.6 ,umol of methane per liter was consumed

during 17 days of travel time between the injection well and

the target wells (Fig. 4A). The natural isotopic abundance

(5'3C) of dissolved inorganic carbon at this location in the

contaminant plume (and also at the target well fence just

prior to the arrival of the tracer) was -18.7%o(Fig. 4B). This

valueshifted to +11.5%oin the center of thetracer cloud and

then dropped back to -17.5%o immediately after the tracer

cloudhad passed the well fence (Fig. 4B). The '3CH4 tracer

contained no detectable CO2.

The amount of methane consumed during a tracer test

(expressed as concentration) was directly related to the

methane concentration within the tracer cloud as it passed

by thewell fence (Fig. 5). This relationship is empirical and

is site specific because it compares the total amount of

methane consumed over the entire transport interval (x1)

with the methane concentration at the end ofthe transport

interval. However, the relationship is reproducible for a

given site. Figure 5 shows a composite of the results from the

contaminated tracer test site for four sampling ports from

tracer test 1 and for two sampling ports from tracer test 2

(those ports for which CH4 was consumed and C/CO was

>0.01 at the CH4 peak). In general, the relationship at this

site indicated a first-order reaction when methane

concen-trations at the well fence were <6 ,uM and a zero-order

reaction when themethane concentration was >15 ,uM.The

maximum amount of methane consumed was 7,uM.The rate

ofmethane consumption could be calculated from this value

for any sample whichhad a methane concentration >15 ,uM

because theconcentration would have been >15 ,uMduring

0.06 z 0 z ,', 0.04 0 z 0 0.02 w o. 0:0 E 3.0DI A 60 0 4.0-I 3.0-01J 2.0222-6 2 0 2 3 6 3

DAYSAFTER INJECTION

FIG. 3. Time course of the relative concentrations ofbromide and methane(A) and theamountof methane consumed(B)atthe tracersite withinanuncontaminatedregionof theaquifer.Thedata arefromasamplingportthatwas7.26mbeneath thelandsurface. The methaneconsumption calculationwasbasedontheamountof methane missing compared with the amount of the conservative tracer(cross-hatchedareainpanel A).

the entire transport interval. This rate (Vmax) was0.4 FM/ day at the contaminatedtracersite.

The computer modelwas able toeffectively simulate the breakthrough curves of both chloride and methane at the contaminated (Fig. 6) and uncontaminatedtracersites. The modelvelocity (v) and thedispersivity (CXL)parameters were calibratedby fitting the chloride breakthrough curves (Fig. 6A) with the Vmax set to zero (Table 3). The velocity and dispersivity were found to be somewhat lower for the uncontaminated site, reflectingthe longertransport time at that site(Fig. 2 and 3). Then, byusingthese values forvand aL, the Michaelis-Menten parameters

(Ki,m

Vmax) were de-termined by adjusting their values to match the model predictionsto themethanebreakthroughcurve.The model-simulated concentration values werefound to be inversely relatedto Vmax andparticularly sensitivetoVmaxatthepeak concentration values (Fig. 6B). Simulated concentrations

were foundto be directly related to the Km (Fig. 6C), and peak concentrations were again most sensitive to its changes.

The model does not have a unique solution for Km and Vmax; similar simulations can be obtained for completely differentsetsofvalues. However,because Km and Vmaxare

mathematically related, a unique Km can be obtained for eachgiven Vmaxbyoneoftwo methods.The Kmvalue can

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0 LLJ w

CD

0a)

wE Z=L I w 0 m C) Co FIG. values f tracer te arefrom beobtai best fit through that be, The fir breakth emphas cloud. liter pei 5, the tracers mean s visual i extrapo comput maximl Smal natural an aqui Natural to asses potenti; bacteria

through

0 0 o 0 C,) z 03 0 20 0 o 0 0 4 8 12 16 20 24 CH4(>M)

FIG. 5. Relationship between theamountofmethaneconsumed foranygiven tracersample andthe measuredmethane

concentra-tion in thatsample. The plotis acomposite ofthe datafrom four depths (8.5, 9.1, 9.7,and 10.2mbelowthe landsurface) fromboth tracer teststhat wereconductedatthecontaminated site. potential to assessmicrobial processesingroundwater

hab-I / /X itats. Because the groundwater environment is similar to the

%/7_

sediments

of surface water environments and to

soils

in

sediet

sunsaturated

ofenvironments

(i.e.,

in each case thevast

major-I ] ,

ity

of the microbial

community

is attached to

particulate

10 12 14 16 18 20 22 24 surfaces), natural-gradient tracer tests represent a unique

opportunity

tomeasurethe in siturateofamicrobial process DAYS AFTER INJECTION in an undisturbed, sediment-dominated ecosystem.

