Environ. Sci. Technol. 1998, 32, 1244-1252
Removal of Atmospheric CCl4 under
Bulk Aerobic Conditions in
Groundwater and Soils
JAMES D. HAPPELL* AND
DOUGLAS W. R. WALLACE
Department of Applied Science, Oceanographic and
Atmospheric Sciences Division, Brookhaven National
Laboratory, Upton, New York 11973-5000
Measured concentrations of relatively nonreactive, anthropogenic halocarbon tracers (CFC-11, CFC-12, CFC-113)
were used to infer the time since recharge, or age, of
groundwater collected from the Upper Glacial and Magothy
Aquifers underlying Brookhaven National Laboratory on
Long Island, NY. On the basis of the reconstructed historical
atmospheric concentrations of CCl4, the initial CCl4
concentration for the precipitation that recharged the aquifer
was estimated as a function of age. Correlation of
measured and estimated initial CCl4 concentrations within
the aquifer, over inferred ages of 0-50 yr, suggested that
CCl4 was being removed in situ with a half-life of 14 ( 4 yr.
Groundwater samples collected at the water table had
CCl4 concentrations that were e50% of equilibrium with
contemporary atmospheric concentrations, suggesting
that removal was also significant in the unsaturated zone.
Soil gas profiles confirmed that atmospheric CCl4 was
being removed from the unsaturated zone, with only ∼25%
of the initial CCl4 being present in the gas phase at a depth
of 30 cm, and with no evidence for removal of CFC-11,
CFC-12, or CFC-113. A time-series of soil gas profiles collected before and after a major rainfall event indicated
that most removal occurred in the top 15 cm of soil. The
flux of CCl4 into the soil was estimated to be ∼8600 (
5100 pmol m-2 d-1, and removal of CCl4 in soils therefore
has the potential to significantly affect the global
atmospheric lifetime of this compound. The observed
degradation in bulk aerobic environments raises
questions concerning the conventional wisdom that CCl4
is degraded significantly only within reducing environments.
and field (20-23) experiments, there are many studies that
conclude that microbially mediated aerobic degradation of
perchlorinated compounds, such as CCl4, does not occur
(18, 20, 22, 23). On the other hand, Castro (24) reports that
two types of bacteria (a Pseudomonas and a CH4 oxidizer),
normally considered aerobic oxidizers, are able to reductively
dehalogenate CCl4 at a rate faster than an anaerobic
methanogen. Additionally, CCl4 distributions in the warm
(>10 °C) ocean surface waters (25-27) and some colder
waters (28) show very clear evidence for the degradation of
background levels of CCl4 at rates far greater than expected
for hydrolysis, even in well-oxygenated waters.
In this paper, we examine the distribution of background
levels of CCl4 in groundwater and soil gas and compare these
distributions with those of closely related anthropogenic
compounds that are much less reactive and that can be used
to provide information concerning the kinetics of CCl4
degradation. Notably we use the measured concentrations
of CFC-11 (CCl3F), CFC-12 (CCl2F2), and CFC-113 (CCl2FCClF2) as nonreactive analogues of CCl4 in soil gas and as
a tool to estimate the time since recharge of groundwater
(see below). TCA (CH3CCl3) concentrations were also
measured as a part of this study. By background levels, we
mean the concentrations found within the atmosphere as a
result of global anthropogenic releases. CCl4 has been
released to the atmosphere as a result of its widespread use
as a solvent, particularly during the first half of the 20th
century. Studies of the deep ocean, which to some extent
records past atmospheric concentrations, suggest that, prior
to the anthropogenic releases of this century, the atmospheric
mole fraction of CCl4 was <0.1 × 10-12 (26), which can be
compared with atmospheric mole fractions, measured at the
time of this study of ∼100 × 10-12. After its harmful properties
became recognized, CCl4 release declined but did not cease
entirely because of industrial uses, particularly as a chemical
feedstock for chlorofluorocarbon (CFC) production. Since
the introduction of the Montreal Protocol, even these releases
have diminished greatly, and the atmospheric burden of CCl4
has recently started to decline (29) as a result of environmental
sinks. The atmospheric lifetime of this compound is
estimated to be 42 yr (29), and the major sink is thought to
be photolysis in the stratosphere. However the possibility
of significant oceanic sinks has more recently been recognized
(25-27), and loss of CCl4 from air within a termite mound
was suggested as evidence for sinks in the terrestrial
environment (30). In this paper, we show data that suggest
that degradation within aerobic soils may indeed be a
significant global sink.
