A case study for demonstrating the application of U.S. EPA ...clemept/publsihed_pdf/npcpaper.pdfA case study for demonstrating the application of U.S. EPA’s monitored natural attenuation

A case study for demonstrating the application of

U.S. EPA’s monitored natural attenuation

screening protocol at a hazardous

waste site

T. Prabhakar Clement a,*, Michael J. Truex b, Peter Lee c

aDepartment of Environmental Engineering, University of Western Australia, Nedlands, WA 6009, AustraliabBattelle Pacific Northwest Division, Richland, WA 99352, USA

cNPC Services Inc., 2401 Brooklawn Drive, Baton Rouge, LA 70807, USA

Received 11 June 2001; received in revised form 12 February 2002; accepted 13 February 2002

Abstract

Natural attenuation assessment data, collected at a Superfund site located in Louisiana, USA, are

presented. The study site is contaminated with large quantities of DNAPL waste products. Source

characterization data indicated that chlorinated ethene and ethane compounds are the major

contaminants of concern. This case study illustrates the steps involved in implementing the U.S.

EPA’s [U.S. EPA, 1998. Technical protocol for evaluating natural attenuation of chlorinated solvents

in ground water, by Wiedmeier, T.H., Swnason, M.A., Moutoux, D.E., Gordon, E.K., Wilson, J.T.,

Wilson, B.H., Kampbell, D.H., Hass, P.E., Miller, R.N., Hansen, J. E., Chapelle, F.H., Office of

Research and Development, EPA/600/R-98/128] monitored natural attenuation (MNA) screening

protocol at this chlorinated solvent site. In the first stage of the MNA assessment process, the field

data collected from four monitoring wells located in different parts of the plume were used to

complete a biodegradation scoring analysis recommended by the protocol. The analysis indicates

that the site has the potential for natural attenuation. In the second stage, a detailed conceptual model

was developed to identify various contaminant transport pathways and exposure points. The U.S.

EPA model and BIOCHLOR was used to assess whether the contaminants are attenuating at a

reasonable rate along these transport paths so that MNA can be considered as a feasible remedial

option for the site. The site data along with the modeling results indicate that the chlorinated ethene

and chlorinated ethane plumes are degrading and will attenuate within 1000 ft down gradient from

0169-7722/02/$ - see front matter D 2002 Elsevier Science B.V. All rights reserved.

PII: S0169 -7722 (02 )00079 -7

* Corresponding author. Department of Civil Engineering, Auburn University, AL 36830, USA.

E-mail address: [email protected] (T.P. Clement).

www.elsevier.com/locate/jconhyd

Journal of Contaminant Hydrology 59 (2002) 133–162

the source, well before reaching the identified exposure point. Therefore, MNA can be considered as

one of the feasible remediation options for the site.

D 2002 Elsevier Science B.V. All rights reserved.

Keywords: Modeling; Groundwater contamination; Bioremediation; Biodegradation; Natural attenuation;

Reactive transport

1. Introduction

Waste disposal operations began at the Petro-Processors Inc. (PPI) Brooklawn site

(known as Brooklawn site) between 1968 and 1970 and continued until 1980. During this

period, various types of hazardous waste material were disposed at the site. The material

included dense nonaqueous phase liquids (DNAPLs) that originated from chlorinated

solvent manufacturing plants and from other refineries. Based on a preliminary site

investigation, which was completed by NPC Services, a draft work plan for implementing

remedial activities was developed in 1984. A hydraulic containment system, and an active

source recovery system coupled with the treatment of the extracted water were selected as

remedial strategies.

Microcosm tests recently performed using the sediment samples collected at the

Brooklawn site indicated that the soil microbes have the potential to degrade various

chlorinated compounds (Acar et al., 1995; Constant et al., 1995; Clover et al., 1998; Pardue,

1999; Truex et al., 2001). Therefore, monitored natural attenuation (MNA) appears to be

one of the additional remedial alternatives available for managing the dissolved plumes at

the site. Previously published natural attenuation studies indicate that the biological activity

required for degrading many chlorinated organic compounds are ubiquitously present in

most anaerobic aquifers (Semprini et al., 1995; Bradley and Chappelle, 1997; Lorah and

Olsen, 1999). If sufficient natural or contaminant-derived organic carbon is available as a

substrate to support the growth of microbial populations, then MNA can be considered as

one of the feasible options for managing chlorinated solvent plumes (Wiedemeier et al.,

1999; Clement et al., 2000). A detailed protocol is now available for assessing the

attenuation processes for applying the technology at field sites (U.S. EPA, 1998, 1999;

Lu et al., 1999). According to the protocol, the first task in implementing MNA at a field

site involves completion of an initial screening assessment study. Subsequent tasks involve

modeling to determine exposures and better quantify fate. The objective of this case study is

to illustrate the efforts involved in implementing the MNA screening process, prescribed in

U.S. EPA (1998), at hazardous waste sites and the benefits of modeling. The natural

attenuation data set collected at the Brooklawn site is used for this purpose.

2. Site characterization data

The surficial features of the Brooklawn site and the location of various monitoring

wells used in the site characterization effort are shown in Fig. 1. The field site is located to

the north of Baton Rouge, LO, USA, approximately 5200 ft (about 1.5 km) away from the

T.P. Clement et al. / Journal of Contaminant Hydrology 59 (2002) 133–162134

Mississippi River. The land to the south of the site is largely undeveloped swampy

lowlands of the Mississippi River floodplain, known as the Devil’s Swamp, and in the

north there is a large industrial development site. A small stream known as the Bayou

Baton Rouge, which originates about 7 miles north of the site, runs along the western

boundary of the site and turns east at the southern site boundary and breaks into several

distributaries which discharge into the Swamp. The dissolved contaminant plume at the

site extends about 4000 ft in the east–west direction and about 1100 ft in the north–south

direction.

2.1. Geological data

The Brooklawn site is located on the interface between ancient Pleistocene sedimentary

deposits and the recent alluvial sediments deposited by the Mississippi River. The interface

is marked by a topographic bluff line that transverses in the east–west direction. The bluff

line approximately parallels the Mississippi River, and is about 30 ft higher than the

adjacent floodplain. This line is an erosional feature carved by the Mississippi River. To

the north of the bluff line, the upland Pleistocene sedimentary deposits predominantly

consist of clayey material, and to the south of the line the floodplain alluvium sediments

predominantly consist of sandy material. Fig. 2 shows a geologic cross section of the site

through the primary contaminant source area. As shown in the figure, the upland

Pleistocene deposits consist of clay from the surface (or from the erosional contact point

with the alluvium) to a depth of about �160 ft MSL. The conductivity values of the clay

Fig. 1. Site details and well locations.

