FS

(ppm)
co
.

i) 6 12 18 24 30 36 42 48 54
Time (days)

13 B
tt
\
i) 8 16 24 32 40 48 56 64
Time (weeks)

Figure 1. Loss of 2,6-DCP in Cherokee Pond (A) and Bolton's
Pond (B) sediment. Arrows denote additions of 2,6-

dichlorophenol.

(under estimation of degradation or transformation rate),
approximation to the direct use of Monod kinetics.

Acknowledgements

We wish to acknowledge the useful and constructive
comments provided by Dr. N. Lee Wolfe and Dr. Susan A.
Moore in reviewing the manuscript

References

1. Alexander, M. 1982. Most-probable-number method for
microbial populations. /n Methods of soil analysis, Part
2. Chemical and microbiological properties -- Agronomy
monograph no. 9 (2nd Edition). ASA-SSSA, Madison
WI.

Battersby, N.S., and V. Wilson. 1989. Survey of the
anaerobic biodegradation of organic chemicals in
digesting sludge. Appl.Environ. Microbiol. 55:433-439.

Boyd, S.A., and D.R. Shelton. 1984. Anaerobic
biodegradation of chlorophenols in fresh and acclimated
sludge. Appl.Environ.Microbiol. 47:272-277.

Brown Jr., J.F., R.E. Wagner, H. Feng, D.L. Bedard,
M.J. Brennan, J.C. Carnahan, and R.J. May. 1987.
Environmental dechlorination of PCBs.
Environ.Toxicol.Chem. 6:579-593.

Caperon, J. 1967. Population growth in micro-organisms
limited by food supply. Ecology 48:715-722.

Genther, B.R.S., W.A. Price, and P.H. Pritchard. 1989.
Anaerobic degradation of chloroaromatic compounds in
aquatic sediments under a variety of enrichment
conditions. Appl.Environ.Microbiol. 55:1466-1471

Gibson, S.A., and J.M. Suflita. 1986. Extrapolation of
biodegradation results to groundwater aquifers:
reductive dehalogenation of aromatic compounds.
Appl.Environ.Microbiol. 52:681-688.

Healy, Jr., J.B., and L.Y. Young. 1978. Catechol and
phenol degradation by a methanogenic population of
bacteria. Appl.Environ.Microbiol. 35:216-218.

Horowitz, A., J.M. Suflita, and J.M. Tiedje. 1983.
Reductive dehalogenations of halobenzenes by
anaerobic lake sediment microorganisms
Appl-Environ.Microbiol. 45:1459- 1465.

Kuhn, E.P., and J.M. Suflita. 1989. Sequential reductive
dechlorination of chloroanilines by microorganisms from
a methanogenic aquifer. Environ. Sci.Technol. 23:848-
865.

Kuhn, E.P., J.M. Suflita, M.D. Rivera, and L.Y. Young.
1989. Influence of alternative electron acceptors on the
metabolic fate of hydroxybenzoate isomers in anoxic
aquifer slurries. Appl.Environ.Microbiol. 55:590-598.

Kuiper, J., and A.O. Hanstveit. 1984. Fate and effects of
4- chlorophenol and 2,4-dichlorophenol in marine
plankton communities in experimental enclosures
Ecotoxicol.Environ Saf. 8:15-33

Lewis, D.L., H.P. Kollig, and R.E. Hodson. 1986
Nutrient limitation and adaptation of microbial

populations to chemical transformations. Appl
Environ.Microbiol. 51:598-603.
O'Conner, O.A., M.D. Rivera, and L.Y. Young. 1989.

Toxicity and biodegradation of phthalic acid esters
under methanogenic conditions. Environ. Toxicol.Chem
8:569-576.

Robinson, J.A., and J.M. Tiedje. 1983. Nonlinear
estimation of Monod growth kinetic parameters from a
single substrate depletion curve. Appl.Environ
Microbiol. 45:1453-1458.