Methods and Techniques
Introduction
There are several mechanisms for the environmental degradation of CCl4 including (a) hydrolysis (1, 2); (b) abiotic
reductive dehalogenation to CHCl3, which occurs in the
presence of iron under anaerobic and aerobic conditions (3,
4); and (c) biologically mediated fortuitous reductive dehalogenation by methanogenic bacteria (5, 6), sulfate-reducing
bacteria (7), or denitrifying bacteria (7-12). While many
halocarbons, including TCA, can be microbially degraded
under aerobic conditions as shown by laboratory (13-19)
* Corresponding author present address: University of Miami,
Rosenstiel School of Marine and Atmospheric Science, Tritium
Laboratory, 4600 Rickenbacker Cswy, Miami, FL 33149-1098; Phone:
305-361-4111; fax: 305-361-4112; e-mail: [email protected].
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Study Site and Hydrogeological Setting. Brookhaven National Laboratory (BNL) is located in the center of Long Island
about 100 km east of New York City (Figure 1). Bedrock is
near the surface toward the north shore of Long Island and
dips to both the north and south forming a groundwater
flow divide ∼1 km north of BNL. Groundwater typically flows
from north to south over the site. Deep recharge, due to a
relatively fast vertical recharge rate, occurs near the groundwater divide. Vertical flow rates decrease moving south of
the divide.
Overlying the bedrock to the south are three different
aquifer units. The deepest is the ∼160 m thick Lloyd Aquifer,
which consists of fine to coarse sand, some gravel, and clay
lenses. The top confining unit for the Lloyd is an ∼100 m
thick layer of Raritan clay. On top of the clay is the ∼300 m
thick Magothy Aquifer. This aquifer contains mostly fine to
S0013-936X(97)00653-6 CCC: $15.00
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Published on Web 03/24/1998
FIGURE 1. Map showing the location of Brookhaven National Laboratory on Long Island, NY. The bottom panel indicates the position
of the groundwater divide in relation to BNL and gives groundwater elevation contours.
VOL. 32, NO. 9, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1245
are not degraded in the unsaturated zone, the soil air
concentration will be same as the tropospheric concentration
for unsaturated zones <10 m thick (39). The source function
for the recharging precipitation (Figure 2b) can be derived
from the atmospheric source function as follows:
Cp ) CaF
FIGURE 2. (A) Atmospheric concentrations of CFC-11, CFC-12, CFC113, TCA, and CCl4 from 1900 to 1995. Concentrations have been
measured since 1975. Before that time, the atmospheric concentrations are estimated based on industry production and release
estimates. CFC-12, CFC-11, and TCA were first released to the
atmosphere at the end of World War II. CFC-113 was first released
in the early 1970s, while CCl4 was released around 1910, much
earlier than the other halocarbon tracers. (B) Source functions
derived for recharging precipitation at 11 °C (see eq 1)
medium sand, some clay, layers of coarse sand, and gravel.
The Magothy is, in places, confined by a overlying clay layer.
The three previous units are of Cretaceous origin, and
overlying them is the Upper Glacial Aquifer. This aquifer is
a Pleistocene glacial deposit, consisting of till in the north
and outwash in the south. The till has clay, sand, gravel, and
boulders; the outwash has fine to very coarse quartz sand
and gravel. In this study, groundwater samples were obtained
from the Upper Glacial and Magothy Aquifers.
The unsaturated zone above the Upper Glacial Aquifer in
the study area consists of fine to very coarse quartz sand and
gravel ranging from 1 to 25 m thick and is covered mainly
by pine barren forest (pines, scrub oaks, blueberries, etc.).
Groundwater Dating. The technique of using CFC-11
(CCl3F), CFC-12 (CCl2F2), and CFC-113 (CCl2FCClF2) as
transient tracers to derive estimates of seawater ventilation
ages is well established (e.g., refs 26 and 31-33). The use of
these same tracers to derive estimates of groundwater
recharge age is becoming increasingly common (34-38).
