U.S. Department of the Interior

U.S. Geological Survey

Scientific Investigations Report 2014–5140

Prepared in cooperation with the

U.S. Environmental Protection Agency

Hydrogeologic Characterization and Assessment of 

Bioremediation of Chlorinated Benzenes and Benzene 

in Wetland Areas, Standard Chlorine of Delaware, Inc. 

Superfund Site, New Castle County, Delaware, 2009–12

Cl Cl

3

Cl Cl Cl

Cl Cl Cl Cl

Cl Cl Cl

Cl

HCB PCB 1245TeCB

1234TeCB

123TCB 12DCB

14DCB

CB B

CO + H O + HCl 2 2

124TCB

1235TeCB 135TCB 13DCB

Cl Cl Cl Cl

Cl

Cl Cl Cl Cl Cl

Cl Cl Cl

Cl

Cl Cl

Cl Cl

Cl Cl Cl

Cl Cl

Cl Cl Cl

1 3

5

5

5

5

5 2 2 4

4

*

* * *

2

2 2

3 3 2
Cover. Diagram showing possible degradation pathways for chlorobenzenes. Refer to figure 2 for explanation.
Hydrogeologic Characterization and 

Assessment of Bioremediation of 

Chlorinated Benzenes and Benzene in 

Wetland Areas, Standard Chlorine of 

Delaware, Inc. Superfund Site, New 

Castle County, Delaware, 2009–12

By Michelle M. Lorah, Charles W. Walker, Anna C. Baker, Jessica A. Teunis, 

Emily H. Majcher, Michael J. Brayton, Jeff P. Raffensperger, and 

Isabelle M. Cozzarelli

Scientific Investigations Report 2014–5140

U.S. Department of the Interior

U.S. Geological Survey

Prepared in cooperation with the 

U.S. Environmental Protection Agency 
U.S. Department of the Interior

SALLY JEWELL, Secretary

U.S. Geological Survey

Suzette M. Kimball, Acting Director

U.S. Geological Survey, Reston, Virginia: 2014

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Although this report is in the public domain, permission must be secured from the individual copyright owners to 

reproduce any copyrighted materials contained within this report.

Suggested citation:

Lorah, M.M., Walker, C.W., Baker, A.C., Teunis, J.A., Majcher, E.H., Brayton, M.J., Raffensperger, J.P., and Cozzarelli, 

I.M, 2014, Hydrogeologic characterization and assessment of bioremediation of chlorinated benzenes and benzene in 

wetland areas, Standard Chlorine of Delaware, Inc. Superfund Site, New Castle County, Delaware, 2009–12: 

U.S. Geological Survey Scientific Investigations Report 2014–5140, 89 p., http://dx.doi.org/10.3133/sir20145140.

ISSN 2328-031X (print)

ISSN 2328-0328 (online)

ISBN 978-1-4113-3848-7
iii

Contents

Abstract..............................................................................................................................................................................1

Introduction........................................................................................................................................................................2

Purpose and Scope ................................................................................................................................................4

Description of Study Area .....................................................................................................................................4

Regional and Site Hydrogeologic Setting.................................................................................................5

Summary of Contamination Extent and Wetland Remedial Alternatives ...........................................6

Background on Degradation Pathways..............................................................................................................7

Methods of Data Collection and Analysis ....................................................................................................................8

Monitoring Network ...............................................................................................................................................8

Hydrogeologic Data..............................................................................................................................................11

Water-Level Measurements .....................................................................................................................16

Slug Tests .....................................................................................................................................................16

Calculations of Hydraulic Gradient, Specific Discharge, and Linear Velocity .................................17

Groundwater and Surface-Water Geochemical Data....................................................................................18

Sample Collection.......................................................................................................................................18

Chemical Analysis.......................................................................................................................................19

Data Quality..................................................................................................................................................20

Experiments to Assess Bioremediation ............................................................................................................21

In Situ Microcosms ....................................................................................................................................21

Laboratory Bioreactors..............................................................................................................................22

Hydrogeologic Characterization ..................................................................................................................................24

Lithology of the Study Area.................................................................................................................................25

Groundwater Hydraulic-Head Distribution and Flow Direction....................................................................26

Hydraulic Conductivity.........................................................................................................................................31

Groundwater Discharge and Velocity ...............................................................................................................31

Effect of Tidal Fluctuation in Red Lion Creek....................................................................................................35

Groundwater Discharge to Red Lion Creek......................................................................................................36

Geochemical Characterization.....................................................................................................................................38

Distribution of Volatile Organic Compounds ....................................................................................................38

Spatial Distribution .....................................................................................................................................38

Temporal Distribution.................................................................................................................................42

Distribution of Redox-Sensitive and Other Geochemical Constituents.......................................................43

Spatial Distribution .....................................................................................................................................43

Temporal Distribution.................................................................................................................................47

Dense Non-Aqueous Phase Liquid (DNAPL) Extent .................................................................................................50

Previous Dense Non-Aqueous Phase Liquid (DNAPL) Investigations ........................................................50

Updated Dense Non-Aqueous Phase Liquid (DNAPL) Site Conceptual Model .........................................52

Columbia and Potomac Aquifers..............................................................................................................53

Wetland Porewater ...................................................................................................................................54

Source Assessment for Wetland Study Area ..................................................................................................55

Columbia Aquifer and Overland Spills as Sources of Contamination ...............................................55

Sediment as a Source of Contamination ................................................................................................58

 Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene................................................59

Evidence from Field Sampling Results ..............................................................................................................59

In Situ Microcosms...............................................................................................................................................62

Microbial Communities in the Wetland Sediment ..........................................................................................66

Laboratory Bioreactor Experiments ..................................................................................................................69

Summary and Conclusions............................................................................................................................................82

Acknowledgments..........................................................................................................................................................83

References Cited.............................................................................................................................................................84
iv

Figures

1. Maps showing location of Standard Chlorine of Delaware, Inc. Superfund 

Site (SCD), New Castle County, Delaware........................................................................................3

2. Diagram showing possible degradation pathways for chlorobenzenes .....................................9

3. Maps showing (A) monitoring network and locations of hydrogeologic 

sections in the wetland study area, and (B) sampling locations in the 

wetland study area, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware.....................................................................................................................................10

4. Schematic diagram of Bio-Trap (Microbial Insights, Rockford, Tennessee) 

sampler used for in situ microcosms ..............................................................................................21

5. Diagram of the anaerobic bioreactor system and sampling locations......................................24

6. Cross section showing hydrogeology and locations of monitoring wells 

and drive-point piezometers along section A–A’, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware......................................................................................26

7. Cross section showing hydrogeology and locations of monitoring wells 

and drive-point piezometers along section B–B’, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware......................................................................................27

8. Map showing groundwater altitude (in meters above mean sea level) 

in the Columbia aquifer, August 8, 2012, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware .........................................................................................................28

9. Map showing groundwater altitude (in meters above mean sea level) 

in the wetland sediments, August 8, 2012, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware .........................................................................................................29

10. Graph showing water-level altitudes of Red Lion Creek and selected 

drive-point piezometers and daily total precipitation, July 2008–August 2012, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware................................................30

11. Cross section showing head distributions and groundwater-flow direction 

in section A–A’, August 8, 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware .................................................................................................................32

12. Cross section showing head distributions and groundwater-flow direction 

in section B–B’, August 8, 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware .................................................................................................................33

13. Hydrograph showing water-level altitudes at selected wetland-aquifer 

pairs and other drive-point piezometers screened in the Columbia aquifer, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware................................................35

14. Map showing location of Red Lion Creek watershed, U.S. Geological 

Survey streamgage 01482298 (Red Lion Creek near Red Lion, Delaware), 

U.S. Geological Survey stage gage 01482320 at River Road Bridge, and 

the watersheds used to estimate discharge to tidal Red Lion Creek ........................................37

15. Maps showing distribution of volatile organic compounds from passive 

diffusion bags deployed in the wetland sediments in the study area, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, 

(A) March 2011, and (B) October–December 2011........................................................................39

16. Graphs showing molar composition of volatile organic compounds 

and total concentrations of chlorobenzene compounds and benzene 

(CBs+benzene) in passive diffusion bag samples collected from the 

northwest and northeast wetland areas during (A, B) March 2011, and 

(C, D) October–December 2011, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware .................................................................................................................41

17. Graphs showing vertical distributions of volatile organic compounds 

(VOCs) from peepers in the northwest and northeast wetland areas 

(A, B) July 2009, and (C, D) October–December 2011, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware .................................................................................42
v

18. Graphs showing temporal changes in molar composition of volatile 

organic compounds and total concentrations of chlorobenzene 

compounds and benzene (CBs+benzene) at site 8 in the northwest 

wetland area from (A) peeper sampling in July 2009, (B) peeper sampling 

in December 2011, and (C) monthly sampling events, September 2011–

August 2012, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware ........................43

19. Graphs showing temporal changes in molar composition of volatile 

organic compounds and total concentrations of chlorobenzene 

compounds and benzene (CBs+benzene) at site 6 in the northeast 

wetland area from (A) peeper sampling in July 2009, (B) peeper 

sampling in December 2011, and (C) monthly sampling events, 

October 2011–August 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware .................................................................................................................44

20. Boxplots showing distributions of pH in the wetland study area 

(A) March 2011, and (B) October–December 2011 synoptic sampling 

results, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware..................................45

21. Graphs showing distributions of chloride, sulfate, ammonia, sulfide, 

methane, and ferrous iron in passive diffusion bag (PDB) samples of 

wetland groundwater collected in the nothwest and northeast areas 

in (A, B) March 2011, and (C, D) October 2011, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware......................................................................................46

22. Graphs showing spatial distribution of nonvolatile dissolved organic 

carbon (NVDOC) in passive diffusion bag (PDB) samples of the wetland 

groundwater collected in (A) northwest, and (B) northeast wetland 

areas in October 2011, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware.....................................................................................................................................47