Schmidt, E., M. Hellwig, and H.-J. Knackmuss. 1986
Microbiological transformation kinetics of xenobiotics in
the aquatic environment. Appl. Environ.Microbiol
46:1038-1044


United States
Environmental Protection
Agency

Center for Environmental Research
Information
Cincinnati OH 45268

Official Business
Penalty for Private Use $300

EPA/600/M-90/003


Table 1. Persistence of 3,5-DCP in Fresh Water Sediment Slurries
Residual 3,5-Dichlorophenol Concentration (mg/L)

Initial
I

Source of — Concentration Time of incubation

Sediment (mg/L) Week Q _Week2 Week4 Week6 Week 8 Week 10 Week 14
Bar H 4.36 3.30 3.48 3.57 4.35 3.01 3.23 2.18
Pond 4.18 4.36 4.36 4.36 4.01 4.14 3.26
Cherokee 414 414 414 3.74 0 0 ti) ti)
Pond 4.14 4.14 4.09 0 o i) O
2-Boat 4.22 4.22 4.22 O to) 0 O i)
Pond 4.22 4.22 4.18 4.12 3.92 4.04 4.06
Sandy Creek 4.27 4.27 4.27 4.00 4.27 3.84 4.24 4.27
Nature Ctr. 4.27 4.27 3.74 4.27 4.01 4.24 3.48
Bolton's 4.18 3.78 4.18 2.87 0 0 0 0
Pond 4.18 4.18 4.18 4.18 4.18 4.18 3.41

Table 2. Persistence of Dichloropheno!l Isomers in Fresh Water Sediment Slurries

Residual Dichloropheno! Concentration (mg/L)

Initial
TT
Concentration ime of Incubation
lsomer (mg/L) Week 0 Week2 Week4 Week6 Week8 Week 10 Week 12

2,3- 4.57 4.57 3.60 0 0 t) 0 0

4.57 3.33 0) 0 O) 0) 0
24- 4.70 4.70 4.45 0 () C) 0 0)

4.70 475 0 0 to) 0 0
2,5- 4.91 491 3.96 4.30 0 to) i) i)

4.35 4.02 4.20 3.92 0 0 0
2.6- 4.12 2.85 4.02 4.04 i i) 0 0

3.03 3.25 4.12 3.35 0) 0) C)
3,4. 4.40 4.32 3.60 4.12 oO O 0 i)

4.40 3.71 4.02 3.98 4.02 3.32 4.02
3,5- 4.14 4.14 4.14 3.74 t) 0 0 0

4.14 414 4.09 0 0 0 0)

useful in investigating the dechlorination of chloroaromatic
compounds in anaerobic saturated zone water. The length
of the lag period in this case is dependent on the values of
K,, umax and X 9. The effects of increasing K, and

Table 3. Lag Times and ®Ts9 Values for PCP, 2,4- decreasing biomass are shown in Figures 3 and 4. The

D, and 2,4,5-T in Fresh Water Sediment effects of decreasing umax would be similar to the effects of
Slurries decreasing biomass. Factors affecting the lag period (such
as those described above) would presumably affect one or
Sediment Lag (days) _2T59 (days) ali of these kinetic parameters. Adaptation is represented by

PCP the theoretical curves in Figure 5

East River, NY 19 26

Lake Borek, USSR 14 15 Incorporating Monod kinetics into a subsurface transport
Cherokee Pond, GA >40 >40 and transformation model requires a supporting database.
24-D Unfortunately, such a database is currently not available.
Wacissa Spr., FL 9 16 This fact was recognized in 1988 with the publication of an
Lake Borek, USSR 50 55 anaerobic protocol (40 CFR, Part 795, Section 795.54) for
Cherokee Pond, GA 22 46 developing anaerobic degradation data for organic
2.4,5-T chemicals of interest in the subsurface environment. A
Cherokee Pond, GA 60 60 geometric sampling approach was recommended for use in