CFC-11, CFC-12, CFC-113, TCA, and CCl4 have no known
significant natural sources. Industrial release records (pre1975) and atmospheric measurements (post-1975) provide
relatively accurate source functions for these compounds
(Figure 2a). The tropospheric concentration of these compounds has increased since their introduction, although
recently the concentrations of CFC-11, TCA, and CCl4 have
started to decrease (29).
Recharging precipitation equilibrates with soil air at the
base of the unsaturated zone. Assuming that the halocarbons
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(1)
where Ca is the atmospheric concentration, Cp is the
concentration of the halocarbon in the recharging precipitation, and F is the appropriate recharge temperature-dependent solubility coefficient (27, 40-42). We assigned a
recharge temperature of 11 ( 2 °C to all samples. This
recharge temperature was estimated from in situ water
temperatures, mean annual air temperatures, and 4He and
Ne measurements made in a subset of 11 wells. The
measured concentration for each CFC in a water sample was
then matched against the corresponding precipitation input
history (Figure 2b) to obtain the sample recharge age.
Obtaining a valid CFC-derived recharge age is based on
the following assumptions: (a) The soil air CFC concentration
is the same as the tropospheric concentration (see above).
(b) The CFC concentrations in groundwater are not significantly affected by hydrodynamic dispersion. This has been
shown to be true for sandy aquifers (34-38). (c) The CFC
tracers are conservative in groundwater (34-38). (d) The
temperature of recharge is known. In waters recharged before
1975, an uncertainty of 2 deg results in an uncertainty in the
age of less than 1 yr. In waters recharged after 1975,
uncertainties in ages resulting from uncertainties in recharge
temperature can range from 1 to 3 yr (43). (e) The aquifer
has not been contaminated by local near-surface sources of
CFCs.
The CFC-derived recharge age for each sample in this
study was determined using the recharge age derived from
various combinations of CFC-11, CFC-12, and CFC-113.
Various combinations of CFC were used for different samples
because, while there were many occasions on which derived
ages from all three compounds were in excellent agreement,
concentrations of individual CFCs in excess of the maximum
anticipated concentration (the concentration in water at 11
°C equilibrated with the maximum tropospheric concentration) were sometimes observed.
Estimation of “Missing” CCl4. The reactivity of CCl4 in
groundwater precludes its use as a dating tool. However a
comparison of the age derived from the nonreactive compounds with the observed concentrations of CCl4 can be
used to determine the kinetics of CCl4 removal within an
aquifer based on what is, in effect, a multi-decade in situ
incubation experiment. To interpret the CCl4 concentrations,
it is first necessary to construct a ‘precipitation input function’
that describes the initial concentration of CCl4 in the
precipitation that has recharged the aquifer over the past
40-50 yr. This is an identical process to that described above
for the age tracers. The absolute amount of degradation
that has taken place within the unsaturated zone and the
aquifer following infiltration is given by the difference
between the initial concentration as calculated from the
precipitation input function and the tracer-derived age and
the concentration measured in the groundwater sample.
Sampling and Analysis
Groundwater. Groundwater was pumped from monitoring
wells with a Grundfos Rediflow 2 submersible centrifugal
pump (Grundfos Pumps Corp., Clovis, CA) using standard
well sampling procedures (at least 3 casing volumes of water
removed before sampling). Between April 1995 and June
1996, 98 wells were sampled (depth range of 1-75 m below
water table). Ninety-four of these wells were located on the
BNL site, with the other four wells located within 4 km of the
site boundary. At least 3 casing volumes of water were purged
from each well at flow rates between 10 and 20 L min-1. Each
syringe was filled and emptied four times, and on the fifth
fill the syringe was capped and stored in a bucket of water.
The entire filling and flushing procedure was done with the
end of the syringe in the outflow tubing to prevent the water
from coming in contact with air. Three, and most times
four, replicate samples were taken from each well.