23. Graphs showing vertical distributions of chloride concentrations in 

peepers from the northwest and northeast wetland areas in 

(A, B) July 2009, and (C, D) October–December 2011, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware .................................................................................48

24. Graphs showing vertical distributions of sulfate concentrations in 

peepers from the northwest and northeast wetland areas in 

(A, B) July 2009, and (C, D) October–December 2011, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware .................................................................................49

25. Graphs showing temporal changes in (A) chloride and (B) sulfate for 

selected sites in the wetland study area, July 2009–August 2012, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware................................................50

26. Graphs showing temporal changes in concentrations of chloride and 

redox-sensitive constituents at site 8 in the northwest wetland area 

from (A) peeper sampling in July 2009, (B) peeper sampling in 

December 2011, and (C) monthly sampling events, September 2011–

August 2012, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware ........................51

27. Graphs showing temporal changes in concentrations of chloride and 

redox-sensitive constituents at site 6 in the northeast wetland area 

from (A) peeper sampling in July 2009, (B) peeper sampling in 

December 2011, and (C) monthly sampling events, September 2011–

August 2012, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware ........................52

28. Graph showing reduction in total volatile organic compound concentrations 

in monitoring wells along the wetland-upland boundary after the installation 

of the barrier wall, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware..............................................................................................................................................53

29. Map showing potential current (2012) distribution of dense non-aqueous 

phase liquids (DNAPLs) in the Columbia aquifer, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware......................................................................................54
vi

30. Map showing potential current (2012) distribution of dense non-aqueous 

phase liquids (DNAPLs) in the wetland porewater, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware, on the basis of computed 

concentrations that exceed 1 percent of the compound’s solubility.........................................55

31. Graphs showing comparison of molar composition (in percent) of water 

samples from aquifer monitoring wells in (A) 1993 and 2004, and 

(B) 2004 and 2012, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware..............................................................................................................................................56

32. Graph showing comparison of molar composition (in percent) in the 

Columbia aquifer immediately below the wetland sediments and 

wetland porewater, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, 2012...........................................................................................................................57

33. Graph showing comparison of volatile organic compound molar ratios 

(micromolar/micromolar) in the Columbia aquifer and wetland porewater 

for the 1986 spill compounds (1,4-dichlorobenzene/1,2,4-trichlorobenzene+

1,2,3-trichlorobenzene), Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, 2012...........................................................................................................................58

34. Graphs showing comparison of (A) molar composition (in percent) and 

(B) molar ratios (micromolar/micromolar) in the Columbia and Potomac 

aquifers, using average concentrations from 2011–12, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware .................................................................................59

35. Graphs showing (A) molar composition of volatile organic compounds 

(VOCs), and (B) concentrations of redox-sensitive constituents in 

standard in situ microcosms (MNA, monitored natural attenuation 

treatment; Lactate, lactate/chitin biostimulated treatment; WBC-2, 

bioaugmented treatment) deployed in the wetland sediment at site 

DP-6, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, 

October 20–December 16, 2009........................................................................................................63

36. Graph showing molar composition of volatile organic compounds (VOCs) 

in standard in situ microcosms (MNA, monitored natural attenuation 

treatment; Lactate, lactate/chitin biostimulated treatment; WBC-2, 

bioaugmented treatment) deployed in the wetland sediment at four sites, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, 

October 26–December 14, 2010........................................................................................................63

37. Graphs showing (A) removal of 13C-chlorobenzene, (B) incorporation 

of 13C in dissolved inorganic carbon (DIC) and phospholipid fatty acid 

(PLFA) biomass, and (C) enrichment of 13C in the PLFA biomass in the 13C-chlorobenzene amended in situ microcosms (MNA, monitored 

natural attenuation; WBC-2, bioaugmented treatment) deployed in the 

wetland sediment at site DP-6, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware, October 20–December 16, 2009........................................................64

38. Graphs showing (A) incorporation of 13C in dissolved inorganic carbon 

(DIC) and phospholipid fatty acid (PLFA) biomass, and (B) enrichment 

of 13C in the PLFA biomass in the 13C-chlorobenzene amended in situ

microcosms (MNA, monitored natural attenuation; Lactate, lactate/chitin 

biostimulated treatment; WBC-2, bioaugmented treatment) deployed 

in the wetland sediment at three sites, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware, October 26–December 14, 2010................................................65

39. Graph showing microbial populations determined by quantitative 

polymerase chain reaction (qPCR) analysis of wetland sediment 

collected at sites DP-9 and DP-6 and in the control (MNA, monitored 

natural attenuation) in situ microcosm deployed in the wetland 

sediment at site DP-6, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, October 20–December 16, 2009 ...........................................................................68
vii

40. Graph showing microbial community composition determined from 

analysis of the chemical structures of phospholipid fatty acid (PLFA) 

in wetland sediment collected at sites DP-9 and DP-6 and in the control 

(MNA, monitored natural attenuation) in situ microcosm deployed in 

the wetland sediment at site DP-6, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware, October 20–December 16, 2009........................................................68

41. Graphs showing amount of (A) total Eubacteria, and (B) Dehalococcoides

species determined by quantitative polymerase chain reaction (qPCR) 

analysis of samples from standard and monochlorobenzene (13C-CB) or 

1,4-dichlorobenzene (13C-DCB) in situ microcosms (MNA, monitored 

natural attenuation treatment; Lactate, lactate/chitin biostimulated 

treatment; WBC-2, bioaugmented treatment) deployed in the wetland 

sediment at four sites, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, October 26–December 14, 2010 ...........................................................................70

42. Concentrations of total chlorobenzene compounds and benzene 

(total CBs+benzene), naturally input from the wetland groundwater, 

and amended 1,1,2,2-tetrachloroethane (TeCA) and its anaerobic 

daughter products in (A) bioreactor C, which was seeded only with 

native microorganisms from the site groundwater, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware, and in bioreactors (B) D, 

and (C) E, which were pre-established with the WBC-2 anaerobic culture.............................71

43. Graphs showing pH measured in (A) bioreactor C, which was seeded 

only with native microorganisms from the site groundwater, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware, and in bioreactors 

(B) D, and (C) E, which were pre-established with the WBC-2 anaerobic 

culture...................................................................................................................................................72

44. Graphs showing oxidation-reduction potentials (ORPs) in (A) bioreactor C, 

which was seeded only with native microorganisms from the site 

groundwater, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware, and in bioreactors (B) D, and (C) E, which were pre-established 

with the WBC-2 anaerobic culture ..................................................................................................73

45. Graphs showing sulfate concentrations measured in (A) bioreactor C, 

which was seeded only with native microorganisms from the site 

groundwater, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware, and in bioreactors (B) D, and (C) E, which were pre-established 

with the WBC-2 anaerobic culture ..................................................................................................74

46. Sulfate and methane concentrations measured in (A) bioreactor C, which 

was seeded only with native microorganisms from the site groundwater, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, and in 

bioreactors (B) D, and (C) E, which were pre-established with the 

WBC-2 anaerobic culture..................................................................................................................75

47. Boxplots showing percent removals of (A) 1,2,3-trichlorobenzene (123TCB), 

(B) 1,2,4-trichlorobenzene (124TCB), (C) 1,4-dichlorobenzene (14DCB), 

(D) 1,3-dichlorobenzene (13DCB), (E) 1,2-dichlorobenzene (12DCB), 

(F) chlorobenzene, and (G) benzene that were calculated from inflow 

and outflow concentrations under anaerobic and aerobic conditions 

in bioreactors C, D, and E ..................................................................................................................77

48. Boxplots showing hydraulic residence time in bioreactors C, D, and E....................................78

49. Graphs showing concentrations of chlorobenzene compounds and 

benzene measured under anaerobic conditions (except in bioreactor C 

on August 13, 2012) at sampling ports along flowpaths through bioreactors 

C, D, and E on selected sampling dates..........................................................................................79
viii

Tables

1. Summary of significant fate and transport properties of contaminants 

in wetland study area, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware.......................................................................................................................................8

2. Monitoring network for the wetland study area, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware......................................................................................12

3. Groundwater sampling and site events that may have influenced sampling 

results in the wetland study area, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware .................................................................................................................15

4. In situ microcosm treatments deployed from October 20–December 16, 

2009 and from October 26–December 14, 2010 at selected wetland sites, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware................................................22

5. Description of phospholipid fatty acid (PLFA) structural groups that are 

used to describe microbial community composition on the basis of fatty 

acid biosynthesis ................................................................................................................................23

6. Calculated vertical gradients from selected drive-point piezometer pairs 

in the Columbia aquifer and wetland sediments, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware......................................................................................31

7. Maximum and minimum estimates of hydraulic conductivity calculated 

from slug tests conducted at selected drive-point piezometers in the 

Columbia aquifer and wetland sediments, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware .........................................................................................................34

8. Variation in vertical specific discharge and linear velocity of groundwater 

from the Columbia aquifer through wetland sediments, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware .................................................................................34

9. Spatial variation in lateral groundwater specific discharge and linear 

velocity of groundwater at Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware.....................................................................................................................................34

10. Area and measured or estimated annual mean discharge for subwatersheds 

in the Red Lion Creek watershed, and estimated groundwater discharge to 

tidal Red Lion Creek............................................................................................................................38

11. Concentrations of selected inorganic constituents measured in groundwater 

samples collected with passive diffusion bags from upland monitoring wells, 

October 2011........................................................................................................................................49

12. Concentrations of chloride and sulfate measured in 2008 in drive-point 

piezometers screened in the Columbia aquifer near the upland-wetland 

boundary east of the Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware..............................................................................................................................................50

13. Traveltime of selected site contaminants in groundwater along a flowpath ...........................57

14. Comparison of calculated and measured porewater volatile organic 

compound concentrations ................................................................................................................60

15. Concentrations of redox-sensitive constituents and volatile fatty acids 

measured in samples from in situ microcosms deployed from October 20–

December 16, 2009 and from October 26–December 14, 2010 at selected 

wetland sites, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware......................66