4. Time course of methane consumption (A) and 813C The results of this study demonstrate that small-scale or dissolved inorganic carbon (DIC) (B) during a second natural-gradient tracer tests can be used to measure methane

st at the contaminated site with

13CH4

(99% pure).The data oxidation in a sand and gravel freshwater aquifer. The loss of asampling portthat was9.7 m beneath the land surface. methane during the tracer tests was not due to physical processes, such as dissolution of dissolved gases into

bub-bles,

because bothaninert gas

(C2F6)

and conservative ions inedbyselectingthe valueeitherbyusingastatistical were transported identically, or to retardation of methane by between the simulation and the methane break- sorption, because the shapes of the methane breakthrough

curve(Fig. 7) or by visually selecting the simulation curves were identical to the shapes of the conservativeion

st matches the peak concentration (Fig. 6B and C). curves. The observed attenuation of methane was most st approach gives equal emphasis for the entire likely the result of methane oxidation by the resident micro-rough curve, while the latter approach places more bial population within the aquifer. This process was clearly

;isonmethane

consumption

inthecenterofthetracer demonstrated as the operative mechanism with

13CH4;

'3C

By

using

theVmax of0.40 ,imol of CH4consumed per enrichment of the dissolved inorganic carbon pool could

rdaydetermined from therelationshipshowninFig. only have resulted from oxidation of the

'3CH4

tracer. The Km for methane consumption at the contaminated symmetry of the CH4 breakthrough curves (Fig. 2 and 3) ;itewasdeterminedtobe 5.2

FiM

when the leastroot suggests that the adaptation period for CH4 as a substrate quare error method was used and 9.0 ,uM when the was relatively short (<24 h) at both the contaminated and method was used. When this value for Vmax was uncontaminated tracer test sites, even though in situ CH4 latedtotheuncontaminated, aerobic

site,

the model concentrations were less than 0.7

,uM

in each case. In Led a Km of 6.0 ,uM for the best match to the addition, the process exhibited a saturation type of kinetics am methane concentration. (Fig. 5), a clear indication of a biologically mediated process.

During

its transport

through

an

aquifer,

a tracercloud is diluted because of

dispersion.

In sand and

gravel

aquifers

DISCUSSION longitudinal dispersion is much greater than transverse ver-1-scale

natural-gradient

tracer teststake

advantage

of tical andtransversehorizontal

dispersion. Consequently,

the

groundwater flowtotransporttracersintroducedinto tracercloud remains narrow but becomes

elongated

in the ifer through an undisturbed section of the

aquifer.

direction of

groundwater

flow

(19, 23).

During

a

typical

1-gradient

tracer testshave been used in this manner small-scaletracertestthetracers areoften

only

detectedat sstherelative transport of

inorganic compounds (19),

a

single

target MLS. The

general

lack of vertical

mixing

is al

organic

contaminants (23),

viruses,

yeasts, and also evident on a

plume-wide

basis at the

Cape

Cod

site,

a (7), and indigenous groundwater bacteria (11)

resulting

in

relatively

steep vertical concentration

gradients

i various

aquifers.

These tracer tests also have the ofdissolved

constituents,

such as oxygen and nitrate

(30).

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15 10 6 C-) N-I 5 0 12 10 8 6 4 2 0 10 8 6 4 2 0 0 10 20 30

0 10 20 30 40

TABLE 3. Valuesfor unknownparametersin themodel usedto

obtain best-fitsimulation

Test site _(rn/day)V OL_ct (m)_i) _(p.M/day)Vmax _GpM)

KM

Contaminated(anoxic) 0.285 0.065 0.4a 9.0b Uncontaminated (oxic) 0.220 0.0104 0.4a 6.0b

aValueobtainedempirically frombreakthrough curves at thecontaminated

tracersite(seeFig.5).

bBest-fit selection based on visualinspection of peak concentrations of breakthroughcurves.

Smallvertical dispersion also meansthat fine-scale vertical resolution is possible during these small-scale tracer tests when it is desirable to restrict the tracer to narrow vertical zones.