It should be noted that the maximum concentrations of
CFC-11, CFC-12, CFC-113, TCA, and CCl4 in groundwater
equilibrated with the appropriate maximum atmospheric
concentration at 11 °C (hereafter referred to as maximum
anticipated concentrations) are ∼107 times lower than the
New York State drinking water standards for these compounds. When working with samples at such low concentrations, potential sources of sample contamination are
always a concern. Therefore, the submersible pump used
for groundwater sampling was checked for halocarbon
contamination, prior to and twice during the study period,
by sampling a deep artesian well that draws water from the
Lloyd Aquifer. The water from this well is several thousand
years old and should therefore not contain any anthropogenic
halocarbons. Samples of water taken from this well without
the use of the pump contained nondetectable (<0.005 pmol
kg-1) amounts of CFC-12, CFC-11, CFC-113, TCA, and CCl4.
All samples taken from this well using the pump also
contained nondetectable amounts of halocarbons, indicating
that the pump and sampling apparatus introduced no
significant contamination into the samples.
Soil Gas. Soil gas samples were obtained using a 185 cm
length of 1/4 in. o. d. stainless tubing. One end of the tubing
was pinched closed and filed to a point. Eight small holes
were drilled into the 2-cm interval above the point, so samples
were drawn from a 2 cm depth range. The sampling probe
could be inserted to various depths up to 180 cm. A depth
profile was generally taken without removing the probe from
the ground by starting with the shallowest sample and
inserting the probe successively deeper. Occasionally the
probe had to be moved to a different hole (within 30 cm of
original) because of impenetrable obstacles in the original
hole.
After the probe was inserted to the depth to be sampled,
an Air Cadet pump (Cole Parmer Instrument Co., Chicago,
IL) connected to the probe with 1/8 in. o. d. stainless steel
tubing was used to withdraw the soil gas at rates between
200 and 300 mL min-1. The outlet side of the pump was
connected to a “T” fitting. One of the two remaining ports
on the “T” could be attached to a 100-mL ground glass syringe,
and the remaining port was attached to 2 m of 1/8 in. stainless
tubing. This arrangement, along with a three-way polypropylene stopcock on the syringe, allowed filling and flushing
of the syringe without any interruption of the soil gas flowing
through the probe and pump or having to detach the syringe
from the pump. The syringe was connected to the “T”
immediately after pumping commenced and was flushed
four times before a sample was taken. The filled syringes
were stored and analyzed for halocarbons as described above.
Two to four samples of atmospheric air for halocarbon
analysis were collected in syringes ∼1 m above the ground
at every site where a soil gas profile was obtained.
Soil gas samples for O2 analysis were field-transferred from
the syringes into evacuated 50-mL serum vials with butyl
rubber stoppers by attaching a needle to the syringe and
piercing the stopper. The vials were pressurized to ∼2 atm.
Analysis of the samples for O2 + Ar (Ar coeluted with O2) was
performed at Florida State University’s Department of
Oceanography by gas chromatography with thermal conductivity detection.
All samples (water and gas) were analyzed for halocarbons
within several hours of collection on a purge-and-trap
FIGURE 3. CFC-derived recharge age vs sample depth below the
water table. The depth error bars represent the screen length. The
age error bars are the standard error of the average age from four
replicate samples. A linear regression (age dependent on depth)
of these data yields an average vertical velocity of 2.1 ( 0.2 m yr-1
(( standard error of slope).
capillary column gas chromatograph with electron capture
detection (44). The limits of detection for the five halocarbons
measured in this study ranged between 0.005 and 0.010 pmol
kg-1. A gas-phase standard, prepared at BNL (45), was used
for calibration purposes, and all halocarbon concentrations
are reported on the Scripps Institution of Oceanography (SIO)
1993 absolute calibration scale.
Results
Distribution of Groundwater Ages. Figure 3 presents a
scatter plot of groundwater age as a function of depth below
the water table, showing that, as expected, older water was
sampled from wells with greater screen depths. Water as
old as 50 yr was sampled, which marks the effective limit of
groundwater dating using halocarbons as water older than
this would have undetectable halocarbon levels. There is
considerable scatter evident in the diagram, some of which
undoubtedly reflects real age vs depth variability within the
aquifer. Some of this variability may arise from the presence
of discontinuous barriers to vertical flow within the aquifer
(such as clay layers, etc.). Some of the scatter is the result
of wells being screened at variable distances from the
groundwater divide, which is located at the northern end of
the site, where vertical flow is presumably strongest (Figure
1). Other wells are located in the vicinity of the headwaters
of the Peconic River, where groundwater flow lines tend to
converge.