16. Amount of 13C incorporation into phospholipid fatty acid (PLFA) biomass 

and dissolved inorganic carbon (DIC) measured in in situ microcosms 

that were preloaded with 13C-labeled 1,4-dichlorobenzene or benzene 

and deployed from October 26–December 14, 2010 at selected wetland sites .......................67

17. Summary of microbial species identified using denaturing gradient gel 

electrophoresis (DGGE) analysis of samplers from the monitored natural 

attenuation (MNA) and lactate/chitin biostimulated treatments of in situ

microcosms deployed from October–December 2010 .................................................................69
ix

Multiply By To obtain

Length

centimeter (cm) 0.3937 inch (in.)

meter (m) 3.281 foot (ft) 

kilometer (km) 0.6214 mile (mi)

kilometer (km) 0.5400 mile, nautical (nmi) 

meter (m) 1.094 yard (yd) 

Area

square meter (m2) 0.0002471 acre 

square kilometer (km2) 247.1 acre

square meter (m2) 10.76 square foot (ft2) 

square kilometer (km2) 0.3861 square mile (mi2)

Volume

liter (L) 0.2642 gallon (gal)

cubic meter (m3) 0.0002642 million gallons (Mgal) 

cubic centimeter (cm3) 0.06102 cubic inch (in3) 

cubic meter (m3) 35.31 cubic foot (ft3)

cubic meter (m3) 1.308 cubic yard (yd3) 

cubic kilometer (km3) 0.2399 cubic mile (mi3) 

cubic meter (m3) 0.0008107 acre-foot (acre-ft) 

Flow rate

cubic meter per second (m3/s) 70.07 acre-foot per day (acre-ft/d) 

cubic meter per year (m3/yr) 0.000811 acre-foot per year (acre-ft/yr) 

meter per second (m/s) 3.281 foot per second (ft/s) 

meter per minute (m/min) 3.281 foot per minute (ft/min) 

meter per hour (m/hr) 3.281 foot per hour (ft/hr)

meter per day (m/d) 3.281 foot per day (ft/d)

meter per year (m/yr) 3.281 foot per year ft/yr) 

cubic meter per second (m3/s) 35.31 cubic foot per second (ft3/s)

cubic meter per day (m3/d) 35.31 cubic foot per day (ft3/d) 

liter per second (L/s) 15.85 gallon per minute (gal/min) 

cubic meter per day (m3/d) 264.2 gallon per day (gal/d) 

cubic meter per day per square 684.28 gallon per day per square mile 

kilometer [(m3/d)/km2] [(gal/d)/mi2]

cubic meter per second (m3/s) 22.83 million gallons per day (Mgal/d) 

cubic meter per day per square

kilometer [(m3/d)/km2]

0.0006844 million gallons per day per square 

mile [(Mgal/d)/mi2]

cubic meter per hour (m3/h) 39.37 inch per hour (in/h)

Mass

gram (g) 0.03527 ounce, avoirdupois (oz)

kilogram (kg) 2.205 pound avoirdupois (lb)

Pressure

kilopascal (kPa) 0.009869 atmosphere, standard (atm)

kilopascal (kPa) 0.01 bar

kilopascal (kPa) 0.2961 inch of mercury at 60 °F (in Hg)

kilopascal (kPa) 0.1450 pound-force per inch (lbf/in) 

kilopascal (kPa) 20.88 pound per square foot (lb/ft2) 

kilopascal (kPa) 0.1450 pound per square inch (lb/ft2) 

Conversion Factors and Datums
x

Multiply By To obtain

Density

kilogram per cubic meter (kg/m3) 0.06242 pound per cubic foot (lb/ft3) 

gram per cubic centimeter (g/cm3) 62.4220 pound per cubic foot (lb/ft3) 

Hydraulic conductivity

meter per day (m/d) 3.281 foot per day (ft/d) 

Hydraulic gradient

meter per kilometer (m/km) 5.27983 foot per mile (ft/mi) 

Transmissivity

meter squared per day (m2/d) 10.76 foot squared per day (ft2/d) 

Vertical coordinate information is referenced to the National Geodetic Vertical Datum of 1929 

(NGVD 29).

Horizontal coordinate information is referenced to the North American Datum of 1983 (NAD 83).

Concentrations of chemical constituents in water are given either in milligrams per liter (mg/L) 

or micrograms per liter (μg/L). 


Hydrogeologic Characterization and Assessment of 

Bioremediation of Chlorinated Benzenes and Benzene 

in Wetland Areas, Standard Chlorine of Delaware, Inc. 

Superfund Site, New Castle County, Delaware, 2009–12

By Michelle M. Lorah1, Charles W. Walker1, Anna C. Baker1, Jessica A. Teunis1, Emily H. Majcher2, 

Michael J. Brayton1, Jeff P. Raffensperger1, and Isabelle M. Cozzarelli1

Abstract

Wetlands at the Standard Chlorine of Delaware, Inc. 

Superfund Site (SCD) in New Castle County, Delaware, are 

affected by contamination with chlorobenzenes and benzene 

from past waste storage and disposal, spills, leaks, and contaminated groundwater discharge. In cooperation with the 

U.S. Environmental Protection Agency, the U.S. Geological 

Survey began an investigation in June 2009 to characterize the 

hydrogeology and geochemistry in the wetlands and assess 

the feasibility of monitored natural attenuation and enhanced 

bioremediation as remedial strategies. Groundwater flow in 

the wetland study area is predominantly vertically upward 

in the wetland sediments and the underlying aquifer, and 

groundwater discharge accounts for a minimum of 47 percent 

of the total discharge for the subwatershed of tidal Red Lion 

Creek. Thus, groundwater transport of contaminants to surface 

water could be significant. The major contaminants detected 

in groundwater in the wetland study area included benzene, 

monochlorobenzene, and tri- and di-chlorobenzenes. Shallow 

wetland groundwater in the northwest part of the wetland 

study area was characterized by high concentrations of total 

chlorinated benzenes and benzene (maximum about 75,000 

micrograms per liter [μg/L]), low pH, and high chloride. In the 

northeast part of the wetland study area, wetland groundwater 

had low to moderate concentrations of total chlorinated benzenes and benzene (generally not greater than 10,000 μg/L), 

moderate pH, and high sulfate concentrations. Concentrations 

in the groundwater in excess of 1 percent of the solubility 

of the individual chlorinated benzenes indicate that a contaminant source is present in the wetland sediments as dense 

nonaqueous phase liquids (DNAPLs). Consistently higher 

contaminant concentrations in the shallow wetland groundwater than deeper in the wetland sediments or the aquifer also 

indicate a continued source in the wetland sediments, which 

could include dissolution of DNAPLs and desorption from the 

sediments.

When highly reducing, methanogenic, or sulfate-reducing 

conditions existed in the wetland groundwater, molar composition of the volatile organic compounds (VOCs) showed 

that chlorobenzene and benzene were predominant, indicating 

biodegradation of the chlorinated benzenes through reductive 

dechlorination pathways. Temporal changes in redox conditions between 2009 and 2011–12 have shifted the locations 

in the wetland study area where reductive dechlorination is 

evident. Microbial community analyses of sediment showed 

relatively high cell numbers and diversity of populations 

(Dehalococcoides, Dehalobacter, Desulfitobacterium, and 

Geobacter) that are known to contain species capable of 

reductive dechlorination, confirming groundwater geochemistry evidence of the occurrence of reductive dechlorination. 

Natural attenuation was not sufficient, however, to reduce total 

VOC concentrations along upward groundwater flowpaths in 

the wetland sediments, most likely due to the additional source 

of contaminants in the upper sediments. In situ microcosms 

that were unamended except for the addition of 13C-labeled 

contaminants in some treatments, confirmed that the native 

microbial community was able to biodegrade the higher 

chlorinated benzenes through reductive dechlorination and 

that 1,2-dichlorobenzene, chlorobenzene, and benzene could 

be degraded to carbon dioxide through oxidation pathways. 

Microcosms that were bioaugmented with the anaerobic 

dechlorinating consortium WBC-2 and deployed in the wetland sediments showed reductive dechlorination of tri-, di-, 

and monochlorobenzene, and 13C-chlorobenzene treatments 

showed complete degradation of chlorobenzene to carbon 

dioxide under anaerobic conditions. 

1

 U.S. Geological Survey.

2

 Cherokee Nation Technology Solutions.
2 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Experiments with a continuous flow, fixed-film bioreactor 

seeded with native microorganisms in groundwater from the 

wetland area showed both aerobic and anaerobic biodegradation of dichlorobenzenes, monochlorobenzene, and benzene, 

although monochlorobenzene and benzene degradation rates 

decreased under anaerobic conditions compared to aerobic 

conditions. In two bioreactors with established biofilms of 

WBC-2, percent removals of all chlorinated benzene compounds (medians of 86 to 94 percent) under anaerobic conditions were as high as those observed for the bioreactors seeded 

only with native microorganisms from the site groundwater, 

and benzene removal was greater in the WBC-2 bioaugmented 

bioreactors. The high percent removals in the WBC-2 bioreactors without the need for an acclimation period indicates 

that the same dechlorinators are involved in the chlorinated 

benzene degradation as those for the chlorinated ethanes and 

ethenes that the culture was developed to degrade. The ability 

of the WBC-2 culture to completely reduce the chlorinated 

benzenes and benzene, even in the presence of high sulfate 

and sulfide concentrations, is unique for known dechlorinating 

cultures. The availability of the established culture WBC-2, as 

well as the ability of the native wetland microbial community 

to degrade the site contaminants under anaerobic and aerobic 

conditions, provides flexibility in considering bioremediation 

options for the wetland areas at SCD. 

Introduction

Major releases of chlorinated benzenes and benzene 

occurred at the Standard Chlorine of Delaware, Inc. Superfund 

Site (SCD) in New Castle County, Delaware, from 1966–2002, 

resulting in contamination of the groundwater underlying the 

site and the wetlands surrounding Red Lion Creek (fig. 1). The 

U.S. Environmental Protection Agency (EPA) issued a Record 

of Decision (ROD) for SCD in 1995 that included an interim 

action for containment of groundwater and treatment for soils 

and sediments either by bioremediation or thermal desorption. 