5 ~ providing data for rapidly and slowly degrading compounds.
@ Te 1s the time to observe a 50% decrease in con-

centration, and should not be mistaken for a hal Ife Samples were to be analyzed at 0, 4, 8, 16, 32, and 64

that 1s related to first-order kinetics (Moore et a weeks. The protocol would be used to determine the length

EPA600/3-89/080) of time (lag) before which detectable degradation could be

observed and the half-life of the chemical following the lag

period. Conceivably these two pieces of information could

be incorporated into a fate model as separate entities. The

methodology provided here is a crude, but conservative


——-}> % of Theoretical

——-> Time (a)

240

Figure 2. Loss of 2,4-dichlorobenzoate ( [] - [, 5 ppm) and the formation of 4-chlorobenzoate (0-0) in Cherokee Pond
sediment. Arrows denote additions of 2,4-dichlorobenzoate.

Table 4. Mean Number (Log, + Standard Deviation, n=5)
of 2,4-Dichlorophenol Dechlorinating Organisms in
Cherokee Pond Sediments Collected from Selected
Sites During Various Seasons

Sampling Date Te Tiogs? Sample Control
9/27/88 2.99 + 0.88 5.18 + 1.20 3.56 +1.01
1/24/89 3.62 +0.09 4544058 3.78 + 0.39
3/27/89 4.22 + 0.31 4.68 +0° 413 + 0.33
5/30/89 4.03 + 0.55 4.37 + 0.49 3.65 + 0.26

a Time =0

» Time = complete dechlorination of 2,4-dichlorophenol to
monochlorophenol
ens3

17. Shelton, D.R., and J.M. Shelton. 1984. General method
for determining anaerobic biodegradation potential
Appl.Environ.Microbiol. 47:850-857.

18. Simkins, S., and M. Alexander. 1984. Models for
mineralization kinetics with the variables of substrate
concentration and population density. Appl. Environ
Microbiol. 47 :1299-1306.

19. Sleat, R., and J.P. Robinson. 1983 Methanogenic
degradation of sodium benzoate in profundal sediments
from a small eutrophic lake. J.Gen. Microbiol. 129:141-
152.

X-0.01mg/L

5
ab
3 3F
=
ao b
ab
rb
0
0
Figure 3.

50 100

——— time (d)

Theoretical plot representing the effect on the
lag of increasing the half saturation constant K,.
So (4.6 mg/l); umax (0.15/d); X_ and K, (mg/l) as
indicated.


United States
Environmental Protection
Agency

Environmental
Research Laboratory
Athens, GA 30613

Research and Development

EPA/600/M-90/003 June 1990

“EPA

ENVIRONMENTAL

RESEARCH BRIEF

Degradation Kinetics of

Chlorinated Aromatic

Compounds in Saturated Subsurface Environments

John E. Rogers', Jacobus Struijs2, Dorothy D. Hale3, and Frank Bryants

Abstract

Results are presented to support the use of Monod kinetics
in describing the anaerobic degradation of chlorinated
aromatic compounds in the saturated subsurface
environment. For all compounds studied, a lag period was
observed before loss of the compound was detected.
However, subsequent additions of compound were
dechlorinated without a lag and at an increased rate. The
length of the lag was dependent on both the characteristics
of the fresh water sediment and the compound investigated.
Preliminary studies indicated that the population size of
dechlorinating organisms in some sediments can increase in
response to the removal of a single chlorine from the
chemical, 2,4-dichlorophenol.

Background

The regulation of hazardous and solid waste is continually
being refined as the understanding of the properties and
fates of chemical waste components increases. This
increased knowledge has resulted in a number of adopted
and proposed amendments to the Resource Conservation
and Recovery Act (RCRA) (PL 98-616). The Hazardous and
Solid Waste Amendments of 1984 directed EPA under
RCRA section 3001(h) to develop additional hazardous
waste characteristics for identifying hazardous wastes. In
June 1986, EPA proposed to modify the Toxicity

T Environmental Research Laboratory, Environmental Protection
Agency, Athens, GA 30613-7799