An effective mean vertical flow rate of 2.1 ( 0.2 m yr-1 for
the entire sampling area can be derived from a linear
regression of age versus depth. The quoted uncertainty is
the standard error of the slope of the regression line. This
can be compared with the mean annual recharge of 60 cm
yr-1 (1/2 mean annual rainfall) and a porosity of 0.30 (T.
Burke, personal communication), which would lead to a
maximum vertical flow rate of 2.0 m yr-1.
Missing CCl4. In all of the groundwater samples there
was less CCl4 measured than was inferred for the estimated
initial CCl4 content of precipitation based on the groundwater
age. In waters older than 20 yr there was <10% of the
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FIGURE 4. Plot of the percent CCl4 remaining in groundwater samples
vs the halocarbon-derived recharge age. Open symbols are for ages
<15 yr. Solid symbols are for ages >15 yr. Error bars are the standard
error of the average age from four replicate samples. The expected
CCl4 concentration in each sample (assuming no removal and input
only from the atmosphere) was calculated using the average
recharge age derived from various combinations of the CFC-12,
CFC-11, and CFC-113. The percent CCl4 remaining was then calculated
from the expected and the measured CCl4 concentrations. A CCl4
removal half-life of 14 ( 4 yr was calculated by fitting an exponential
equation of the form y ) ae-bx (a ) 24.2 ( 2.2, b ) 0.0484 ( 0.0135,
r 2 ) 0.29, n ) 98) to the data. This plot shows that after ∼20 yr only
10% of the expected amount of CCl4 in the groundwater remains.
The fact that there was never more than 50% CCl4 remaining
suggested that CCl4 was also being removed in the unsaturated
zone overlying the aquifer.
estimated initial CCl4 remaining. Although dissolved O2 was
not routinely measured as a part of this study, samples from
20 wells were analyzed for O2 using a standard Winkler
titration (46). These wells sampled water from 1 to 75 m
below the water table, and 17 of the 20 wells contained
detectable O2 with concentrations ranging from 3 to 360 µmol
kg-1 (1-104% saturated).
This indicates that CCl4 was being removed from a largely
aerobic aquifer in apparent contradiction of the finding by
laboratory studies that CCl4 degradation only takes place
under suboxic or anaerobic conditions. The correlation
between the missing CCl4, expressed as the percentage of
the initial CCl4 remaining within the aquifer, and the
groundwater age is presented in Figure 4. A CCl4 removal
half-life of 14 ( 4 yr (( standard error of slope of the regression
line) was calculated by fitting an exponential equation of the
form y ) ae-bx to these data. Alternatively a linear fit suggests
a half-life of ∼8 ( 1 yr. The plot shows clearly that after 20
yr, e10% of the initial amount of CCl4 remains in the
groundwater. Figure 4 also shows that, even in shallow wells
that contain recently recharged water, no sample contained
more than 50% of its estimated initial CCl4 content. This
strongly suggests that CCl4 was also being removed in the
unsaturated zone, after infiltration, but before recharge of
the aquifer.
Evidence for slight removal of CFC-11 and TCA was seen
in some wells (data not shown). A half-life for CFC-11 was
determined to be 411 ( 80 yr, suggesting that it is stable
enough to be a tracer on decadal time scales. A half-life for
TCA could not be determined due to the many samples that
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FIGURE 5. Halocarbon soil gas depth profiles at (A) 38-01 on 9/10/96
and (B) 58-01 on 9/11/96. The legend in panel B also applies to panel
A. Both profiles were obtained from grassy areas. The depth to the
aquifer was 3.2 m at 38-01 and 3.4 m at 58-01. All values are relative
to the respective average air concentration from samples obtained
on the same day at the same site. The calculated diffusive flux of
atmospheric CCl4 and TCA into the soil (see eq 2), in pmol m-2 d-1,
were (A) 6126 and 9276 and (B) 4547 and 4218.
contained TCA in excess of the maximum anticipated
concentration as a result of additional sources such as spills
and landfill leacheates.