Predominant contaminants identified at the site include 

1,2,3-trichlorobenzene (123TCB), 1,2,4-trichlorobenzene 

(124TCB), 1,3,5-trichlorobenzene (135TCB), 1,2-dichlorobenzene (12DCB), 1,3-dichlorobenzene (13DCB), 1,4-dichlorobenzene (14DCB), chlorobenzene (also referred to as 

monochlorobenzene) (CB), and benzene (U.S. Environmental 

Protection Agency, 2011). Although installation of a groundwater interception and treatment system has been completed 

around the main industrial area of the SCD (designated by the 

barrier wall in fig. 1), wetland and sediment areas impacted 

by contaminated groundwater discharge and past surficial 

spills remain outside this barrier wall. Implementation of in 

situ bioremediation was identified for the wetland areas under 

the selected remediation alternative for the SCD during the 

Feasibility Study, and a laboratory treatability investigation of 

bioremediation of wetland sediments was completed in 1993 

(Roy F. Weston, Inc., 1993a). This feasibility investigation, 

however, was inconclusive on the potential bioremediation 

technologies that could be used to enhance degradation and 

their effectiveness in the wetland areas (U.S. Environmental 

Protection Agency, 1993), and the contingency remedy of low 

temperature thermal desorption was considered instead (Black 

& Veatch, 2003; HydroGeoLogic, Inc., 2009). The extremely 

high estimated cost of a thermal remedy and the recognition 

of advances in bioremediation since 1993 (Black & Veatch, 

2003; HydroGeoLogic, Inc., 2009) led to the decision to reinvestigate the potential of bioremediation for the SCD wetland 

areas. In cooperation with EPA, the U.S. Geological Survey 

(USGS) expanded its ongoing Potomac aquifer investigation 

at the site (Degnan and Brayton, 2010) and began a preliminary investigation in June 2009 of the feasibility of monitored 

natural attenuation or of combined natural attenuation and 

enhanced bioremediation as remedial strategies for the contaminated wetland areas.

Biodegradation is recognized as the key destructive 

process for successful use of monitored natural attenuation as 

a groundwater remediation method for organic contaminants 

(U.S. Environmental Protection Agency, 1998). An understanding of the native microbial communities and their possible degradation mechanisms can also be applied in engineered 

bioremediation solutions that can lead to more complete or 

rapid biodegradation at a site than through a monitored natural 

attenuation remedy (Lovley, 2003). Bioremediation techniques 

can be divided into two general categories—biostimulation, 

which involves the addition of electron acceptors or donors 

(usually an organic carbon donor for anaerobic degradation), 

nutrients, or other compounds that have been determined to 

stimulate the activity of microorganisms naturally present at 

the site, and bioaugmentation, which involves the addition 

of bacteria known to degrade the contaminants of interest. 

Bioaugmentation can be applied to increase the population 

size, and thus degradation capacity, of bacteria already present 

at the site or to add contaminant-degrading bacteria that were 

not present. 

Since the bioremediation treatability investigation was 

conducted with the wetland sediments in 1993 for the SCD, 

substantial progress has been made in the understanding of 

biodegradation processes for chlorinated solvents, including 

chlorinated benzenes, and the microorganisms involved, as 

well as in hydrogeologic investigation and bioremediation 

technologies applicable to wetland areas (Lorah and Olsen, 

1999a, b; Kassenga and others, 2003; Lorah and others, 2007; 

Lorah, Majcher, and others, 2008; Majcher and others, 2007, 

2009; Braeckevelt and others, 2008; Imfeld and others, 2008). 

Both aerobic and anaerobic degradation processes have been 

identified for chlorinated benzenes (Ramanand and others, 

1993; Adrian and others, 2007; Field and Sierra-Alvarez, 

2008). In general, higher chlorinated benzenes such as 

124TCB are degraded more readily under anaerobic conditions by reductive dehalogenation reactions, whereas DCBs 

and especially CB are degraded more readily by aerobic bacteria through an oxidation reaction. One of the most frequent 

problems with bioremediation of groundwater contaminated 
Introduction 3

with chlorinated benzenes has been the incomplete degradation of higher chlorinated benzenes to CB and the typical 

recalcitrance of CB under anaerobic conditions (Field and 

Sierra-Alvarez, 2008; Fung and others, 2009). Reductive 

dehalogenation of DCBs to CB and of CB to benzene has 

been reported primarily in laboratory studies (Fung and others, 

2009), although a few field studies, including a wetland study, 

have recently reported evidence of anaerobic degradation 

of CB (Kaschl and others, 2005; Kästner and others, 2006; 

Braeckevelt and others, 2008). Because of the abundance of 

organic carbon substrates and the influence of wetland plants, 

which transport oxygen to the sediment through their roots, 

wetland sediments typically contain a high diversity of redox 

zones and microbial populations that can support a variety of 

contaminant degradation pathways (Amon and others, 2007; 

Lorah and others, 1997). Anaerobic and aerobic degradation 

reactions have been shown to occur in the same wetland sediment for chlorinated ethenes (Lorah and others, 2001) and for 

CB (Braeckevelt and others, 2007; Braeckevelt and others, 

2008).

Anaerobic degradation of chlorinated benzenes occurs 

by reductive dechlorination, the same process identified 

for chlorinated ethenes and chlorinated ethanes, such as 

trichloroethene and tetrachloroethane, respectively. In fact, 

FALL LINE

Base from U.S. Geological Survey

digital data, 2004, 1:24,000

DELAWARE RIVER

C&D CANAL

NEW

CASTLE

KENT

COASTAL

PLAIN

SUSSEX

Wilmington

Delaware City

Dover DELAWARE

DELAWARE

MARYLAND

MARYLAND

NEW JERSEY

DELAWARE

B AY

39°30'

39°00'

38°30'

75°30' 75°00' 75°38'20" 75°38'10" 75°38'00" 75°37'50"

PENNSYLVANIA

Indian River

Bay

Rehoboth Bay

Georgetown

Red Lion

Creek

Dragon

Creek

Location of Standard

Chlorine of Delaware, Inc.

Superfund Site (SCD)

0

0 6 MILES

10 KILOMETERS

39°

36'

20"

39°

36'

10"

39°

36'

00"

39°

35'

50"

Red Lion Creek

BARRIER WALL

WESTERN

DRAINAGE

GULLEY

SETTLING

POND

EASTERN

DRAINAGE

DITCH

FORMER

AIR PRODUCTS

FFACILITY

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE (SCD)

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE (SCD)

Unnamed 

Tributary

OCCIDENTAL CHEMICAL

FFACILITY

Governor Lea Road

CATCH

BASIN

NO. 1

9

9

0

0

0.20 KILOMETERS

0.10 MILES

Imagery from U.S. Geological Survey,

Delaware Valley Regional Planning

Commission, and Department of Homeland

Security, 2010, NAD 1983 State Plane

Delaware Transverse Mercator Projection

NN

River Road River Road

MAIN INDUSTRIAL

AREA OF SCD

Haul Road

Figure 1. Location of Standard 

Chlorine of Delaware, Inc. Superfund 

Site (SCD), New Castle County, 

Delaware (modified from Degnan 

and Brayton, 2010).
4 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

many of the anaerobic bacteria that have been identified 

as able to degrade chlorinated benzenes are in the same 

genera (such as Dehalococcoides, Dehalobacter, and 

Desulfitobacterium) that have been extensively studied for 

reductive dechlorination of the chlorinated ethenes and ethanes 

(Hölscher and others, 2003). For example, Dehalococcoides

sp. strain CBDB1 has been shown to derive energy from 

dechlorination of 123TCB, 124TCB, all tetrachlorobenzene 

isomers, pentachlorobenzene, and hexachlorobenezene, as 

well as polychlorinated dibenzodioxins (Bunge and others, 

2003; Adrian and others, 2007). Although reductive dehalogenation of DCBs has been observed, no specific bacteria that 

derive energy from this reaction have been identified (Fung 

and others, 2009). Even though several anaerobic microbial 

cultures are commercially available for degradation of chlorinated ethenes and ethanes, none are currently available for 

degradation of chlorinated benzenes. 

The USGS wetland study at SCD includes two major 

objectives: (1) wetland characterization, including determination of the hydrogeology and the distributions of groundwater 

contaminants and other geochemical constituents, and (2) 

determination of biodegradation processes naturally occurring 

in the wetland sediment and the effects of amendments to 

enhance biodegradation, including biostimulation and bioaugmentation. An anaerobic consortium, called WBC-2, that 

was developed by the USGS from wetland sediment at a 

site in Maryland to degrade chlorinated ethanes and ethenes, 

was used for bioaugmentation testing in this study. WBC-2 

contains several Dehalococcoides strains, including CBDB1, 

as well as Dehalobacter, Acetobacteria, and Bacteroidetes

species that have been identified by others in chlorobenzene 

degrading communities (Geosyntec Consultants Inc., 2006; 

Lorah and others, 2008, Lorah and others, 2012). This report 

describes the results of field and laboratory investigations that 

occurred between 2009–12 for the ongoing wetland study at 

the SCD. The goal of the wetland study was to assist the EPA 

in identifying an effective remedial strategy for the contaminated wetland areas. Bioremediation can be a cost-effective 

strategy with the potential for in situ applications that would 

have minimum adverse impact on the wetland ecosystem. 

Purpose and Scope

The purpose of this report is to provide a hydrogeologic 

and geochemical characterization of the wetland study area 

and to assess the evidence for natural and enhanced biodegradation of chlorinated benzenes and benzene in the 

SCD wetland area from field and laboratory investigations 

conducted in 2009–12. The hydrogeologic characterization 

includes description of the lithology of the wetland sediments 

from sediment cores, evaluation of head distributions and 

groundwater-flow directions from synoptic and continuous 

water-level measurements in piezometers, and calculation of 

groundwater-flow velocities using head distributions and estimates of hydraulic conductivities from slug tests. In addition, 

a preliminary estimate of the groundwater discharge rate to 

Red Lion Creek was made to assess the potential significance 

of groundwater on contaminant transport to surface water. The 

geochemical characterization includes evaluation of the distributions of volatile organic compounds (VOCs) and of redoxsensitive and other geochemical constituents in the wetland 

groundwater using samples collected from passive diffusion 

samplers and piezometers. Samples collected during synoptic 

sampling of many sites across the wetland were used to assess 

areal distributions of constituents, whereas samples collected 

monthly at selected sites were used to assess temporal changes 

in the wetland geochemistry. 