2 National Institute of Public Health & Environmental Hygiene,
Bilthoven, The Netherlands

3 Technology Applications, inc., c/o Environmental Research Laboratory,
U.S. Environmental Protection Agency, Athens, GA 30613-7799

Ds»

Characteristic. The existing characteristic uses the National
Interim Primary Drinking Water Standards (DWS) as toxicity
thresholds for listed individual chemical components of the
waste. These thresholds are, in turn, combined with a
generic dilution/attenuation factor (100 times) to determine
the regulatory threshold. The proposed characteristic would
use chronic toxicity reference levels, combined with a
compound-specific dilution/attenuation factor, to calculate
the regulatory threshold for individual waste toxicants in the
waste leachate. A subsurface (unsaturated and saturated
zone) fate model is used to calculate the dilution/
attenuation factors.

In the proposed rules, the model includes mathematical
equations that rely on compound-specific hydrolysis and
soil adsorption data; these equations are coupled to others
using parameters describing a wide range of subsurface
environments. The resulting model calculates the degree of
attenuation and dilution a compound would undergo as it
migrates to a subsurface drinking water source. Chemical
hydrolysis is the only transformation mechanism currently
considered in the proposed rule. Although the EPA
recognizes that biodegradation is an important
transformation process, it was considered insufficiently
understood to be included in the model at that time.

Research at EPA's Environmental Research Laboratory at
Athens, GA, has been aimed at identifying key kinetic
expressions that describe anaerobic degradation in the
saturated zone. The main emphasis of this research is to
increase our understanding of the kinetics of anaerobic
degradation to provide the scientific basis necessary for
reliably including biodegradation in the future regulation of
hazardous and solid wastes.



5
X9=0

4 \\ conan
23 00003
a
| 2

1

0

0 1 2 3 4 5

———> time (mo)

Figure 4. Theoretical plot indicating the effect of decreasing
bacterial population on lag. S, (4.6 mg/l); umax

(0.35/d); K, (8 mg/l); X, (mg/L) as indicated.

S(mgi!)

——— time (mo)

Figure 5. Theoretical plot indicating adaptation. S, (4.6 mg/L)
each addition; umax (0.15/d); K, (10 ma/l); X, (0.1
pg/!).

20.

21.

22.

23.

24.

25

26.

Smouse, P.E. 1980. Mathematical models for
continuous culture growth dynamics of mixed
populations subsisting on a heterogenous resource
base. 1. Simple competition. Theor.Popul.Biol. 17:16-
36.

Stephenson, T., J.N. Lester, and R. Perry. 1984.
Acclimation to nitrilotriacetic acid in the activated sludge
process. Chemosphere 13:1033-1040.

Stewart, F.M., and B.R. Levin. 1973. Partitioning of
resources and the outcome of interspecific competition:
a model and some general considerations. Am.Nat.
107:107-198.

Struijs, J., and J.E. Rogers. 1989. Reductive
dehalogenation of dichloroanilines by anaerobic
microorganisms in fresh and dichlorophenol-acclimated
pond sediment. Appl.Environ.Microbiol. 55:2527-2531.

Vashon, R.D., W.J. Jones, and A.G. Payne. 1982. The
effect of water hardness on nitrilotriacetate removal and
microbial acclimation in activated sludge. Water Res.
16:1429-1432.

Walker, R.L., and A.S. Newman. 1956. Microbial
decomposition of 2,4-dichlorophenoxyacetic acid.
Appl.Microbiol. 4:201-206.

Wiggins, B.A., S.H. Jones, and M. Alexander. 1987.
Explanations for the acclimation period preceding the
mineralization of organic chemicals in aquatic
environments. Appl.Environ.Microbiol. 53:791-796.