_.4 The neteffect of dispersion is that the concentration ofa 40 traceris continually being reduced. Consequently, when a

traceris the substrate ofamicrobially mediated reaction, the rate of the reaction also changes, probably in accordance with Michaelis-Mentenkinetics, as the tracer is transported. However, the rate ofreaction is constant inthose cases in which the tracer concentration is high enough to result in a zero-order reaction during the entire transport interval. In thatsituation, theamountoftracerconsumed isindependent of thetracerconcentration (Fig. 5), and ina manner analo-gous to the saturation type ofkinetics, such consumption could be termed maximum consumption. The Vmax of the reaction canthen becalculated for each sample that results inamaximumconsumption value by dividing the consump-tion term by the transport time. In this study, a Vmax for methane oxidation at the contaminated site (0.4 ,uM/day) wascalculated by using this approach.

90l 70 60 _E IV 0 10 20 30 40

FIG. 6. Model simulation(curves) and observed relative

concen-trations (data points) of the breakthrough of chloride (A) and

methane(B andC)atthetracersitewithin the contaminant plumeat

9.7mbelowthe land surface. PanelsB andC show the sensitivity of

the model simulation to different values for Km (micromolar) and Vmax(micromolarperday).

50 40 30 20 10 0 _ 0.2 0.6 1.0 1.4 1.8 2.2

Vmox (u.M day1)

FIG. 7. Best-fitrelationship between Km and Vmax for methane oxidation as simulated by the transport model for the methane

breakthrough curve shown in Fig. 2. The solid line is a linear

regression of theKm values yielding the lowestroot mean square

error(RMSE) for each given Vmax. The dashed linesareequal-value

contoursof therootmean square error.

6 C) C-) C-RMSE 0.6 ,0.5 // 0.54 0.4 / / I 0.4 // / / 0.5

///

/ /7 s /7 77l 7/K~z /7K~~~ -/7 K -1.1

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The interpretation of these types of tracer studies in groundwater systems can be significantly augmented by applying an advection-dispersion transport model that also accountsfor substrateconsumption by microbial processes to simulate the tracer test results. The model used in this study calculates the amount of methane consumed during discrete time intervals(0.1 day)basedonMichaelis-Menten kinetics, and it calculates dilution of the methane in the tracercloud because of dispersion. When this approach is used,the model caneffectively simulate the methane break-through curves (Fig. 6). Although the modeling approach doesnotallow foraunique solution for both kinetic param-eters,ifthere isinformationtodetermineoneofthe param-etervalues independently, the model can be used to deter-mine the other value. Then the rateofmethaneoxidation can bepredictedatanypointwithin the transport interval. As we demonstrate above, it is possible to determine the

V.max

value ofaprocessand then use the model to determine the Km.Thismakes the tracertest-model combination a power-ful toolforstudyingmicrobial processes in situ in a nondis-ruptive manner.

Although the model does not account for microbial growth, this factor didnotappear tobe necessary to model methane oxidation within the context of this study. More than95% ofthemicrobial populationin theCape Cod aquifer isattachedtosolid surfaces (12) and is presumed to remain immobile as the tracer cloud passes by it. In general, the contact time with the tracer was short, which tended to

diminish growth as afactor, especially when methane was the substrate andwaspresentonlyatrelatively low concen-trations.Thelongestcontacttime with the tracer was for the

population

at the well fence (longer transport time means greaterdispersion). In thecaseofthecontaminatedsite,the maximum contact time was 16 days (Fig. 2); however, the methane concentration exceeded the calculatedKm (9 ,LM) for only 4 days. For comparison, when the frequency of

dividing

cells was used as an instantaneous estimate, the generation times oftheunattached bacteriain the

contami-nant plume were up to 5.8 days (10). Obviously, there are many other substrates andevenothersituationswith meth-ane as the tracer for which growth would be a significant factor and could not be ignored. Thus, incorporation of growth kinetics into the transport equation should be

con-sidered forfutureapplications ofthis model.

Methaneoxidationhasbeensuggestedas asuitable mech-anism for in situ bioremediation ofgroundwater

contam-inated with halogenated aliphatic compounds (35). Indeed, there are currently bioremediation efforts in progress that areattemptingtoexploitthis process foraquiferrestoration (25, 27). Therefore, it is pertinentto notethat themicrobial

populations

in both an uncontaminated

portion

and a con-taminated portion ofthe Cape Cod aquifer possessed the

ability

to oxidize methane, even

though

it is not

naturally

present inthegroundwater. If this occurrence wereindeed widespread, this predisposition for methane oxidation in groundwatercouldbean

important

factorwhen the

activity

isconsidered as one ofthe

options

for in situ remediation. Thepotentialformethanotrophic activitywithin the

aquifer

isprobablyenhancedby alimited

availability

of

degradable

organic substrates,

even within the center of the contami-nantplume (2). Hence,

heterotrophic activity

is

likely

tobe low comparedwith many other habitats(28, 30), which isa situationthat tendstofavor the establishment of

autotrophic

and methylotrophic populations.