T.P. Clement et al. / Journal of Contaminant Hydrology 59 (2002) 133–162 135

Fig.2.Representativegeological

cross

sectionacross

thesourcearea.


ranges from 10�6 to 10�4 ft/day. Some interlayers of silt and sand are present within the

Pleistocene clay unit. One of the interlayered silt zones, the ‘‘�40 ft silt’’ layer, seems to

be ubiquitously present across the entire site. The hydraulic conductivity of the silt zone is

about 0.1 to 1 ft/day. Additionally, an ‘‘intermediate sand unit’’, which is a 40-ft-thick

sandy silt layer, also runs across the site starting at a depth of about �80 ft MSL. The

‘‘400-ft’’ aquifer, which is a highly permeable sand aquifer (with conductivity values in

the range 10 to 100 ft/day), underlies the entire site starting at a depth of �160 ft MSL.

The alluvial sediments in the floodplain region are generally sandy-to-silty in texture

near the surface. With depth, the sediments become interlayered and intermixed with fine

sandy silts to medium sand separated by thin discontinuous clay layers which are generally

less than 5 ft thick. Clay mixed sands with appreciable amounts of organic material were

found in numerous soil borings within the floodplain region. The hydraulic conductivity of

the alluvium sand ranges from 1 to 10 ft/day. The thickness of the top alluvial unit

increases with distances away from the site. As shown in Fig. 2, along the clay–alluvium

interface, the alluvial unit pinches out and intersects with the Pleistocene clay unit; and

with distances away from the site, towards the river, the alluvial unit become thicker and

intersects the ‘‘�40 ft silt’’ layer and later intersects the ‘‘intermediate sand’’ layer.

2.2. Background geochemical data

Water samples collected from wells P-0936-1 and P-1620-1, which are located outside

the plume (see Fig. 1), were selected to assess the background geochemistry of Brooklawn

groundwater. Note that one of the selected wells (P-0936-1) is located upstream in the

Pleistocene clay unit and the other well (P-1620-1) is located downstream in the alluvium

unit. The measured geochemical constituents of the Pleistocene and alluvium units are

summarized in Table 1. The geochemical data indicate that the alluvium, which is

Table 1

Background geochemical characteristics of the groundwater

Constituent Alluviuma Pleistocene clay unitb Units

Chloride 29 34 mg/l

Specific conductance 616 604 Amho

Dissolved methane not detected not detected mg/l

Dissolved oxygen not detected 2 mg/l

Dissolved hydrogen 5.2 2 nM

Inorganic carbon 81 26.7 mg/l

Iron, ferrous 7.9 0.1 mg/l

Nitrite not detected 0.04 mg/l

Nitrate 0.3 0.8 mg/l

Oxidation/reduction potential �40 �46 mV

pH 6.6 7.3 standard units

Sulfide not detected 0.1 mg/l

Sulfate not detected 8 mg/l

Temperature 20 18 jCTotal organic carbon 11 1.5 mg/l

a Average concentration from wells P-1620-1 and P-2953-1.b Concentration from well P-0936-1.


expected to be the primary transport pathway, can be characterized as an anaerobic system

with neutral pH and moderate ionic strength and hardness. The redox characteristics of the

Pleistocene clay zone also indicate reducing conditions. Although low levels of dissolved

oxygen were measured in P-0932-1, which is located outside the site boundary, several

other wells located within the plume in the Pleistocene clay were devoid of oxygen, thus

indicating anaerobic conditions. Measured organic carbon content of the alluvium sedi-

ments ranged from 0.39% to 1.13% (Valsaraj et al., 1999).

2.3. Source locations

Based on historic site characterization data, the spatial extents of various DNAPL

contamination zones were delineated. The boundaries of the source zones are shown in

Figs. 1 and 2. The bulk of the DNAPL mass is present in the eastern portion of the site.

However, some minor isolated pits and disposal drains were also present in the western

region. Historical data show that most of the waste products were initially disposed into

several earthen pits located in the Pleistocene clay unit along the upland portion of the site.

During remedial investigations, these upland waste NAPLs were found to be confined

with only minimal NAPL mass migration beyond the pits. However, large amounts of

NAPL were later disposed into excavated pits in the lower floodplain area, which are

surrounded by constructed levees. The NAPL products disposed in this area have migrated

beyond the limits of the pits. Further, during a flood event in the 1970s, a portion of the

levee failed and NAPL products were directly discharged into the small drainage channels

along the southwest corner of the site. This event created a ‘‘spill area’’ (see Fig. 1), which

was delineated based on soil core data collected from the channels. Further site

investigations revealed that the presence of NAPL was mostly limited to shallow regions,

with isolated deeper occurrences in the northern area of the channel adjacent to the

Brooklawn site.

2.4. Contaminant characterization data

The contamination currently present at the site consists of large quantities of pooled or

trapped NAPL-phase products and various dissolved plumes that emanate from the free

phase. The DNAPL present at the site is a complex mixture of several organic compounds.

Table 2 shows the groundwater concentrations of various contaminant species measured in

two monitoring wells located within the dissolved plume region. The first well, P-1426-6,

is located down gradient from the primary DNAPL source area, and the second well,

PBB21-1N, is located near the edge of a secondary spilled source region. The data show

that dissolved chlorinated ethene and chlorinated ethane compounds are at concentration

levels of one to two orders of magnitude greater than all other compounds. High values of

chloroform and 1,2-dichloropropane were also measured in some of the groundwater

samples. Chloroform can be biodegraded and is itself an anaerobic biotransformation

product of carbon tetrachloride (Beelen and Keulen, 1990; Picardal et al., 1995).

Researchers have shown that chloroform can be degraded by anaerobic methanogenic

enrichments and by other non-methanogenic anaerobic cultures (Bagley and Gossett,

1995). Literature information also suggests that 1,2-dichloropropane can be completely


dechlorinated under anaerobic conditions (Loffler et al., 1997). Since the dissolved levels

of chlorinated ethene and ethane compounds were good indicators of the status of overall

contamination, further analyses of chloroform and dichloropropane were not included in

this screening study. Non-chlorinated contaminants such as the BTEX (benzene, toluene,

ethylbenzene and xylene) compounds were also not included because the measured BTEX

levels in down gradient wells were much lower than chlorinated compounds. In this initial

MNA screening study, we primarily focus on the fate and transport of the following

chlorinated ethane and ethene components: 1,1,2,2-tetrachloroethane (TeCA), 1,1,2-

trichloroethane (TCA), 1,2-dichloroethane (DCA), chloroethane (CA), tetrachloroethene

(PCE), trichloroethene (TCE), dichloroethene (DCE), and vinyl chloride (VC). These

components were previously identified as the contaminants of concern for the site based

on their chemical, physical, and other transport properties (NPC, 1996). Since chlorinated

solvents are prevalent at most DNAPL waste sites, the U.S. EPA (1998) MNA guidelines

also primarily focus on analyzing these contaminants.