Laboratory Procedures
Collection and Treatment of Samples

Sediment and water samples were collected from five ponds
near Athens, GA, in September 1986. Bolton's Pond and
Cherokee Trailer Park Pond were sampled periodically over
the 3 year period beginning September 1986. The samples
were collected and treated as follows. Sterile Mason jars
were filled to capacity with sediment (0-10 cm) and
overlying water and capped near the sediment/water
interface. Additional site water was collected near the
sediment surface in sterile 1-L Erlenmeyer flasks. Both
water and sediment samples then were transported to the
laboratory and placed in an anaerobic glovebox that
maintained an atmosphere of 25%N9:5%Hb>. All further
Manipulations of the water and sediment were conducted in
the chamber. Sediments were washed with site water
through a 1-mm sieve (U.S. Standard Testing Sieve No.18)
to remove organic debris and stones. Sediments
subsequently were stored in crystallizing dishes for 3 weeks
before being used in serum bottle microcosms. Such
treatment allowed removal of residual oxygen from the
sediments and restoration of methanogenic activity

Transformation Assays

Wet sediment containing the equivalent of 10 g of dry
sediment weight was added to individual 125-ml serum
bottles. Sufficient site water then was added to bring the
final volume to 100 ml. One-to-five milliliters of an aqueous
stock solution (200 ppm) of chlorinated substrate was added
to separate reaction vessels to yield a final concentration of
2 to 10 ppm. Bottles were caoped with butyl rubber
stoppers, crimp sealed, and incubated in the dark in the
anaerobic chamber at 25°C. The loss of the substrate then
was followed over time. At specific intervals, 1.0-mi
subsamples were removed and combined with 1.0 ml of
hexane (PCBs), pentane (chlorinated benzenes), acetonitrile
(chlorinated anilines, phenols and benzoates), or ethanol
(PCP, 2,4-D and 2,4,5-T) to terminate biological activity and
to dissolve any test chemical sorbed to the sediment. The
subsamples combined with acetonitrile or ethanol were
centrifuged at 3500 rpm (IEC HN-S CENTRIFUGE) and
subsequently filtered through 0.22-um filters before analysis
or storage (4°C). Subsamples treated with hexane and
pentane were also centrifuged, with the solvent phase
subsequently analyzed for residual test chemicals.

Autoclaved sediment slurries served as sterile controls. In
most cases, sediments were autoclaved (Sybron/Castle
3020) at 120°C (1.4 atm) for 30 minutes on 3 consecutive
days. Some sediment samples, however, were autoclaved
only once for 30 minutes before the addition of a chlorinated
substrate. The more rigorous autoclaving changes the
sediment properties and causes the production of
compounds that interfere with the chromatographic
determination of some chlorinated substrates and products.

HPLC Analysis

Chlorinated substrates in sediment slurries were quantitated
as follows. Sediment slurries mixed with equal volumes of
acetonitrile or ethanol were centrifuged for 10 minutes at
3500 rpm (IEC HN-S CENTRIFUGE). The supernatant
solutions were filtered (0.22 um, Millipore, GVWP) before
analysis by reversed phase HPLC. The chromatographic
system consisted of a Rainin pump system coupled to a C-

18 Dynamax Microsorb column (0.46 x 25 cm), a Knauer UV
absorbence detector operated at 280 nm (chlorinated
phenols, anilines and benzoates, and 2,4-D) or 290 nm (PCP
and 2,4,5-T) and a Shimadzu C-R3A integrator. The
chromatography solvent for chlorinated phenols and
benzoates and 2,4-D was methanol:water:acetic acid
(60:38:2 v/v/v). For chlorinated anilines, the solvent
composition was 70:28:2 (v/v/v). The chromatographic
solvent for PCP and 2,4,5-T was acetonitrile:water:acetic
acid (60:38:2 v/v/v). Residual substrates and Products were
identified by comparing their retention times with those of
authentic standards.