Soil Gas. The soil gas profiles indicated a rapid decrease
in CCl4 concentration over the top 30 cm of the soil relative
to the atmospheric concentration, with a relatively constant
concentration (∼25% of atmospheric value) being found
deeper in the soil (Figures 5a,b and 6a,d). In three out of
four of these profiles there were also noticeable decreases in
TCA concentrations with depth, although the relative decrease in TCA (∼50%) was smaller than that for CCl4 (∼75%).
CFC-12, CFC-11, and CFC-113 concentrations remained
relatively constant with depth at levels that were close to
measured atmospheric values. Oxygen concentrations in
the soil gas from 5 to 120 cm deep at sites 38-01 and 58-01
ranged from 19.2 to 21.9% (av ( SE ) 20.6 ( 0.2, n ) 13).
This again indicates that CCl4 and sometimes TCA were being
removed from the soil gas under conditions that were
predominantly aerobic. These findings are similar to those
reported by Khalil and Rasmussen (30), who observed TCA
and CCl4 removal from the air in termite mounds, with 25
and 50% of the respective halocarbons removed in the top
25 cm of the mounds.
Gases can be transported through the unsaturated zone
by diffusion and advection in both the gas and liquid phases.
Advection of soil air can occur because of changes in
atmospheric pressure, pressure changes due to wind, and
changes in soil temperature. Advection can also occur due
to changes in soil water content, which occur during
infiltration, or from changes in the water table height (36).
At site 26-02, we obtained a time-series of halocarbon soil
gas profiles before, during, and after a large (7-cm) rainfall
event (Figures 6a-e). Such a large infiltration would
presumably increase the halocarbon transport to the soil gas
due to the advective motion of water percolating through
the soil. Before the rain, CCl4 decreased with depth (Figure
6a) in a manner similar to our other profiles (Figure 5) showing
the effects of removal. Four samples obtained over the top
30 cm during the rain (Figure 6b) had CCl4 concentrations
that were significantly higher, within 5% of the atmospheric
CCl4 concentration, indicating increased supply and/or
reduced removal. Twelve hours after the rain had ended
(Figure 6c), a CCl4 minimum (70% atmospheric) was observed
at 10 cm with CCl4 concentrations below 30 cm within 2-3%
of atmospheric values. The appearance of a minimum
suggests rapid consumption of CCl4 within the upper 0-15
cm of the soil. Between 36 and 84 h after the end of the rain,
CCl4 depth profiles (Figure 6d,e) were similar to those
measured immediately before the rain started. TCA in the
soil gas behaved in a fashion similar to CCl4, although the
relative changes were smaller. CFC-12, CFC-11, and CFC113 concentration profiles exhibited relatively little response
to the rain (Figure 6a-e), except for a possible slight decrease
in CFC-11 to ∼80% of its atmospheric concentration at depths
greater than 30 cm, 36 h after the end of rainfall (Figure 6d).
Discussion
Groundwater and Soil Sinks. This study and others (38, 47,
48) indicate that CFC-derived recharge ages are excellent
tools for deriving model-independent in situ reaction rates.
While our data do not allow us to determine the mechanisms
involved, they do indicate relatively rapid removal of CCl4
from groundwater in Long Island’s Upper Glacial and
Magothy Aquifers under bulk aerobic conditions (half-life of
14 ( 4 yr). Ten aerobic groundwater samples obtained from
the karst region of northern Florida as part of this study also
indicated removal of CCl4, with 1-60% of the expected CCl4
remaining. This evidence from two different types of aquifers
suggests that in situ removal of CCl4 under bulk aerobic
conditions is likely to be common in groundwater.
Soil gas profiles also indicated the removal of atmospheric
CCl4 (Figures 5 and 6) under bulk aerobic conditions. Again,
our data do not allow us to determine the mechanisms
involved or whether similar removal mechanisms operate in
both the unsaturated and saturated zones. However the
relatively rapid reestablishment of CCl4 profiles after raininduced advective input (Figure 6a-e) and the evidence for
the most active removal zone being in the upper 15 cm (Figure
6c) are suggestive of a biological removal mechanism.