Field and laboratory studies were used to obtain evidence 

of natural or enhanced biodegradation processes. For field 

investigations, ratios of parent and potential daughter VOCs in 

different areas of the wetland were calculated with geochemical data from piezometers and 1.22-meter (m; 4-feet [ft])-long 

peepers (porous diffusion samplers with 22 rows of sampling 

ports) and compared to ratios observed in upland wells and 

sediment samples that were indicative of the contaminant 

sources. In situ microcosms that included treatments to evaluate monitored natural attenuation, biostimulation with organic 

donor, and bioaugmentation with WBC-2 also were used to 

provide field evidence of biodegradation. Microbial community analysis of samples collected from the in situ microcosms 

and the wetland sediment provided distributions of possible 

microbial species involved in contaminant degradation. 

Anaerobic bioreactors, operated under flow-through conditions 

with varying amendments, were used to provide laboratory 

evidence of natural and enhanced biodegradation processes by 

comparing reactors seeded with WBC-2 to a control bioreactor seeded only with native microorganisms derived from the 

inflow of site groundwater.

Description of Study Area

The area of the USGS wetland study at SCD is composed 

of wetlands in the northern extent of the SCD boundary along 

the southern border of Red Lion Creek, as well as wetlands 

that extend along the northern border of Red Lion Creek, 

primarily north of the installed barrier wall and east of the 

unnamed tributary of Red Lion Creek (fig. 1). In order to more 

fully understand factors that contribute to the geochemistry 

and hydrogeology of the wetlands, a part of the bordering 

uplands were included in the study area, particularly along 

Haul Road (fig. 1). The USGS study area is entirely contained 

within the boundaries of the SCD and largely contained within 

what was referred to as the “eastern wetlands” in the Remedial 

Approach to Wetlands investigation conducted in 2009 

(HydroGeoLogic, Inc., 2009). The SCD is a 0.26 square kilometer (km2

) former chemical manufacturing facility bounded 

to the north by Red Lion Creek, to the east by Occidental 

Chemical Corporation (and Delaware Route 9), to the west by 

the former Air Products, Inc. and an unnamed tributary of Red 

Lion Creek, and to the south by Governor Lea Road (fig. 1). 
Introduction 5

Environmental responses, investigations, and remedial actions 

have been ongoing at the SCD since 1981, and it was placed 

on the EPA National Priority List in 1987. 

Subsurface contamination is a result of on-site production 

of chlorinated benzenes in the central-upland part of the SCD 

from 1966–2002. A variety of chlorinated benzenes (CB, 

14DCB, 12DCB, and lesser amounts of 13DCB, 123TCB, 

and 124TCB) were produced by combining (through reaction 

and distillation) chlorine and benzene, which were purchased 

from adjacent industrial facilities and stored and disposed at 

the SCD until closure in 2002. Numerous spills and longterm leaks contributed to subsurface contamination that has 

extended spatially to the wetland study area and Red Lion 

Creek. 

Regional and Site Hydrogeologic Setting

Three distinct lithologic formations are predominant in 

the vicinity of the SCD—coarse to medium sands and gravels 

of the Quaternary Columbia Formation, followed by marine 

fine silt and clay of the Upper Cretaceous Merchantville 

Formation, and then alternating layers of clay, silt, and fine to 

medium sand of the Cretaceous Potomac Formation (Ramsey, 

2005). Pleistocene erosion due to lowering of sea level 

resulted in downcutting of rivers into the Potomac Formation. 

These channels were refilled with undifferentiated sediments 

(sand, gravel, and clay) of the Columbia Group (Phillips, 

1987). Holocene sediments overlying the Columbia also are 

present as marsh deposits and consist of black to dark grey 

organic rich silty clay with beds of peat ranging in thickness 

between 0.3 and 12 m (Ramsey, 2005). 

The water-table aquifer in the Columbia Formation 

(referred to herein informally as the Columbia aquifer) is 

12–23 m thick locally, discontinuous, and infrequently used 

for small domestic water supplies (Roy F. Weston, Inc., 1992). 

Groundwater in the water-table aquifer is recharged by precipitation and generally discharges to bordering wetlands and 

surface-water features. The depth to water ranges from about 

6.1 m below land surface in the uplands to near land surface in 

the wetlands, which have an elevation close to mean sea level. 

At the SCD, the direction of flow for shallow groundwater is 

generally from south to north. Deeper groundwater is present 

in confined and semi-confined aquifers that vary in thickness 

and spatial extent. Historical long-term water use has led to a 

documented regional cone of depression that affects all levels 

of the Potomac aquifer (Martin, 1984) and water levels within 

various Potomac aquifer sands can vary depending on production well demands. Increases in salinity in the Potomac aquifer 

also have been documented due to the strong influence of 

production wells on flow direction. Vertical flow gradients of 

the deeper flow system in the area are consistently downward 

from the Columbia to the Potomac aquifer, and downward 

between sand layers within the Potomac Formation 

(Martin, 1984). 

The Merchantville Formation clay, stratigraphically 

beneath the Columbia aquifer, was initially thought to be a 

nearly continuous impermeable layer that prevented contamination from migrating deeper at the site. This conclusion 

was partially drawn from the lack of water-level response in 

Columbia wells to an aquifer test conducted in 1990 (Roy F. 

Weston, Inc., 1992). Further investigations indicated that the 

clay is discontinuous and that holes may have been eroded 

through the clay by paleochannels prior to and during the 

deposition of the Columbia Formation (Roy F. Weston, Inc., 

1992; Degnan and Brayton, 2010). 

Flow from the upland part of the SCD to the USGS study 

area is impeded by an approximately 1-m-thick soil-bentonite 

containment barrier wall that extends up to 22 m below land 

surface. This barrier wall was installed in 2007 to prevent 

migration of contaminants from the upland part of the site 

downgradient toward Red Lion Creek (U.S. Environmental 

Protection Agency, 2011) in compliance with the ROD (U.S. 

Environmental Protection Agency, 1995a). The barrier wall 

was installed into the confining units of the Merchantville clay 

and (or) upper Potomac sequences, although these units have 

been demonstrated to be laterally variable in thickness and, in 

the case of the Merchantville, discontinuous in the vicinity of 

the site. 

Groundwater withdrawals from the Potomac aquifer for 

industrial and public supply use occur within 5 kilometers 

(km) of the SCD. Public supply use is typically 0.011 cubic 

meters per second (m3/s), or 0.25 million gallons per day 

(Mgal/d), whereas aggregate industrial withdrawal by the 

Delaware City Refining Company (DCRC) is typically 

0.24 m3/s (5.5 Mgal/d) (Doug Rambo, Delaware Department 

of Natural Resources and Environmental Control, written 

commun., 2013). Water-level declines up to 61 m below sea 

level have been documented in parts of the Potomac aquifer 

since industrial withdrawals began in the 1950s (Martin, 

1984). Most of the decline near the SCD is due to long-term 

sustained use by the DCRC. No impact on the shallow flow 

system in the Columbia aquifer has been observed from industrial pumping; however, lower head values in the Potomac 

aquifer relative to the Columbia aquifer have enhanced a vertically downward groundwater gradient in the deeper flow system of the Columbia aquifer. Historically, aggregate groundwater use by the refinery has been fairly constant, however, 

production wells were inactive during an ownership transition 

period from May 2010 to January 2011. During this period, 

selected wells were occasionally pumped to ensure that the 

pumps remained in good working condition, with only very 

limited withdrawals to support plant maintenance (Rebecca 

Gudgeon, DCRC, oral commun., 2011). Industrial groundwater withdrawal wells located closest to the SCD (wells P-6A, 

P-5B, and R-15) typically pump a combined 0.044 m3/s during 

normal operation. These wells were operated in limited capacity from December 2009 to May 2010, pumping less than 

378 cubic meters per day (m3/d) from when the plant was idled 

until January 2011. A groundwater recovery was observed during this period with a corresponding reduction in the magnitude of the vertically downward gradients. 
6 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Surface runoff at the SCD moves generally north into 

Red Lion Creek, or into small streams feeding the creek. 

Shallow groundwater discharges to the creek and to wetland 

areas surrounding the creek. Water from Red Lion Creek 

drains during low tide through a tide control structure located 

approximately 0.8 km downstream from Delaware State Route 

9. Starting in the late 1950s, the creek was no longer subject 

to tidal influence because gates on the tide control structure 

closed during high tide to prevent water from entering from 

the Delaware River (Mickowski, 1986). However, the creek 

and wetlands of the SCD have been tidally flooded periodically since the 1950s, including during the 1986 spill discussed below. Despite several attempts at repair, the tide 

control structure is currently (June 2014) not functioning and 

approximately one-third of the wetlands are submerged under 

high tide conditions. 

Summary of Contamination Extent and Wetland 

Remedial Alternatives 

Industrial activities at the SCD resulted in extensive 

contamination of groundwater, soils, and sediments that has 

been documented over the last 30 years (Roy F. Weston, Inc., 

1993a, Black & Veatch, 2007; HydroGeoLogic, Inc., 2009; 

Christopher Wolfe, HydroGeoLogic, Inc., written. commun., 

2012) and led to required interim and final remedial actions, 

as specified in the site ROD (U.S. Environmental Protection 

Agency, 1995a; 2011). Subsurface contamination at the SCD 

is likely the result of leaks and several spills of chlorinated 

benzenes, including a leaking catch basin that was discovered 

and repaired in 1976, a spill of 5,000 gallons of CB in 1981, 

and a much larger spill of 400,000 gallons of 14DCB and 

169,000 gallons of TCBs in 1986 (U.S. Environmental 

Protection Agency, 1995a). In addition to the known compounds released in spills, previous studies have detected 

hexachlorobenzne (HCB) and tetrachlorobenzenes (TeCBs) as 

compounds of concern in soil and sediment at the site (U.S. 