Gas Chromatograph Analysis

Concentrations of chlorobenzenes were quantified by gas
chromatography. A 1.0-m! sediment slurry sample was
added to 1.0 ml of n-pentane containing 5 ppm lindane as
an internal standard. The mixture was vigorously mixed for
30 seconds and then centrifuged at 3900 X g. A 5-pl sample
from the solvent phase was injected into an HP 5890 gas
chromatograph connected to an HP 3399A integrator. The
temperature of the injection port was 270°C and the
temperature of the ECD detector was 300°C. Nitrogen was
used as the carrier and the auxiliary gas. An OV-! column
(30 m X 0.25 mm with 0.5 um film) was used The column
temperature was maintained under isothermal conditions at
180°C for quantification of substrate. A programed
temperature gradient (initial temperature of 40°C for 4
minutes, followed by 10°C/min increase to 180°C which
was held for 15 minutes) was used in the identification of
degradation intermediates.

Identification of Products

The identities of intermediate metabolites were established
by using a combination of co-chromatography (HPLC and
GC) with authentic standards and gas chromatography-mass
spectrometry (GC-MS). GC-MS was used to confirm
molecular weights and numbers of chlorine substituents.
Analyses were conducted on iso-octane extracts of
sediment slurries. Sediment slurry samples (5 ml) were
mixed with 1 ml of isooctane, and after centrifuging (3500
rpm), the isooctane layer was separated from the water
phase and used without further Purification. The extracts
were analyzed with a Finnigan 4500 GC-MS, interfaced to
the Finnigan Incos data system. The gas chromatograph
was equipped with a DB-5 30-m x 0.25-mm capillary
column.

Most Probable Number (MPN) Determinations

Numbers of 2,4-dichlorophenol (DCP) dechlorinating
microorganisms in pond sediments were estimated with
most probable number (MPN) techniques (1). The dilution
and incubation medium was prepared by adding 2,4-DCP to
sterile pond water to give a final concentration of 10 mg/L.
Sterile pond water was prepared by filtration through 0.22-
um membranes followed by autoclaving (30 minutes). Either
4-fold or 10-fold dilution series were used with 3 replicate
MPN tubes. Dilution series were Prepared directly from
sediment slurries used in transformation assays. Autoclaved
control tubes were prepared using autoclaved sediment.
Following a 4-week incubation period, a 1-ml subsample
was removed from each MPN tube, mixed with a 1-ml
volume of acetonitrile and analyzed for 2,4-DCP. Tubes
were considered positive if the 2,4-DCP. peak area was not


more than three times the peak area of the 4-chlorophenol
(CP) product. Numbers of dechlorinating organisms (with a
95% confidence interval) were estimated from the number
of positive tubes in selected consecutive dilutions using an
MS-DOS turbo pascal program utilizing American Society of
Microbiology guidelines for MPN determinations.

Computer Simulations

Theoretical simulations of the fate of chlorinated aromatic
compounds in sediment slurries were prepared using a
Monod growth model (26). Two basic assumptions were
made. First, the growth rate of the dechlorinating
microorganisms can be expressed by

u = umax S/(K, + S) (1)

where u is the specific growth rate, S is the concentration of
chlorinated substrate, umax is the maximum specific growth
rate, and Kg is the half saturation constant (2). Second, the
following mass balance equation applies,

S. + BYY = S + BY (2)

where Sp, is the concentration of substrate at zero time, B, is
the concentration of bacteria at zero time, and Y is the
growth yield factor. Y is treated here as a constant, as has
been the practice of others (2-5). Because only changes in
S are of interest, the term X can be substituted for B/Y. If B
is given in cells/L and Y in cells/mg (number of cells formed
per mg parent compound dechlorinated) then X has the
units of mg (parent compound)/L. Thus, X corresponds to
the amount of "substrate" required to produce a population
density of B. By analogy, X, corresponds to the amount of
“substrate” that could be equated to the formation of the
initial population of specific degraders at tirne zero, i.e. Bo.
Because Y is considered a constant, u can be represented
as

u = 1X dX/dt (3)
Combining Equation 3 with Equation 1 yields
dX/dt = umax S X/(K, + S) (4)

Because of the mass balance Equation 2, dX/dt can be
assumed to equal -dS/dt, yielding

- dS/dt = umax S X/(K, + S) (5)

Plots of S versus t were obtained by simultaneously solving
Equations 4 and 5 using TUTSIM software.