The diffusive flux of atmospheric CCl4 into the soil was
calculated from the soil gas profiles according to
(δC
δZ)
F ) Dgaτg
(2)
where F is the flux in pmol m-2 d-1, Dg is the free air diffusion
coefficient () 0.686 m2 d-1 at 10 °C), a is the soil porosity
() 0.20), τg is the tortuosity () 0.19), and δC/δZ is the CCl4
gradient (in pmol m-3 m-1) over the top 10 cm of soil. Dg,
a, and τg were estimated from Cook and Solomon (39). These
fluxes ranged from 4500 to 19 100 pmol m-2 d-1, with an
average flux of 8600 ( 5100 pmol m-2 d-1 ((1 SD). Removal
fluxes for TCA were calculated with the same equation (Dg
) 1.372 m2 d-1) and ranged from 4100 to 17 500 pmol m-2
d-1, with an average flux of 9300 ( 5300 pmol m-2 d-1((1
SD).
If removal of these compounds is a common feature in
soils and proceeds at a rate similar to that calculated from
the soil gas profiles in this localized study, a rough estimate
of the global removal of CCl4 by soils can be made. Using
the five biomes, biome areas, and season lengths (nonfrozen
days) given by Shorter et al. (14), a global CCl4 soil removal
rate of 27 kt yr-1 is estimated. On the basis of an atmospheric
burden of 2700 kt and an atmospheric lifetime of 42 yr (49),
the global degradation rate of CCl4 in 1995 was 64 kt yr-1.
This degradation is generally attributed entirely to photolysis
in the stratosphere. In a similar manner, the TCA soil removal
rate is estimated to be 26 kt yr-1. Using an atmospheric
burden of 2400 kt and a lifetime of 5 yr (49), the global
destruction rate for TCA was 480 kt yr-1 in 1995, with 85, 10,
and 5% of the loss attributed to reaction with tropospheric
OH, transport to the stratosphere, and the oceans, respectively. As has been done by Shorter et al. (14) for CH3Br, we
estimate the uncertainty on our global flux estimates for CCl4
and TCA to be (75% due to uncertainties in the dependence
of the soil flux estimates on biome type, biome areal extent,
soil moisture, soil organic content, and temperature. This
results in a global soil flux and uncertainty of 27 ( 20 kt yr-1
for CCl4 and 26 ( 19 kt yr-1 for TCA. Our data therefore
suggest that removal of atmospheric CCl4 and TCA by soils
may be a significant term in their global budgets, accounting
for roughly 40 ( 30% and 5 ( 4% of the global removal rates,
respectively. Even though the soil sink for TCA appears to
be relatively small on the basis of this limited study, it may
be important because the global mean OH radical concentration in the atmosphere, and hence the lifetimes of other
trace gases (HCFCs, HFCs, CH4) reacting with OH, are
deduced from the TCA atmospheric lifetime. Experiments
to determine the extent, rate, and mechanisms of CCl4 and
TCA removal by soils are needed in order to better constrain
this term in the global budgets of these anthropogenic gases.
Speculations on Removal Mechanisms. Adsorption to
the soil is an unlikely possibility for CCl4 removal because
the CFCs measured as part of this study would likely be
affected similarly by such a mechanism, yet there is no
evidence for their adsorption (Figures 5 and 6). Carbon
tetrachloride removal under bulk aerobic conditions may be
proceeding abiotically as has been shown to occur in the
laboratory in the presence of iron and sulfur compounds (4).
Tanhua et al. (25) have suggested that CCl4 degradation
observed in oxygenated Black Sea surface waters may be
catalyzed by transition metal complexes. They speculate
that the reduction of CCl4 is thermodynamically favorable
even in the presence of O2 because of its relatively high redox
potential. Supporting the possibility of reduction of CCl4
under aerobic conditions, Castro (24) found an aerobic
Pseudomonas bacterium and a CH4 oxidizing bacterium that
were able to reduce CCl4 at rates faster than an anaerobic
methanogen. Although we cannot provide quantitative data
for the CCl4 reduction product CHCl3 because our standard
did not contain CHCl3, we could identify the CHCl3 peak in
our chromatograms, and the peak area did increase with
depth. The increase in CHCl3 does lend some support to the
hypothesis that reductive processes are removing CCl4.