Quantifyingandcharacterizingratesof methane oxidation in aquatichabitats aredifficult tasks. For thisstudy, tracer

methane concentrationswere

sufficiently

low to

preclude

a Vmax determinationattheaerobic siteby usingthe maximum

consumption

method

(Fig. 5). Consequently,

the Vmaxvalue forthecontaminated sitewasused

(somewhat arbitrarily)

to determinetheKmfor theuncontaminatedsite. The

similarity

of the methane

consumption

curvessuggeststhat therewas not a large difference in the

methanotrophic

populations

between the contaminated anduncontaminated

sites,

despite

a50-fold difference in the total unattached microbial popu-lation between thetwosites(12,

30).

Rudd and Hamilton

(26)

foundaKm for aerobic methane oxidation inaCanadian lake (4.7

p,M)

that was similar tothe aerobic Km found in this

study

(6.0 ,uM),buttheVmaxwasmorethan 10-fold

higher

in the lake(7.7 versus0.4 ,uM

day-'),

probably reflecting

the factthat methanewas

naturally

presentwithin the lake. The Kmformethaneoxidationinlandfill soilswas2.5to6.2

,uM

(34),while twostudies

involving

surface sedimentsofLake

Washington

found valuesof8.3 to 10.7 ,M (21) and 5.1 to 10.0 ,uM (16). These values agree well with the values obtainedinour

groundwater study (Table 3).

Thereis very little information available

concerning

thekineticsof anoxic methane oxidation. Iversen etal.

(15) reported

afirst-order

rate constantof 0.06

day-'

forthewatercolumnofasaline

lake,

which is similar to the value found in this

study

(VmaxlKm

= 0.05

day-').

Anoxic methane oxidation hasbeen the focus ofseveral recent

investigations

(14,

15,

20, 24). There is

growing

evidence that this process occurs in surface watersystems

(marine

and

freshwater),

both in sediments and in water columns.

Although

it has not been

directly

proven, sulfate reduction is often considered to be the

electron-accepting

stepin these habitats

(5, 14).

Littleattentionhasbeen

given

tonitrateas a

potential

electronacceptorforanoxicmethane

oxidation,

even

though

itwould haveanenergy

yield

that is morethananorderof

magnitude higher

than thatofsulfate reduction.

However,

therearereportsofmethaneoxidation

coupled

todenitrification in thewastewater treatment liter-ature

(32).

Although

the electron acceptor could not be ascertained inthis

groundwater study,

nitrate seemstobea

likely candidate, given

its

availability

and energy

yield.

Denitrification does occur at the contaminated tracer site and within the

depth

interval that the tracer was

injected

(31).

The

denitrifying

zonealso containssulfate

(40 ,uM)

and soluble iron (2

,uM)

but no detectable

hydrogen

sulfide or

ferrous sulfide and no iron

oxyhydroxides (22). Hence,

sulfate reduction and iron reduction do not appear to be involved.

In summary,small-scale

natural-gradient

tracer testswere

used to measure methane oxidation at two locations in a

sand and

gravel

aquifer,

a

pristine

oxic location and an

anoxic zone of a contaminant

plume.

Methane-oxidizing

activity

wasevidentatboth siteseven

though

methanewas

not

naturally

present in the

groundwater.

The amount of methane consumed was

directly

related to the methane

concentration,

whichwasalso

strongly

influenced

by

disper-sion

during

the transport process. The kinetics of methane oxidationwerecalculated

by

using

aniterative

groundwater

advection-dispersion

model thatcontained a methane

con-sumption

term based on Michaelis-Menten kinetics. Our results suggest that methaneoxidation could be

exploited

as

a bioremediation tool in this

aquifer,

if it were

warranted,

andatthesametime

they

demonstrate the

utility

of

natural-gradient

tracer testsin

characterizing

microbialprocessesin situ in

groundwater

environments.

on January 17, 2021 by guest

http://aem.asm.org/

(8)

ACKNOWLEDGMENTS

Wethank D. LeBlanc and R. Harvey for valuable discussions and assistance in equipping the tracer test sites and J. Duff, M. Ceazan, and M. Brooks for field and analytical assistance. We also thank R. Harvey and D. Lovleyfor helpful reviews.

This study was funded by the U.S. Geological Survey Toxic WasteGroundwater Contamination Program.

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