3. Biodegradation assessment

The MNA assessment process involves a six-step screening method (U.S. EPA, 1998).

As shown in Fig. 3, the first step in the MNA assessment is to use the site data to answer

the important question ‘‘Is biodegradation occurring at the site?’’ In order to address this

Table 2

Concentration of chlorinated constituents in the alluvium

Compound Concentration at well P-1426-6 (Ag/l) Concentration at well PBB21-1N (Ag/l)

1,2,4-Trichlorobenzenea <10 <20

1,2-Dichlorobenzenea,b <5 <5

1,2-Dichloropropanea 4640 24000

Bis(2-Chloroisopropyl)ethera 40.7 20.9

Carbon Tetrachloridea <5 <5

Chlorobenzenea,c <5 53

Chloroforma 70.1 4040

Hexachlorobenzenea <10 <20

Hexachlorobutadienea <10 <20

Hexachloroethanea <10 <20

1,1,2,2-Tetrachloroethane 153 6960

1,1,2-Trichloroethane 9710 98100

1,1-Dichloroethene 683 2670

1,2-Dichloroethane 15700 99600

Tetrachloroethene 117 3420

Trichloroethene 3710 13200

cis-1,2-Dichloroethene 2790 13100

trans-1,2-Dichloroethene 763 5

Vinyl chloride 25700 58200

a Compound is not included for further study in this initial screening for MNA.b The concentration of all isomers of dichlorobenzene were similar.c Chlorobenzene is included in the table because it is a biotransformation product of other chlorinated

benzenes.


question, the average concentrations of various geochemical parameters measured in

alluvium wells P-1620-1 and P-2953-1, shown in Table 1, were first used to set the

background levels. Groundwater geochemical data collected from four wells including P-

1223-2, P-1426-6, PBB21-1N, and P-1535-2, which are expected to represent different

types of biochemical and/or hydraulic regimes present at the site, were used to assess the

biodegradation conditions. It can be seen from Fig. 1 that the first well, P-1223-2, is

located to the south of the primary NAPL source area. This well is adjacent to the highly

contaminated source area and was (prior to pumping, between 1970 and 1995) down

gradient from the source. Currently, the well is within the hydraulic capture zone of the

Fig. 3. Monitored natural attenuation screening process flow sheet [adapted from U.S. EPA, 1998].


extraction wells and contains significant concentrations of chlorinated solvents. The

second monitoring well, P-1426-6, is located about 300 ft south of the primary NAPL

source area. This well location was also previously (prior to pumping) down gradient from

the source, but is now outside the hydraulic capture zone of the pumping wells. The

hydraulic gradient at this location changes seasonally due to fluctuations in the Mississippi

River stage. The third monitoring well, PBB21-1N, is located at the edge of a NAPL

source area in a bayou channel in the ‘‘spill zone’’. This well is close to the NAPL source

region. The hydraulic gradient in this region has significant seasonal variations and is

influenced by the water levels in the bayou channel. The fourth well, P-1535-2, is located

in the swampy area down gradient from the previous monitoring location. The hydraulic

gradient at this location also has considerable fluctuations.

Using the U.S. EPA (1998) MNA framework, a specified number of ‘‘points’’ were

assigned depending on the concentration of the geochemical indicators observed in the

wells. Points were awarded only if the concentration of a geochemical indicator was within

the range specified in the screening criteria and if the indicator was not a constituent of the

original contaminant source. The points were added and interpreted based on U.S. EPA

guidelines to determine whether biodegradation is occurring at the selected location. If the

total score was above 15 points, the location was deemed to have a good potential for

natural attenuation. Further details of this scoring scheme are discussed in U.S. EPA

(1998).

The results of the biodegradation assessment for the four selected Brooklawn site wells

are summarized in Table 3a and b. As shown in the table, the biodegradation assessment

scores for all four locations are found to be greater than 15, indicating a good potential for

natural biodegradation. Most of the points assigned to each location were due to the

detection of geochemical indicators of the presence of anaerobic environments—such

environments can support biological activities required for mediating dechlorination

processes. A few points were also assigned for the presence of certain degradation

products. For example, in wells P-1426-6 and PBB21-1N, cis-DCE, a daughter product of

chlorinated ethene and chlorinated ethane biotransformation, was present at concentration

levels much greater than the concentrations of the other dichloroethene isomers. The

concentration of cis-DCE was below detectable levels at the other two monitoring

locations.

Points could not be assigned for the presence of several other biodegradation products

(other than cis-DCE) even though they were present at very high levels. This is because

several of these possible biodegradation products were already present in the NAPL

source, although at very low levels. For example, VC, which is a by-product of the

reductive dechlorination process, was measured as 25.7 mg/l in well P1426-6 and 58.2

mg/l in PBB21-IN. However, the measured VC values were consistently low near the

source region. Further, the measured weight percentage of VC in the NAPL source was

about 0.026%, which yields a maximum effective VC solubility of 0.9 mg/l. Therefore, the

high VC concentration levels observed in some of the down gradient wells could have

been due to biodegradation. However, because VC is contained within the original NAPL

source, points cannot be awarded for the presence of VC according to the U.S. EPA (1998)

guidelines. This demonstrates the conservative nature of the scoring strategy, which could

result in underestimation of the overall biodegradation potential.


Table 3

Bioattenuation screening parameters and scoring

Contaminant/geochemical

indicator

Criteria In NAPL

source?

PBL-1223-2

(shallow)

concentration

PBL-1223-2

(shallow) score

P-1426-6 (deep)

concentration

P-1426-6

(deep) score

Dissolved oxygen (mg/l) <0.5 N/A 0 3 0.92 0

Nitrate (mg/l) <1 N/A 0.4 2 1 2

Iron(II) (mg/l) >1 N/A 44.2 3 3.09 3

Sulfate (mg/l) <20 N/A <5 2 <1.0 2

Sulfide (mg/l) >1 N/A 0.006 0 0.003 0

Methane >0.5 N/A no data 0 4 3

Oxidation/reduction

potential (mV)

<50 or <�100 N/A �58 1 no data 0

pH 5<pH>9 N/A 6.4 N/A 6.8 N/A

Total organic carbon (mg/l) >20 N/A no data 0 21.2 2

Temperature >20 N/A 22.3 1 no data 0

Carbon dioxide >2� N/A no data 0 no data 0

Alkalinity >2� N/A 9.9 0 38.8 0

Chloride (mg/l) >2� N/A 87 2 355 2

Hydrogen >1 nM N/A <0.08 0 4 3

Volatile fatty acids >0.1 N/A no data 0 no data 0

BTEX (mg/l) >0.1 Yes 0.02 0 0.065 0

TCE (Ag/l) Yes 5.32 0 3710 0

1,1,2-TCA (Ag/l) Yes <5 0 9710 0

1,2-DCA (Ag/l) Yes <5 0 15700 0

trans-1,2-DCE (Ag/l) Yes <5 0 763 0

cis-1,2-DCE (Ag/l) No no data 0 2790 2

1,1-DCE (Ag/l) No 21.8 2 683 2

VC (Ag/l) Yes 1010 0 25700 0

Chloroethane (Ag/l) No <5 0 <5 0

Ethene (mg/l) >0.01 or >0.1 No no data 0 16 3

Ethane (mg/l) >0.01 or >0.1 No no data 0 0.001 2

1,1,2,2-Tetrachloroethane (Ag/l) Yes <5 0 153 0

Tetrachloroethene (Ag/l) Yes <5 0 117 0

Total 16 26

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al./JournalofContaminantHydrology59(2002)133–162

142

Contaminant/geochemical

indicator

Criteria In NAPL

source?