Environmental Protection Agency, 1995a). All of the known 

contaminant releases were upgradient of the wetland study 

area. The CB that was spilled in 1981 was the result of an 

overflowing railroad tanker. Spill contents traveled along the 

railroad tracks and entered the drainage gully, which led to 

Red Lion Creek (Standard Chlorine of Delaware, Inc., 1982). 

Booms and dams were placed to prevent migration of CB 

to Red Lion Creek, and soils and sediments were excavated 

in the vicinity of the unnamed tributary. However, further 

investigation determined that the creek was impacted with 

chlorinated benzenes, and contamination was found in site 

groundwater and creek bottom sediments (Standard Chlorine 

of Delaware, Inc., 1982). In the case of the 1986 spill, the wetlands were directly impacted by released product because the 

spill occurred during high tide and resulted in a “fan” of product being swept into the wetlands and along the southern bank 

of Red Lion Creek (Roy F. Weston, Inc., 1988). Similar to the 

1981 spill, product also traveled along the unnamed tributary 

to Red Lion Creek (first through the western drainage gully) 

and entered the wetlands to the west of the USGS study area 

and along its western boundary. Remedial actions conducted 

after the 1986 spill included the construction of booms and 

an earthen dike in the vicinity of the drainage gully to prevent 

tidal influence during manual removal of solid product along 

the tributary. Chemical product also was recovered along the 

southern bank of Red Lion Creek and in the wetlands (Roy 

F. Weston, Inc., 1988). During removal activities, a geotextile filter fence was installed at the mouth of the wetlands in 

an attempt to prevent any further mobilization of solidified 

spill products to the creek (Roy F. Weston, Inc., 1988). The 

tidally influenced wetlands between the earthen dike and filter 

fence proved particularly challenging for product recovery. 

Dredging operations were conducted in this area, but were 

abandoned after it was determined that no additional removal 

could be performed without damaging the root mass of the 

wetland (Roy F. Weston, Inc., 1988). 

Estimates of contaminant extent have been made based 

on numerous investigations dating from the years immediately 

following the first major spill until the present (Roy F. Weston, 

Inc., 1993a; U.S. Environmental Protection Agency, 1995a; 

Black & Veatch, 2005, 2007; HydroGeoLogic, Inc., 2009; 

Christopher Wolfe, HydroGeoLogic, Inc., written commun., 

2012). Most investigations have focused on the Columbia 

aquifer; however, some sampling has been conducted more 

recently at the SCD in the Potomac aquifer (Black & Veatch, 

2005, 2007; Christopher Wolfe, HydroGeoLogic, Inc., written 

commun., 2012) and the wetland sediments (HydroGeoLogic, 

Inc., 2009). Based on these investigations, contamination was 

documented in the Columbia aquifer as a large plume containing numerous chlorinated benzene compounds plus benzene, 

extending across a large part of the SCD from the central area 

where the spills occurred to the southern boundary of the creek 

(U.S. Environmental Protection Agency, 1995a). The interim 

action for groundwater specified in the 1995 ROD (U.S. 

Environmental Protection Agency, 1995a) of a containment 

barrier wall and recovery and treatment system was reported 

to be operating as designed to prevent continued migration of 

potential dense nonaqueous phase liquids (DNAPLs) (U.S. 

Environmental Protection Agency, 2011). In addition to the 

interim remedy, a final remedy for contaminated groundwater 

and impacted soil and sediment was required in the ROD for 

SCD (U.S. Environmental Protection Agency, 1995a, 2004). 

Biological treatment of SCD soils and sediment was selected 

as the preferred remedy (U.S. Environmental Protection 

Agency, 1995a); however, treatability testing had not demonstrated bioremediation to be a viable method at the time the 

ROD was approved (Roy F. Weston, Inc., 1993b). 

Although spill contaminants were known to impact 

the wetlands bordering Red Lion Creek, sampling and 

evaluation of remedial alternatives was not conducted in 

this part of the SCD until 2008. In 2008, HydroGeoLogic, 

Inc. conducted an investigation of wetland sediments and 

evaluated potential remedial actions compatible with the 

wetlands (HydroGeoLogic, Inc., 2009) in order to fulfill the 
Introduction 7

requirements of the ROD. In what was called the “western 

wetlands” or the area to the west of the unnamed tributary to 

Red Lion Creek, particularly north of the containment dike, 

total concentrations of chlorinated benzenes and benzene 

in the sediment ranged from 800 milligrams per kilogram 

(mg/kg) to 14,000 mg/kg (HydroGeoLogic, Inc., 2009). South 

of the dike, the maximum concentration in sediment was near 

400 mg/kg (HydroGeoLogic, Inc., 2009). In the “eastern wetlands” or the area to the east of the unnamed tributary, concentrations were typically not detected to an order of magnitude 

less than the concentrations in the western wetlands south of 

the containment berm, particularly in the easternmost samples 

(HydroGeoLogic, Inc., 2009). However, one sample (EW-07), 

which is from a site in the USGS study area, had a concentration greater than 2,000 mg/kg (HydroGeoLogic, Inc., 2009). 

Contamination was greater in the surficial sediment samples 

(0–15 centimeters or cm) collected in the “eastern wetlands” 

than in deeper (15–30 cm) samples collected at the same location (HydroGeoLogic, Inc., 2009).

In situ treatment of the contaminants in the wetland is 

advantageous due to the difficulty in operating large machinery and excavating in the soft sediment, sensitive environment. 

Following the inconclusive bioremediation treatability study, 

other treatment alternatives were evaluated for the wetlands, 

including in situ chemical oxidation using laboratory tests and 

a pilot field test (HydroGeoLogic, Inc., 2009). Whereas some 

oxidants tested resulted in mass reductions of chlorinated 

benzenes and benzene in areas where the oxidant was mixed 

with sediment in the field test, secondary water-quality effects 

(such as increases in certain metals) were observed and no 

downgradient effect from the oxidant addition was noted 

(HydroGeoLogic, Inc., 2009). Other in situ alternatives including bioremediation, phytoremediation, and in situ capping 

were suggested as possible components of a remedy, given 

the advances in these technologies over the last 10 years 

(HydroGeoLogic, Inc., 2009). 

Background on Degradation Pathways

Most of the chlorinated benzenes present at the SCD, particularly the higher chlorinated compounds, have a high affinity for the solid phase in soils and sediments and low solubility 

in porewater as shown by their chemical properties (table 1). 

Notably, low aqueous solubilities and relatively high partitioning coefficients for organic carbon (table 1) may indicate that 

aquifer and organic-rich wetland sediments act as a reservoir 

and ongoing source of contamination to groundwater and 

porewater. Despite the persistent nature of the various chlorinated benzenes spilled and detected at the SCD, this class of 

compounds is considered to be a candidate for natural attenuation given the reported aerobic and anaerobic biodegradation 

of compounds from HCB through CB (Spain, 1997). The preferential degradation of higher chlorinated benzenes in anaerobic environments and lesser chlorinated benzenes in aerobic 

environments (Field and Sierra-Alvarez, 2008) indicates that 

wetlands could provide a unique environment for complete 

degradation of chlorinated benzenes. Wetland sediments commonly provide predominantly anaerobic conditions in the subsurface while oxygen delivery to the root zone permits aerobic 

processes to occur in conjunction with anaerobic processes 

(Lorah and others, 2001; Amon and others, 2007; Braeckevelt 

and others, 2007, 2008). This section summarizes the aerobic 

and anaerobic pathways of all the SCD chlorinated benzenes, 

including CB, DCBs,TCBs, TeCBs, and HCB. 

Aerobic degradation has been widely reported (Spain, 

1997) particularly for TeCBs and all lesser chlorinated 

benzenes (Rapp and Gabriel-Jurgens, 2003; Field and SierraAlvarez, 2008). The mechanism of biodegradation occurs 

primarily through oxidation dechlorination by microbial 

deoxygenase enzymes, producing short-lived chlorocatechol 

intermediates and end products of carbon dioxide (CO2), 

water, and hydrochloric acid (Potrawfke and others, 1998; 

Field and Sierra-Alvarez, 2008). Complete mineralization 

(transformation to CO2) of chlorinated benzenes has been 

demonstrated primarily in laboratory studies based on inorganic chloride release, oxygen uptake, and 14CO2

 release from 

labeled contaminants (studies summarized in Field and SierraAlvarez, 2008). Due to the increased solubility of the lesser 

chlorinated benzenes and thus increased availability to microorganisms, aerobic degradation rates are more rapid for these 

compounds compared to anaerobic degradation. Depending on 

the microorganisms that are present, biodegradation may be 

cometabolic, or the direct use of chlorinated benzenes as a sole 

energy source may occur (Field and Sierra-Alvarez, 2008). 

Whereas the majority of wetland sediments are anaerobic at 

depth, aerobic degradation pathways are significant in wetland 

sediments like those found at the SCD, particularly in the root 

zone and during transport of precipitation downward into shallow wetland sediments (Lin, 2003). 