Results and Discussion

The object of this project was to develop kinetic models for
predicting the anaerobic degradation of hazardous organic
chemicals in the saturated zone, and in particular, the
degradation of chlorinated aromatic compounds. Studies on
the anaerobic degradation of dichlorinated phenols were
conducted to develop the appropriate kinetic models. Where
possible, we refer to the work of others to illustrate how
Monod kinetics may also appropriately describe the
degradation of a wide variety of compounds in the saturated
zone.

To provide perspective and to ensure that our conclusions
have breadth of application, the degradation of the six
dichlorophenol isomers was investigated in the sediments
from five ponds near Athens, GA. In addition to the
dichlorophenol isomers, the degradation of PCP, 2,4-D and
2,4,5-T was investigated using sediments collected

throughout the United States (in Georgia, Florida, and New
York) and the Soviet Union. Although the data are not
presented here, similar results also were observed with
chlorinated anilines and benzenes. In all cases, the initial
dechlorination of the test compounds was preceded by a lag
period. Similar results have been observed for a variety of
chlorinated (3,6,9,10) and nonchlorinated aromatic
compounds (6,8,11,2,17,14,19). The length of the lag was
observed to be both sediment- (Tables 1 and 3) and
compound- (Tables 2 and 3) dependent. In those cases
where the test compound was added a second and third
time (Figures 1 and 2) following the complete dechlorination
of the previous addition, dechlorination was faster and no
lag was apparent. When such an increase in the onset and
rate of activity is observed, the sediments are considered to
have adapted to the dechlorination or degradation of the
particular compound under investigation (3,6,9,10,23).

In some adapted sediments or sludges, the compound loss
(3,6,9) and the formation of methane and carbon dioxide is
immediate (no lag) . Few, if any, degradation intermediates
are observed. At the other end of the spectrum are those
adapted sediments that remove only a single chlorine from
the compound to form a stable intermediate (9,10,23). In this
research, for example, 2,4-DCP was converted to 4-CP in
the sediment from one pond, and 2,3-, 2,4- and 2,6-DCP.
were converted to 3-, 4-, and 2- CP in the sediment from
another. None of the monochlorophenols was further
degraded. The conversion of 2,3- and 2,6-DCP to a mixture
of monochlorophenols and phenol in some sediments was
indicative of an intermediate level of adaptation. The direct
conversion of dichlorophenols to methane and carbon
monoxide was not indicated with either sediment.

That sediments can adapt to convert dichlorophenols to
monochlorophenols suggests that the dechlorination
process provides a selective advantage for the survival of
dechlorinating organisms. Brown et al. (4) have calculated
that the free energy released during reductive dechlorination
is exergonic. Therefore, one could expect that if this release
of energy could be coupled to the utilization of organic
compounds in sediment, the conversion of 2,4-DCP to 4-CP
observed here could support biological growth, resulting in
adaptation. Results indicate that the number of
dechlorinating microorganisms, as measured by MPN
techniques, increases following the addition of
dichlorophenols to sediment slurries (Table 4). Although an
increase in MPN units was not always observed, adaptation
was consistently observed. The relative importance of
induction (i.e., an increase in enzymatic activity) and growth
were not evaluated here. It is apparent from these studies,
however, that both are occurring. Considering that the sterile
controls in these studies showed no loss of compound,
abiotic processes can reasonably be ruled out.

Several other mechanisms have also been considered in
explaining the length of lag periods. Others have suggested
that a lag period is required for mutation and genetic
transfer (16,25). Also, certain environmental factors have
been implicated. These include limiting nutrient
concentrations (13,24), preferential use of organic (12,13) or
inorganic (7) compounds before degradation of the test
chemical, recovery from toxic chemicals (21), and the
predation of degrader populations by protozoa (26).

Because we observed both adaptation and an increase in
the units of biological activity, Monod growth kinetics can be