PBB21-1N

concentration

PBB21-1N

score

P-1535-2

concentration

P-1535-2

score

Dissolved oxygen (mg/l) <0.5 N/A 0.02 3 0 3

Nitrate (mg/l) <1 N/A 0.8 2 0.4 2

Iron(II) (mg/l) >1 N/A 88.5 3 39 3

Sulfate (mg/l) <20 N/A 21 0 0 2

Sulfide (mg/l) >1 N/A 0.001 0 0.001 0

Methane >0.5 N/A 2.35 3 0 0

Oxidation/reduction potential (mV) <50 or <�100 N/A no data 0 �53 1

pH 5<pH>9 N/A 6.15 N/A 7.6 N/A

Total organic carbon (mg/l) >20 N/A 111 2 30.8 2

Temperature >20 N/A no data 0 24 1

Carbon dioxide >2� N/A no data 0 no data 0

Alkalinity >2� N/A 69.5 0 146 1

Chloride (mg/l) >2� N/A 440 2 110 2

Hydrogen >1 nM N/A 3.7 3 >8 3

Volatile fatty acids >0.1 N/A no data 0 no data 0

BTEX (mg/l) >0.1 Yes 0.2 2 0.035 0

TCE (Ag/l) Yes 13200 0 413 0

1,1,2-TCA (Ag/l) Yes 98100 0 3400 0

1,2-DCA (Ag/l) Yes 99600 0 2770 0

trans-1,2-DCE (Ag/l) Yes <5 0 84.8 0

cis-1,2-DCE (Ag/l) No 13100 2 no data 0

1,1-DCE (Ag/l) No 2670 2 170 2

VC (Ag/l) Yes 58200 0 4080 0

Chloroethane (Ag/l) No <5 0 <5 0

Ethene (mg/l) >0.01 or >0.1 No 16 3 0 0

Ethane (mg/l) >0.01 or >0.1 No 0.385 2 0 0

1,1,2,2-Tetrachloroethane (Ag/l) Yes 6960 0 25.4 0

Tetrachloroethene (Ag/l) Yes 3420 0 <5 0

Total 29 22

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143

4. Development of conceptual models

As shown in Fig. 3, in the second stage of the MNA screening process, a conceptual

model must be developed to provide the context for understanding the transport processes

occurring at the site. Fig. 4 shows the details of the source area, major contaminant

transport pathways, and various other surficial features. In our conceptual model develop-

ment effort, contaminant transport from the major source area where the bulk of DNAPL

mass currently resides is only considered. Fig. 4 also shows the location of various

monitoring wells and boreholes; data from these wells were used to build the computer

simulation model. All model simulations completed in this study focused on predicting the

fate-and-transport of contaminants under natural gradient conditions. In order to be

consistent with this scenario, the data collected prior to the start of groundwater extraction

activities were only used.

4.1. Conceptual model for contaminant transport pathways and exposure points

Fig. 5 is a conceptual cross section model of the site, which shows the relative locations

of potential exposure points along various transport pathways. The Mississippi River and

the ‘‘400-ft’’ aquifer were identified as the exposure points of concern. As shown in the

figure, the conceptual model considers two distinct transport pathways: (1) a horizontal

Fig. 4. Conceptual model for contaminant transport and other site details.


Fig.5.Descriptionofsourcelocationsandtransportpaths.


path towards the Mississippi River and (2) a vertical path towards the deep ‘‘400-ft’’

aquifer. The total distance along the horizontal transport path, via the alluvium, from the

leading edge of the source to the Mississippi River is about 5200 ft. The total distance

along the vertical transport path, via the Pleistocene clay, from the bottom of the source to

the 400-ft aquifer, is approximately 140 ft. In addition to these two major transport

pathways, it is possible that the plumes might reach the 400-ft aquifer through an inclined

pathway via the alluvium where the clay layer might be discontinuous. Since little

characterization information is available beyond the site boundary, it is difficult to

determine the transport characteristics of this inclined pathway, which would include

transport via the alluvial zone and some portions of discontinuous clay layers. Preliminary

analyses suggest that such a lengthy transport pathway may not be critical because alluvial

sediments are naturally less permeable in the vertical direction as compared to the

horizontal direction and, hence, are expected to yield more resistance to vertical transport.

Also, the observed vertical hydraulic gradient between the alluvium and 400-ft aquifer is

very small. Therefore, within the context of this initial screening effort, transport along this

inclined pathway was not considered.

The model shown in Fig. 5 was conceptualized based on the borehole data shown in

Fig. 2 and the time series of groundwater levels shown in Fig. 6. As shown in Fig. 6, the

transient groundwater levels in the alluvium and in the ‘‘400-ft’’ aquifer were similar, and

the trend also closely followed the seasonal variations in the Mississippi River. Water

levels in the Pleistocene clay unit do not respond to the river variations. These field

observations are consistent with our conceptual model.

The water levels observed in the alluvium (P-1620-2) and the 40-ft-silt zone (S-UG-1)

wells showed an average hydraulic gradient towards the river. The interrelationships

between the water level response patterns observed in the alluvium, silt layer, and

Fig. 6. Seasonal variations in groundwater levels observed in different hydraulic units.


Pleistocene clay unit were used to develop a conceptual framework for the transfer of

contaminant mass from the source zone to the alluvium unit. Note that the site character-

ization data, summarized in Fig. 2, reveals that a portion of the DNAPL source disposed in

the alluvium pits was in close contact with the silt layer. Therefore, in our conceptual

contaminant mass transfer model, it is assumed that the contaminants from the DNAPL

source seep into the 40-ft silt layer and later are transported by the horizontal groundwater

level gradient between the silt and alluvium zones.