Anaerobic biodegradation through reductive dechlorination has been reported for HCB (Ramanand and others, 1993; 

Nowak and others, 1996) through CB (Fung and others, 2009) 

and to benzene by mixed cultures (fig. 2). Dechlorination 

will proceed through both cometabolic and dehalorespiration 

mechanisms, which dictate preferential pathways of degradation (Field and Sierra-Alvarez, 2008). The microbial strain 

CBDB1, a close relative of the Dehalococcoides ethogenes

strain known to degrade tetrachloroethene and trichloroethene, was the first isolated bacterium that was shown to derive 

energy through the reductive dechlorination of chlorinated 

benzenes (dehalorespiration mechanism) (Adrian and others, 

2000). CBDB1 can dechlorinate TeCB and TCB to various 

DCB isomers (Adrian and others, 2000). After the isolation 

of this first strain, other Dehalococcoides spp. and Chloroflexi

spp. also have been shown to degrade higher chlorinated 

benzenes by dehalorespiration to various DCB isomers but 

not completely to benzene (Adrian and others, 2000; Nelson 

and others, 2011). Anaerobic reductive dechlorination of these 

reduced products, the DCB isomers, will proceed under both 

methanogenic and sulfate-reducing conditions (Field and 

Sierra-Alvarez, 2008; Nelson and others, 2011). Preferential 
8 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

pathways and rates of anaerobic biodegradation vary, with the 

more highly chlorinated compounds degrading most rapidly, 

and particular isomers of TCBs and DCBs degrading more 

rapidly than others on the basis of their structures (fig. 2). The 

anaerobic degradation rate of CB is slowest, and production 

of benzene from CB has been verified only recently (Fung 

and others, 2009), attributable largely to Dehalobacter spp. 

(Nelson and others, 2011). Anaerobic biodegradation of benzene can occur under conditions ranging from iron-reducing 

to methanogenic (Lovley, 2000), but proceeds more rapidly 

under aerobic conditions. In wetlands, an increased abundance 

of organic acids and increased rates of biodegradation of 

chlorinated benzenes have been observed in the presence of 

wetland plants (such as Typha) (Lin, 2003).

Methods of Data Collection and 

Analysis

From 2009–12, an investigation was performed in the 

wetland study area at the SCD to characterize the hydrogeology and geochemistry of the wetland groundwater and to 

assess the evidence for natural and enhanced biodegradation of 

chlorinated benzenes and benzene. A monitoring network consisting of both conventional piezometer and passive diffusionbased devices was used to characterize wetland porewater and 

groundwater in the underlying aquifers. The following sections describe the methods used to characterize groundwater 

hydrogeology and geochemistry in the wetland and to conduct 

in situ microcosm and bench-scale bioreactor testing to assess 

the potential for biodegradation of chlorinated benzenes and 

benzene in the groundwater.

Monitoring Network

The groundwater monitoring network in the study area 

included existing and newly installed drive-point piezometers, 

existing monitoring wells, and several types of passive diffusion samplers, including low density polyethylene (LDPE) and 

dialysis passive diffusion bags (PDBs) and porous membrane 

samplers (peepers) (fig. 3). Data collected from the entire SCD 

monitoring network were used for hydrogeologic analysis (fig. 

3a), whereas the sampling network included only selected sites 

close to the wetland study area (fig. 3b, table 2). The groundwater sampling network included previously installed monitoring wells (designated with the prefix “MW”) in the upland 

area, shallow drive-point piezometers near the upland-wetland 

Table 1. Summary of significant fate and transport properties of contaminants in wetland study area, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware.

[Vp, vapor pressure; S, solubility; Calc. Sscl, calculated super-cooled solubility; NAPL, nonaqueous phase liquid; Csat, saturated concentration; KH, Henry’s Law 

Constant; Kow, octanol-water partioning coefficient; Koc, organic carbon partitioning coefficient; g/mol, grams/mole; mm Hg, millimeters mercury; mg/L, milligrams per 

liter; μg/L, micrograms per liter; atm-m3/mol, atmospheres-meters cubed per mole; L/kg; liter per kilogram; %, percent]

Compound

Molecular 

mass

(g/mol)

Mean

Vp1

(mm 

Hg)

Median

Vp1

(mm 

Hg)

Mean

aqueous 

S1

(mg/L)

Median 

aqueous 

S1

(mg/L)

VanNoort2

Calc. 

Sscl

(mg/L)

Chiou3

Calc. 

Sscl

(mg/L)

Cohen 

and 

Mercer4

effective 

solubility

(mg/L)

NAPL 

interaction

or 

presence 

1% Csat 

(μg/L)

Mean

KH1

(atm-m3/

mol)

Median

KH1

 (atm-m3/

mol)

Mean

Log 

Kow1

Median 

Log 

Kow1

Mean 

log 

Koc1

Median 

log 

Koc1

Benzene 78.11 94.8 94.8 1,770 1,770 17,500 5.57E-03 5.57E-03 2.13 2.13 1.85 1.85

Chlorobenzene 112.6 11.30 11.85 446 472 4,460 3.53E-03 3.65E-03 2.89 2.84 2.24 2.25

1,2-Dichlorobenzene 147 1.31 1.38 132 156 1,560 1.82E-03 1.91E-03 3.41 3.42 3.09 2.89

1,3-Dichlorobenzene 147 2.15 2.15 121 124 1,250 3.20E-03 3.20E-03 3.48 3.51 3.19 2.75

1,4-Dichlorobenzene 147 1.38 1.70 66 76.4 173 137.2 102 1,020 2.28E-03 2.43E-03 3.42 3.44 2.64 2.79

1,2,3-Trichlorobenzene 181.4 0.17 0.17 15 15.1 46 169.3 10.5 105 2.13E-03 2.13E-03 4.05 4.05 4.52 4.20

1,2,4-Trichlorobenzene 181.4 0.39 0.39 131 49 8.7 88 2.39E-03 1.42E-03 4.00 4.02 4.20 4.18

1

 Malcolm and others, 2004; Agency for Toxic Substances and Disease Registry, 2012; Toxnet, 2013, U.S. Environmental Protection Agency, 1996.

2

 VanNoort, 2009. 

3

 Chiou and others, 1986. 

4

 Cohen and Mercer, 1993.
Methods of Data Collection and Analysis 9

boundary installed in 2008–09 prior to this wetland study 

(designated with the prefix “EM”), and 11 piezometer sites 

installed between April and December 2011 as part of the 

wetland study (table 2; fig. 3). 

 The “MW” wells, which are screened in the shallow 

Columbia aquifer (fig. 3a, table 2),were installed in the late 

1990s to delineate contamination north of the industrial part 

of the SCD and continue to be sampled by HydroGeoLogic, 

Inc. as part of the biannual sampling (Christopher Wolfe, 

HydroGeoLogic, Inc., written. commun., 2012). The five 

“EM” drive-point piezometer sites were installed as part of the 

USGS Potomac Aquifer Study to investigate potential groundwater discharge to the wetlands (fig. 3a, table 2). These drivepoint piezometers, mostly screened in the Columbia aquifer, 

were constructed with 5.08-cm-diameter stainless steel casing 

and a 30.5-cm-long slotted screen (0.025 cm). It should be 

noted that polyvinyl chloride (PVC) piezometers (with a 

1.90-cm diameter) also were installed prior to this study at 

sites 7, 8, and 9, but were not used for the wetland study (and 

thus are not shown in table 2); these sites were utilized only 

for passive diffusion samplers and a newly installed piezometer at site 8 (DP-8A) (table 2; figs. 3a, b). 

Most of the new piezometers installed for the wetland 

study were placed along expected groundwater flowpaths 

(transects) toward the southern bank of Red Lion Creek. 

(fig. 3a). The western transect consisted of drive-point 

piezometers (DP) at sites 10, 11, and 13. The eastern transect 

consisted of drive-point piezometer nests at sites 14, 15, 

and 16. Additional new piezometers were installed at sites 

near the upland-wetland boundary (sites 8, 104, 107, 6, 5); 

at an observed seep area in the wetland (site 12); and on the 

northern bank of Red Lion Creek (sites 17 and 18). These 

transect sites generally consist of one or more shallow drivepoint piezometers screened within the wetland sediment and 

one drive point screened just below the wetland sediment in 

the Columbia aquifer. The initial drive points screened in the 

shallow wetland sediment, labeled “S,” were installed in May 

2011 and constructed of 1.9-cm-diameter PVC casing and a 

15.2-cm-long screen made out of high density polyethylene 

filter with a pore size of 60 microns (table 2). Later in 2011, 

the “A” and “AA” drive-point piezometers were installed in 

HCB

PCB

1235TeCB

1245TeCB

1234TeCB

135TCB

124TCB

123TCB

hexachlorobenzene

pentachlorobenzene

1,2,3,5-tetrachlorobenzene

1,2,4,5-tetrachlorobenzene

1,2,3,4-tetrachlorobenzene

1,3,5-trichlorobenzene

1,2,4-trichlorobenzene

1,2,3-trichlorobenzene

1,3-dichlorobenzene

1,4-dichlorobenzene

1,2-dichlorobenzene

chlorobenzene

benzene

carbon dioxide

water

hydrochloric acid

Cl Cl

3

Cl Cl Cl

Cl Cl Cl Cl

Cl Cl Cl

Cl

HCB PCB 1245TeCB

1234TeCB

123TCB 12DCB

14DCB

CB B

CO + H O + HCl 2 2

124TCB

1235TeCB 135TCB 13DCB

Cl Cl Cl Cl

Cl

Cl Cl Cl Cl Cl

Cl Cl Cl

Cl

Cl Cl

Cl Cl

Cl Cl Cl

Cl Cl

Cl Cl Cl

1 3

5

5

5

5

5 2 2 4

4

*

* * *

2

2 2

3 3 2

EXPLANATION

Compounds Pathways reported previously in the following references:

Known site spill compounds from 1981 and 1986. * Can also proceed via aerobic degradation.

1

5

2

4

3

Middledorp and others, 1997; Masunaga and others, 1996

Ramanand and others, 1993; Nowak and others, 1996

Lin, 2003

Nelson and others, 2011 (Dehalobacter species responsible

for dehalorespiration.)

Adrian and others, 2000 (Strain CBDB1 responsible

for dehalorespiration.)

13DCB

14DCB

12DCB

CB

B

CO

H O

HCl

2

2

Figure 2. Diagram showing possible degradation pathways for chlorobenzenes (modified from Field and Sierra-Alvarez, 2008).
10 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

the wetland sediments for water sampling and collection of 

continuous water-level data, respectively. The “A” and “AA” 

drive-point piezometers were constructed of 5.1-cm-diameter 

PVC with a 15.2-cm-long, 0.025-cm slotted screen. The “B” 

drive-point piezometers, most of which were screened in the 

Columbia aquifer, were constructed of 5.1-cm-diameter 304 

stainless steel casing and a 30.5-cm- or 22.9-cm-long slotted 

screen (0.025 cm). Because the wetland sediment thickness 

varied across the site, the bottom of the screens ranged from 

0.55 m to 0.98 m below land surface for the “A,” “AA,” and 

“S” drive points and from 2.3 m to 10.1 m below land surface 

for the B drive points (table 2). 