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FIGURE 6. Plot of halocarbons in soil gas at site 26-02 on (A) 9/16/96, 24 h before rain; (B) 9/17/96, during 7 cm rainfall; (C) 9/19/96, 12
h after end of rain; (D) 9/20/96, 36 h after rain; and (E) 9/22/96, 84 h after rain. The legend in panel B applies to all plots. Profiles A-D
were obtained from a grassy area, while profile E was obtained in a wooded area within 10 m of the others. The top of the aquifer was
located 1.5 m below the ground surface. All values are relative to the respective air concentrations in samples obtained on the same
day. The average concentration (in pmol mol-1 ( SE) of all air samples (23) obtained during this study was 580.6 ( 12.1 for CFC-12, 275.0
( 2.9 for CFC-11, 81.9 ( 0.6 for CFC-113, 97.5 ( 3.8 for TCA, and 105.8 ( 0.9 for CCl4. The calculated diffusive flux of CCl4 and TCA into
the soil (see eq 2) were, in pmol m-2 d-1 (A) 4993 and 9682, (B) 6169 and 5306, (C) 19130 and 17474, (D) 8811 and 15428, and (E) 10223 and
4142.
Alternatively, and even though the bulk environment of
the soils and groundwater sampled during this study was
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aerobic, there is a high likelihood that microbial activity within
suboxic or anoxic microzones is responsible for the CCl4
degradation. Such microenvironments may support denitrifying or methanogenic bacteria that have been shown to
degrade CCl4 in both the field and laboratory studies.
Degradation of CCl4 in soils, in contrast to the production
of CH4, for example, is an inherently irreversible process as
there are no bacteria available in the soil that can synthesize
CCl4. Hence, even though the population of denitrifying and
methanogic bacteria is likely to be relatively small in these
bulk aerobic environments, their activity may be sufficient
to degrade low concentrations of CCl4 such as those found
in the atmosphere. Similar anaerobic microzones may occur
in the ocean within suspended organic particles (50), and be
responsible for the observed oceanic degradation of CCl4.
How can our findings be reconciled with previous field
studies that found no evidence for CCl4 degradation within
oxic environments (18, 20, 22, 23)? All of these previous
studies examined CCl4 concentrations that were at least 5
orders of magnitude higher than the concentrations measured during this study. We hypothesize that the absolute
number and activity of microorganisms that are capable of
CCl4 degradation in these predominantly aerobic environments is limited, perhaps by the availability of anoxic
microzones, in which case the CCl4 degradation capacity
cannot simply be scaled up to degrade higher concentrations
of CCl4. Because of the slower diffusion of O2 in water as
compared to air, increased soil moisture increases the
likelihood of development of suboxic and anoxic microzones.
Supporting the hypothesis of removal in anoxic microzones,
the greatest change in CCl4 soil gas concentration over the
top 5 cm of soil, and hence the largest flux of CCl4 into the
soil, occurred 12 h after the end of a 7-cm rainfall (Figure 5c).
If this hypothesis is correct, then the absolute removal rate
of CCl4 inferred from our study is far too small to significantly
reduce the much higher levels of CCl4 encountered in the
prior field studies. Obviously more experiments and field
studies are required to determine the mechanism(s) and/or
organisms involved in CCl4 removal and to rule out the
possibility of a truly aerobic removal mechanism. Such
experiments would be critical for determining whether there
is any potential for the in situ removal of environmentally
objectionable quantities of CCl4 from bulk aerobic environments.
Acknowledgments
This work was supported by a grant from the Office of
Environmental Restoration (OER) at Brookhaven National
Laboratory and by the Department of Energy through
Contract DE-ACO2-76CH000016. Tom Burke of OER was
especially helpful in arranging access to the monitoring wells
and sampling equipment. Jeff Chanton (at Florida State
University) kindly analyzed soil gas samples for O2 + Ar.
Gerhard Bönisch and Peter Schlosser (at the Lamont Doherty
Earth Observatory) performed the 4He and Ne analyses. We
would like to thank Guangwei Che, Katherine Grimes, and
Jeffery Happell for assisting with field sampling and sample
analysis on Long Island. Candace Schwartz collected the
groundwater samples from northern Florida. The map was
provided by Micheal Scorea of the USGS.
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