4.2. Estimation of contaminant transport velocity

Groundwater levels observed in the silt (S-UG-1) and alluvium (P-1620-2) wells were

used to estimate a yearly averaged horizontal transport velocity. Table 4 provides the water

levels measured in these wells over a 12-month period. As shown in the table, the yearly

average of head differences between these two wells is 2.15 ft and the distance between the

wells is 800 ft; this yields an average groundwater gradient of 0.003. Using an estimated

porosity value of 0.3 and conductivity value of 5 ft/day for the alluvium, the average

transport velocity for the horizontal pathway, towards the river via the alluvium aquifer, is

0.05 ft/day.

Similarly, the vertical hydraulic gradient was estimated based on the average head

difference between the �40-ft silt well (S-UG-1) and the ‘‘400-ft’’ aquifer (well D-UG-1).

While the vertical transport path has a much shorter length than the horizontal transport

path, the material between the contamination and the ‘‘400-ft’’ aquifer is low permeable

clay. Based on measured hydraulic conductivity value of 0.0001 ft/day for the clay

material, an estimated vertical hydraulic gradient of 0.019 ft/ft and an estimated porosity of

0.4, the vertical transport velocity is approximated as 4.75�10�6 ft/day. Contaminants

Table 4

Calculation of average horizontal hydraulic gradient for the alluvium

Month Hydraulic head (ft),

Well S-UG-1

(Pleistocene interface)

Hydraulic head (ft),

Well P-1620-2

(Alluvium)

Hydraulic head

difference (ft)

April 31.48 35.15 �3.67

May 32.56 33.93a �1.37

June 31.54 30.71 0.83

July 31.06 31.25 �0.19

August 30.14 24.53 5.61

September 28.02 18.4 9.62

October 26.56 18.43 8.13

November 26.27 19.61 6.66

December 26.04 22.55 3.49

January 26.47 23.87 2.6

February 27.24 30.79 �3.55

March 30.41 32.86 �2.45

Average head difference (ft) 2.15

Approximate gradient (based on 800 ft distance between wells) (ft/ft) 0.003 (toward river)

a Estimated value since no data is available for the month.


traveling at this velocity would require over 80,000 years to traverse the 140 ft of clay

zone present between the current location of the contaminant source and the top of the

‘‘400-ft’’ aquifer. Moreover, monitoring data from the 400-ft aquifer also indicated that the

aquifer is uncontaminated. Therefore, further analysis of the vertical pathway was not

considered in this initial screening assessment.

5. Biodegradation and reactive transport

The U.S. EPA screening model, BIOCHLOR, was used to model the attenuation

processes occurring at site. BIOCHLOR is an analytical computer code that is intended for

use as a screening-level model to determine if remediation by natural attenuation is

feasible at a chlorinated solvent site (Aziz et al., 2000). The code uses a novel analytical

solution strategy to solve the multi-species sequential reactive transport problem (Sun et

al., 1999; Sun and Clement, 1999; Clement, 2001). BIOCHLOR has the ability to simulate

uniform flow with three-dimensional dispersion, linear adsorption, and biodegradation via

reductive dechlorination reactions. The model can predict migration patterns of a parent

chlorinated solvent species (either TCA or PCE) and its daughter products.

BIOCHLOR assumes first-order kinetics to model the biological decay reactions. The

use of first-order kinetics is appropriate when the biodegradation rate is primarily a

function of the concentration of the contaminant, when the number of microorganisms that

can degrade the contaminant is constant over time within the region of interest, and when

all other nutrients critical to the biodegradation processes are in abundance. For most field-

scale natural attenuation modeling applications, the first-order assumption may be

considered as a reasonable approximation (U.S. EPA, 1998; Aziz et al., 2000) provided

the electron donor is not limiting. Moreover, the assumption of first-order kinetics is often

acceptable for biodegradation at low pollutant concentration levels (Schmidt et al., 1985),

which is typically encountered in most groundwater remediation problems.

The disposal operations at the site started between 1968 and 1970, and the field data

used for this paper were collected nominally between 1992 and January 1995. Therefore,

the simulations were completed for 25 years so that direct comparison to the field data is

possible. The physical dimensions of the contaminant source were determined based on

the scale of the primary source shown in Figs. 4 and 5. The BIOCHLOR model requires

several basic transport parameters as input values, which include advection velocity,

dispersion coefficients, and retardation factors for the contaminant species. Further, the

model requires first-order degradation rate coefficients for the selected chlorinated solvent

reductive dechlorination sequence (either chlorinated ethenes or chlorinated ethanes). In

the following sections, the methods used for selecting appropriate retardation values,

biodegradation rate constants, source zone concentration levels, and other flow and

transport parameters are summarized.

5.1. Retardation parameters

The effective transport velocity of the contaminants is greatly influenced by the

adsorption characteristics of the porous medium. A linear equilibrium-partitioning model


was assumed to quantify the adsorption characteristics of Brooklawn sediments. Estimates

for the partition coefficient Kd (and subsequently the retardation factor R) were obtained

using an empirical correlation function. Valsaraj et al. (1999) performed sorption experi-

ments using three types of Brooklawn sediments and developed the following function:

Kd ¼ 100:81K0:56ow foc

where Kow is the octanol–water coefficient and foc is the fraction of organic carbon of the

soil. Table 5 presents the value of Kow and the calculated values for Kd and R values for the

chemicals considered in this initial MNA screening study. An foc value of 0.39%, which

was the value measured by Valsaraj et al. (1999) for the sandy alluvial material, was used

to compute the retardation parameters summarized in Table 5.

5.2. Biodegradation rate parameters

The Brooklawn site contamination includes a mixture of both chlorinated ethene and

chlorinated ethane compounds. The BIOCHLOR model is capable of simulating the

degradation of either the chlorinated ethene reaction chain (starting with PCE) or the

chlorinated ethane reaction chain (starting with TCA), not the mixture. However, when

both ethene and ethane species are present at the site, then it is difficult to interpret ethene

daughter products because some of the chlorinated ethenes can be produced from

chlorinated ethane decay reactions. Fig. 7 shows possible reaction pathways for degrada-

tion of PCE and TeCA and the subsequent chlorinated ethane and chlorinated ethene

daughter products (Lorah and Olsen, 1999). In this work, the BIOCHLOR simulations

were first completed for each individual ethene and ethane reaction chain. Subsequently,

approximations were made to quantify the influence of chlorinated ethenes produced from

chlorinated ethane decay reactions.

First-order kinetic models are assumed to be sufficient to describe all the reaction steps

represented in Fig. 7, which encompasses both chlorinated ethane and chlorinated ethene

Table 5

Estimated value of retardation coefficients

Compound Log10 of octanol–water

coefficient log Kowa

Partition

coefficient Kd (l/kg)b

Retardation coefficient

(1+qKd/n)c

TeCA 2.39 0.55 3.9

TCA 2.12 0.62 4.3

DCA 1.47 0.17 1.9

CA – – 1 (estimate)

PCE 2.88 1.0 6.5

TCE 2.42 0.57 4.0

DCEd 1.48 0.25 2.4

VC 0.6 0.05 1.3

a Values from Schwarzenbach et al. (1993).b Estimated from Valsaraj et al. (1999) correlation.c Assuming q=1.6 kg/l and n=0.3.d Value listed is for 1,1-dichloroethene.


decay reactions. In addition, since many of the biotic reactions need organic carbon as

substrate to support the direct dechlorination or co-metabolic reactions, it is assumed that

sufficient organic carbon is available at the site. This assumption is reasonable at

Brooklawn site because the site is located down gradient from a swamp that has large

amounts of decaying material that can supply large amounts of carbon. In addition, carbon

sources could also be derived from other waste organic material co-disposed with

chlorinated compounds.

To assess the biodegradation kinetics of a mixture of chlorinated ethane and chlorinated

ethene compounds, the degradation rate coefficients must be considered in conjunction

with the fraction of each potential product produced from different parent compounds. As

shown in Fig. 7, chlorinated ethane species can degrade and produce less-chlorinated

ethane species or chlorinated ethene species. Table 6 presents the estimated fraction of

each daughter product that could be expected to be formed from different parent

compound. These estimates are based on laboratory data (Lorah and Olsen, 1999; Chen

et al., 1996).

Table 7 summarizes the assumed values of first-order decay rate coefficients for

modeling anaerobic destruction of chlorinated ethane and chlorinated ethene species.

These values were used in the BIOCHLOR model to perform the initial fate-and-transport

analysis for the Brooklawn site. Note that rate coefficients are only presented for reductive

dechlorination along one class of daughter products (e.g., only ethanes or only ethenes)

because these are the only sequences that the BIOCHLOR model can describe. Also, to

match the required inputs for the BIOCHLOR model, TCA is assumed to be the parent

Fig. 7. Anaerobic transformation pathways for the chlorinated ethane and ethene compounds.


compound of the chlorinated ethane series. Further, all DCE isomers are combined

together and assumed to be the daughter product of biodegradation.

5.3. Source concentrations

Dissolved concentration levels of chlorinated ethane and chlorinated ethene species

measured or estimated to be present near the source zone are listed in Table 8. The first

column in the table lists field measured concentration data. These data are averages of

dissolved-phase concentrations in the groundwater samples collected from wells W-0822-

3, W-1025-2, and W-0626-1, where the water is in direct contact with NAPL products (see

Fig. 2 for well locations). The concentrations in the second column were estimated based

on a solubility analysis using measured NAPL composition data (Truex et al., 1999). It is

interesting to note that in Table 8 the measured concentration values and the calculated

effective solubility values compare favorably whenever the field data are above the

detection limit.

As shown in Fig. 7, chlorinated ethenes can be produced from biodegradation of the

chlorinated ethanes present in the NAPL source. The source area concentration of the

‘‘parent’’ chlorinated ethanes (TeCA and TCA), which can potentially produce chlorinate

ethenes, are approximately 60 and 337 mg/l, respectively. Clearly, these high concen-

trations of ethane parent compounds can yield significant amounts of chlorinated ethenes

Table 6

Estimated fractional percent conversion for chlorinated ethanes and chlorinated ethenes

Reaction pathway Estimated yield

(mol/mol)

Based on data from:

PCE!TCE 1.0 Lorah and Olsen (1999)

TCE!DCE 1.0 Lorah and Olsen (1999)

DCE!VC 1.0 Lorah and Olsen (1999)aTeCA!TCA 0.35 Lorah and Olsen (1999)

TeCA!TCE 0.02 Lorah and Olsen (1999)

TeCA!DCE 0.63 Lorah and Olsen (1999)aTCA!DCA 0.2 Chen et al. (1996)

TCA!VC 0.8 Chen et al. (1996)aDCA!CA 0.7 Chen et al. (1996)

DCA!ethane 0.3 Assumed

a As expected, sum of all TeCA or TCA or DCA yields is equal to unity.

Table 7

First-order rate coefficients

Reaction Value (1/day) Basis

TCA!DCA 0.013 One tenth of the rate estimated for Lorah and Olsen (1999) microcosm data

DCA!CA 0.001 One tenth of the rate estimated for Lorah and Olsen (1999) microcosm data

CA!ethane 0.014 One tenth of the rate estimated for Lorah and Olsen (1999) microcosm data

PCE!TCE 0.005 Set to be equal to TCE conversion rate

TCE!DCE 0.005 One tenth of the rate measured in microcosms using Brooklawn sediments

DCE!VC 0.005 Set to be equal to TCE conversion rate

VC!ethene 0.0006 One tenth of the rate measured in microcosms using Brooklawn sediments


down gradient from the source. The conversion fractional percent data presented in Table 6

can be used to estimate the maximum possible concentration of each chlorinated ethenes

produced from the ethane compounds present in the source. These estimates are

summarized in the third column of Table 8. Note if the biodegradation rates of ethane

reactions are high (when compared to ethene reactions), the mass of ethene produced from

ethane may simply be added to the original ethene source levels. This approximation is

employed in this study to indirectly simulate the combined ethane–ethene reactions.

5.4. Transport parameters

The transport parameters used in the model simulations are summarized in Table 9. The

transport properties of the alluvium were estimated based on field measurements. Methods

used for estimating the hydraulic gradient values were discussed in Section 4. The values

of retardation factors for different chlorinated compounds were estimated based on site-

specific sorption data shown in Table 5. The value for porosity was estimated from

literature values for similar types of geologic materials. The longitudinal dispersivity value

was estimated based on the guidelines presented in Gelhar et al. (1992). The ratio of the

longitudinal to transverse dispersivity was assumed to be 0.1, and ratio of the longitudinal

to vertical dispersivity was assumed to be 0.01.

Table 8

Contaminant concentrations (in mg/l) near the source zone

Constituent Measured values Estimated valuesa As decay productsb

TeCA 57 52 N/A

TCA 337 210 N/A

DCA 600 184 N/A

CA <25 0 N/A

PCE <25 3.5 N/A

TCE <25 6.2 1

DCE <25 0.8 21

VC <25 1.0 133

a Estimated based on solubility analysis.b Calculated using the chlorinated ethane concentrations from the first column and yield data from Table 6.

Table 9

Parameters used in BIOCHLOR simulations

Property Value

Hydraulic conductivity (ft/day) 5

Hydraulic gradient 0.003

Longitudinal dispersivity (ft) 50

Porosity 0.3

Average retardation for ethenes 3.6

Average retardation for ethanes 2.4

Model area (zone-1) length (ft) 5000

Model area width (ft) 2000

Source thickness (ft) 45

Source width (ft) 1000


6. BIOCHLOR simulations

Three sets of BIOCHLOR simulations were completed to assess whether adequate

bioattenuation is occurring at the Brooklawn site, and to evaluate the potential that a

steady-state condition may be established such that the contaminant plumes will not reach

the receptors. Wherever possible, the simulation results were compared against observed

data collected from the test boreholes drilled during a site investigation effort completed in

between September 1994 and January 1995. These test boreholes, BL-1225-1, BL-1426-1,

BB-1429-1, and BB-1626-1 (see Fig. 4), are located approximately 150, 300, 400, and 500

ft, respectively, from the source.

Simulation 1 predicted chlorinated ethane transport for a transport period of 25 years so

that the model results can be compared to the measured contaminant levels. The source-

zone concentrations used in this simulation were set based on field-measured values (see

Table 8, column 1). Both Simulations 2 and 3 predicted chlorinated ethene transport using

two types of source conditions. Since measured (above detection limit) source-zone

concentrations are unavailable for chlorinated ethenes, the first set of ethene simulation

(Simulation 2) was performed using source concentration levels computed based on the

concentration estimate from a solubility analysis (see Table 8, column 2).

In the second set of ethene simulations (Simulation 3), it was assumed that all

chlorinated ethane degradation reactions would occur at a rapid rate that would allow

complete ethane degradation close to the source zone. To reflect this condition, the source-

zone concentrations of chlorinated ethenes in Simulation 3 were set equal to the sum of

concentrations used in Simulation 2 and the maximum amount of ethene concentrations

that could be produced as products of ethane biodegradation (i.e., add ethene concentration

levels in column 2 and column 3 of Table 8). Analysis of field data along with model

simulations showed that this was a reasonable approximation for the Brooklawn site; this

point is discussed in more detail in the following section. The ultimate objective of

Simulation 3 was to assess the impact of chlorinated ethane degradation on predicted

chlorinated ethene levels. Since BIOCHLOR cannot be used to simulate coupled ethane–

ethene degradation reactions, this simplified approximation was made to assess the

coupled degradation condition.

6.1. Simulation 1—chlorinated ethane reaction chain

Fig. 8 presents the model-predicted TCA, DCA, and CA concentration profiles after 25

years of migration. The figure shows the predicted contaminant profiles down gradient

from the source zone along the plume centerline under three transport conditions: (1) no

attenuation, (2) attenuation by adsorption, and (3) attenuation by adsorption and biode-

gradation. The results show that the attenuation mechanisms including sorption and

degradation considerably retard plume migration. Also, all the biodegradation reactions

including production and destruction of intermediate compounds, such as CA, occur

within 500 ft down gradient of the source.

Field data for TCA and DCA indicate that the concentration of each species is less than

1 mg/l at distances of 150 and 300 ft, respectively, down gradient from the source. With

only adsorption-related attenuation, the model predicts the TCA and DCA concentrations


Fig. 8. Comparison of field data against BIOCHLOR simulations TCA and its daughter products (t=25 years).


as 50 and 450 mg/l, respectively, at these locations. This comparison indicates that the

TCA and DCA field data more closely match the concentration profiles predicted by

BIOCHLOR under the condition where both biodegradation and adsorption are occurring.

Since field data are not available for comparison to predicted CA concentration profiles,

these profiles must only be considered as preliminary estimates.

The BIOCHLOR model assumes TCA as the parent ethane contaminant and predicts

sequential dechlorination of TCA to DCA and later to CA. Thus, TCA produced from its

parent ethane compound TeCA, and also the possibility of partial conversion of TCA and

DCA compounds to VC (see Fig. 7), which has been measured in microcosm tests (Chen

et al., 1996; Lorah and Olsen, 1999), were not considered in this simulation.

6.2. Simulation 2—chlorinated ethene reaction chain

Fig. 9 presents model-predicted PCE, TCE, DCE, and VC concentration profiles after

25 years of migration. Similar to the previous analysis, these simulations were also

completed under different conditions including no attenuation, attenuation by adsorption,

and attenuation by adsorption and biodegradation. Source zone concentration levels used

in this simulation are based on the estimates from a solubility analysis (Table 8, column 2).

Under the assumed conditions, the model predicted that VC would be produced and

degraded within 300 ft down gradient from the source when biodegradation was assumed

to occur in the aquifer. With only adsorption-related attenuation, the VC concentration was

predicted to continuously reduce toward zero within 300 ft down gradient from the source.

The concentration of PCE measured in the field was less than 0.1 mg/l at a distance of

150 ft down gradient from the source. With only adsorption-related attenuation, Simu-

lation 2 predicted 2.5 mg/l of PCE at this location. The predicted PCE profile seems to

match the field data only when biodegradation and adsorption were assumed to simulta-

neously occur in the aquifer. The field data for TCE, DCE, and VC were more difficult to

interpret from Simulation 2 because each of the ethene species can also be produced from

degradation of chlorinated ethane species. Fig. 9c and d clearly indicates that the field-

measured DCE and VC values significantly diverge from the model profiles. In particular,

the field-measured VC concentration levels are much higher than those predicted by the

model either with or without biodegradation. As illustrated in Fig. 7, chlorinated ethene

compounds, particularly VC, can be produced as a by-product of chlorinated ethane

degradation. These by-product effects are quantified in the next set of simulations.

6.3. Simulation 3—coupled chlorinated ethane and ethene reactions

BIOCHLOR model results for the ethane series (shown in Fig. 8) indicates that TCA

concentration levels decrease from over 300 to less than 1 mg/l within 150 ft down

gradient of the source. Field data also confirm that the TCA concentration reduces to less

than 1 mg/l close to the source zone. Since degradation of TCA can yield ethene

compounds, it is possible that significant amounts of chlorinated ethenes are being

produced through biological processes within the 300-ft region from the source.

The NAPL present at the Brooklawn site contains large amounts of TeCA. As shown in

Fig. 7, degradation of TeCA would yield chlorinated ethene compounds. At biologically


Fig. 9. Comparison of field data against BIOCHLOR simulations PCE and its daughter products (t=25 years).

T.P.Clem

entet


156

active field sites, it is reasonable to assume that TeCAwill be transformed within the same

area of the aquifer as TCA because typical degradation rates of TeCA transformation are

equal to or greater than the rates observed for TCA transformation (Lorah and Olsen,

1999). Field measured TeCA and TCA concentration levels at the Brooklawn site also

appear to support this assumption.

To evaluate the impacts of transformation of TeCA and TCA on the transport of

chlorinated ethene species, Simulation 3 was conducted using an ‘‘effective’’ ethene

source. Simulation 3 assumes rapid ethane degradation close to the NAPL zone. By

invoking this assumption, the concentration levels of the source are set to be equal to the

sum of concentrations used in Simulation 2 and the maximum amount of ethene

concentrations that could be produced as byproducts of ethane biodegradation (Table 8,

column 2+colum3 3). Results of this simulation are summarized in Fig. 10. Since the

source concentration of the first ethene species, PCE, was identical to the value used in

Simulation 2, PCE results are not discussed here.

Comparison of TCE profiles shown in Fig. 10 against the profiles shown in Fig. 9

indicates that the Simulation 3 profiles are not significantly different from those predicted

by Simulation 2. This result is as expected because, as illustrated in Fig. 7, TCE

concentrations are not significantly impacted by chlorinated ethane biodegradation

products. The predicted profiles of DCE and VC compounds, however, are impacted

considerably. As shown in Fig. 10, use of the ‘‘effective’’ ethane by-product-based ethene

sources yielded model results that closely match DCE and VC field data.

6.4. Determination of steady-state plume conditions

Additional simulations using conditions similar to those used in Simulations 1 through

3 were completed for 50- and 100-year periods. The purpose of these simulations was to

assess whether the contaminant plumes would reach steady-state conditions and cease

migrating prior to reaching the exposure points. Although results are available for all the

species and for all three simulation conditions (Truex et al., 1999), the results for critical

contaminant species are only presented in this paper. Fig. 11 shows the predicted DCA

(using simulation conditions 1) and VC plume profiles (using simulation conditions 3)

after 25, 50, and 100 years of transport. The predicted profiles with and without

biodegradation are also shown in Fig. 11. The results indicate that under active

biodegradation conditions, chlorinated solvent plumes at this site would reach steady

state after about 25 years of transport. In addition, the model also predicted that the plumes

would degrade to very low non-detectable concentration levels within 1000 ft from the

source area.

6.5. Model limitations

It is important to note that the tools and the assessment process used in this study have

several limitations. The U.S. EPA model BIOCHLOR has two inherent limitations: (1) it

averages the retardation effects and (2) it can only model a single sequential reaction

chain. These approximations add considerable uncertainly to the modeling results, and

therefore, the modeling step should simply be considered as a part of the conceptualiza-


Fig. 10. Comparison of field data against BIOCHLOR simulations ethene simulation with adjusted source

concentrations (t=25 years).


tion process. Also, it is essential to recognize that Simulation 3 only provides an estimate

of the impact of chlorinated ethane degradation on the transport of the chlorinated

ethenes.

Since BIOCHLOR assumes uniform flow, only an averaged one-dimensional velocity

can be used in the model. However, over the modeled domain, the groundwater head

distribution will vary nonlinearly, particularly near the down gradient transient boundary

close to the Mississippi River. Therefore, the averaged velocity estimate used in the study

Fig. 11. Steady-state analysis using BIOCHLOR.


could be a lower estimate and may only reflect the transport conditions closer to the

source. However, since the contaminants are degraded within a short distance from the

source (close to well P-1620-2, which was used for estimating the head gradient), this

should be a reasonable approximation. Use of a detailed numerical reactive transport code,

such as RT3D (Clement, 1997; Clement et al., 2000) coupled with a transient MOD-

FLOW-based flow model, would help perform more realistic simulations under these

transient, coupled reactive transport conditions.

7. Summary and conclusions

A field study was completed to demonstrate the steps involved in implementing the

U.S. EPA (1998) monitored natural attenuation screening protocol at a chlorinated solvent

contaminated field site. A natural attenuation assessment data set collected at a Superfund

site located in Louisiana was used in the study. Data from four monitoring wells, located

in different parts of the contaminated area, were selected to complete the natural

attenuation scoring analysis recommended by the protocol. The estimated scores are

above the acceptable minimum score proposed in the protocol, thus indicating that the

contaminants in the aquifer have the potential for natural attenuation. The background

geochemical data also indicate that the site is highly anaerobic and has large amounts of

natural carbon and, hence, has the potential for supporting chlorinated solvent degrada-

tion reactions.

A recent report by the National Research Council (NRC, 2000 p. 244) concluded,

‘‘scoring systems are susceptible to misuse and because approaches to natural attenuation

have been advanced in recent years, the committee recommends the abandonment of

scoring systems in screening site for natural attenuation. Instead, the committee recom-

mends site-specific conceptual models and footprints.’’ Clearly, the U.S. EPA (1998)

scoring process employed in the study is an empirical assessment method. Despite this

limitation, if applied carefully, the method does provide a means for systematically

assembling a diverse set of bioremediation footprints in a compact format that can be

communicated with regulators and stakeholders. However, as recommended by NRC

(2000) and U.S. EPA (1998) reports, a detailed site-specific conceptual model should

always be included to further support the result of the scoring process.

A detailed conceptual framework was developed and was used to build a computer

model for the site. The U.S. EPA’s screening tool BIOCHLOR was employed for this

purpose. As per the U.S. EPA (1998) guidelines, two MNA evaluation criteria must be

applied to assess the modeling results. These two criteria pose the following questions

(U.S. EPA, 1998): (1) Has the plume moved a shorter distance than would be expected

based on the known (or estimated) time since the contaminant release and the contaminant

velocity in groundwater, as calculated from site-specific measurements of hydraulic

conductivity and hydraulic gradient, and estimates of effective porosity and contaminant

retardation? (2) Is it likely that site contaminants are attenuating at rates sufficient to meet

remediation objectives for the site in a time period that is reasonable compared to other

alternatives? If the answers to both these questions are affirmative, then the site can

proceed with full-scale evaluation of natural attenuation.


Based on the MNA screening data and modeling results presented in this paper, the

following conclusions can be made.. Comparison of field data against model predictions for contaminant transport showed

that the dissolved plume would appear to migrate slower than would be predicted if only

adsorption-related attenuation mechanisms or if no attenuation mechanisms were acting on

the plume constituents. This result meets the first MNA evaluation criterion.. Using the available biodegradation rate information, the screening-level reactive

transport model (BIOCHLOR) results indicated that the dissolved plume would reach a

steady-state condition such that contaminants would stop migrating through the aquifer

well before reaching the identified exposure points. This result meets the second MNA

evaluation criterion.

Thus, the MNA screening criteria outlined in U.S. EPA (1998) were sufficiently met for

the Brooklawn site; therefore, MNA can be considered as one of the feasible remediation

options.

Acknowledgements

This project work was supported by the NPC Services. The authors wish to thank the

NPC Services staff members including Robert Bolger, James Spencer, and Carl Douglas

for providing project support and necessary site characterization data. This paper, in part,

was completed when Dr. Clement was at the Centre for Water Research, University of

Western Australia. We like to thank Prof. Norman Jones, Dr. Greg Davis, Mr. Colin

Johnston, and the two anonymous reviewers for their constructive comments.

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Documents

A case study for demonstrating the application of U.S. EPA ...clemept/publsihed_pdf/npcpaper.pdfA case study for demonstrating the application of U.S. EPA’s monitored natural attenuation