Additional shallow “A” drive-point piezometers also 

were installed for water-quality and hydrologic sampling 

at sites 8, 104, 107, 6, and 5 near the edge of the wetlandupland boundary. These piezometers were constructed of 

5.1-cm-diameter PVC with a 30.5-cm-long slotted screen 

(0.025 cm). Drive point DP-6AA with a 15.2-cm-long 

screened interval also was installed at site 6 to collect continuous water-level data. 

Several types of passive-diffusion based samplers were 

included in the monitoring network for this investigation. 

Peepers, installed at selected sites in July 2009, September 

2011, and November 2011 (table 3), are passive-diffusion samplers with closely spaced sampling cells to capture changes in 

redox conditions and water quality at a single location across 

a range of depths (Lorah and others, 1997; Spencer and others 

2002). The 122-cm-long peepers used in this study were constructed of thick acrylic plastic with 22 equally spaced rows 

of sample cells, each cell holding approximately 22 milliliters 

(mL) of water. Both sides of the sample cells were overlain 

with 0.2-micrometer (μm) polysulfone filter paper, held in 

75°38'20" 75°38'10" 75°38'00" 75°37'50"

39°36'20"

39°36'10"

A

EXPLANATION

MONITORING SITES AND IDENTIFIERS

A

B

B'

A'

OR-6C

Refer to figure 3b for

enlarged map area

Surface-water gage

Surface-water sampling site

Line of section A-A'

Barrier wall

Columbia Formation

Potomac Formation

Drive-point piezometer nest

MW-24

PW-14

DP-8

A A'

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

9

9

N

River Road River Road

Red Lion Creek

Haul Road

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

Unnamed Unnamed

Tributary

Tributary

0

0

100 200 METERS

300 600 FEET

PZ-6 PZ-5

PZ-7

PZ-8

PZ-28 PZ-30

PZ-9

DP-7

PZ-11 PZ-13

MW-8

EW-41C

EW-61C

PZ-18

PZ-19

MW-9

PZ-10

MW-19 MW-20

PMW-45

MW-24

MW-40

OP-1

MW-25

PZ-12 PZ-15 PMW-47

DP-11, SW-11

DP-10, SW-10

DP-13, SW-13

DP-12, SW-12

DP-104

SW-104

DP-9

DP-8, SW-8

PZ-14

PZ-16

MW-6

EW-51C

PZ-29

MW-7

PMW-46

PW-14

PW-17

MW-21

MW-22

SW-2,3

EM-2

EM-1

EM-4

DP-5, EM-5, SW-5

DP-16, SW-16

EM-6

DP-15, SW-15

DP-6, SW-6

DP-14, SW-14

DP-107, SW-107

DP-18, SW-18

DP-17, SW-17

PZ-17

MW-5 CW-11

MW-23

Surfacewater gage

MW-37

MW-38

MW-36

MW-35

Wetland dividing line for

northwest and northeast

wetland areas

Figure 3. (A) Monitoring network and locations of hydrogeologic sections in the wetland study area, and (B) sampling locations in 

the wetland study area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.
Methods of Data Collection and Analysis 11

place by thin acrylic plastic frames. The peepers were installed 

just below the land surface and allowed to equilibrate with the 

adjacent porewater for a minimum of 2 weeks. 

PDBs, constructed of either LDPE tubing or dialysis 

membrane tubing to obtain samples for different analyses, 

have been shown to be an effective way to characterize wetland groundwater (Vroblesky, 2001; Ehlke and others, 2004; 

Majcher and others, 2007). PDBs installed directly in the wetland sediment were used as a relatively rapid method to obtain 

a large number of groundwater samples across a broad area of 

the wetland, allowing spatial characterization and determination of locations for piezometer installation. The LDPE PDBs 

were constructed of heat-sealed, LDPE plastic that was filled 

with deionized water. The dialysis PDBs were constructed 

from regenerated cellulose tubing with a molecular weight 

cutoff (MWCO) of 8,000 Daltons that was filled with deionized water and closed with cable ties. The dialysis tubing was 

pretreated by the manufacturer to remove any trace amounts of 

hydrogen sulfide and metals and was rinsed thoroughly before 

use in order to remove any remaining preservatives. Before 

installation, the LDPE and dialysis PDBs were purged with 

nitrogen for at least 12 hours to displace any dissolved oxygen 

within the bags. Both PDBs were placed in the saturated wetland sediments, approximately 20–40 cm below land surface, 

in adjacent mesh stainless steel screens attached to PVC rods 

to mark the location. 

Hydrogeologic Data

Hydrogeologic data were collected from across the 

wetland study area in order to gain a clearer understanding 

of groundwater flow and groundwater/surface-water interactions within the wetland area of the SCD. Data collected from 

parts of the SCD outside of the wetland study area as part of a 

broader site monitoring network also were used in parts of this 

75°38'00"

39°36'20"

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

EXPLANATION

SAMPLER TYPE AND IDENTIFIER

Passive Diffusion Bag (PDB) sampler

Drive-point piezometer

Drive-point piezometer and PDB

Drive-point piezometer and peeper

Drive-point piezometer, peeper, and PDB

Drive-point piezometer and PDB (monthly sample)

Drive-point piezometer, peeper, and PDB (monthly sample)

Wetland dividing line for northwest and northeast wetland areas

143

N

Red Lion Creek

Haul Road

0 50

150

100 METERS

0 300 FEET

143

142

141

144

139

133

136

129 128 121 122

123

15 125

126

127

145

146

147

148

150

151

152

153

154

155

156

7

157 158

6

149

160

110 109

108

107

106

120

118

119

117

116 112

113

114

14

104

16

18

17

132 131

138

137

135

12

134

11

10

13

9

8

140

102

B

Figure 3. (A) Monitoring network and locations of hydrogeologic sections in the wetland study area, and (B) sampling locations in 

the wetland study area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.—Continued
12 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12 Table 2. Monitoring network for the wetland study area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.—Continued [NAD83, horizontal coordinate information referenced to North American Datum of 1983; NGVD29, vertical elevation referenced to National Geodetic Vertical Datum of 1929; MP, reference elevation of measuring point referenced to land surface; als, above land surface; bls, below land surface; casing material: PVC, polyvinyl chloride; SS, stainless steel; CS, carbon steel; sampler type: PP, peristaltic pump; PDB, passive diffusion bag; PS, pumped syringe from well tubing; NC, not collected; N/A, not applicable; m, meters; in., inches; ° ’ ”, degrees, minutes, seconds] Site name Site identification number Construction date Latitude (NAD83) Longitude (NAD83) Land surface (m above NGVD29) MP height (m als) Well depth (m bls) Screened depth (m bls) Geologic unit1 Casing material Casing diam- eter (in.) Sampler type Continuous record Drive-point piezometers EM-1S 393610075375902 5/7/2008 39°36’10.35” 75°37’58.07” 1.19 0.78 1.25 0.94-1.25 Wetland sediments SS 2 PP NC EM-1D 393610075375901 5/7/2008 39°36’10.35” 75°37’58.07” 1.14 1.24 3.05 2.85-3.15 Wetland sediments SS 2 PP NC EM-2S 393611075375602 5/7/2008 39°36’11.27” 75°37’56.32” 0.18 0.85 1.15 0.85-1.15 Columbia Formation SS 2 PP NC EM-2D 393611075375601 5/7/2008 39°36’11.27” 75°37’56.32” 0.11 1.08 3.05 2.68-2.99 Columbia Formation SS 2 PP 06/09-05/12 EM-4D 393612075375901 8/25/2008 39°36’11.57” 75°37’59.23” 1.02 0.61 2.86 2.56-2.86 Columbia Formation SS 2 PP NC DP-5A 393614075375802 5/3/2012 39°36’13.50” 75°37’58.91” 0.44 0.75 0.76 0.52-0.67 Columbia aquifer— Wetland sediments interface PVC 2 PDB, PS NC EM-5D 393614075375801 8/25/2008 39°36’13.50” 75°37’58.91” 0.47 0.97 3.10 2.79-3.10 Columbia Formation SS 2 PP, PS, PP NC DP-6A 393615075375805 5/11/2011 39°36’14.62” 75°37’58.15” 0.22 0.73 0.73 0.43-0.73 Wetland sediments PVC 2 PDB, PS NC DP-6AA 393615075375804 12/13/2011 39°36’14.62” 75°37’58.15” 0.23 1.30 0.76 0.61-0.76 Wetland sediments PVC 2 NC 12/11-11/12 EM-6S 393615075375801 8/25/2008 39°36’14.53” 75°37’58.28” 0.35 0.69 1.27 0.97-1.27 Columbia aquifer— Wetland sediments interface SS 2 PDB, PP NC EM-6D 393615075375802 8/25/2008 39°36’14.53” 75°37’58.28” 0.38 0.82 3.34 3.03-3.34 Columbia Formation SS 2 PDB, PP NC DP-8A 393616075380701 5/11/2011 39°36’16.66” 75°38’06.9” -0.03 0.50 0.66 0.37-0.67 Wetland sediments PVC 2 PDB, PS NC DP-10S 393617075380405 6/15/2011 39°36’16.80” 75°38’03.55” 0.35 1.05 0.62 0.46-0.61 Wetland sediments PVC 0.75 PP NC DP-10A 393617075380407 12/13/2011 39°36’16.80” 75°38’03.55” 0.34 1.12 0.70 0.55-0.70 Wetland sediments PVC 2 PS NC DP-10AA 393617075380403 11/16/2011 39°36’16.80” 75°38’03.55” 0.35 1.22 0.73 0.58-0.73 Wetland sediments PVC 2 NC 01/12-10/12 DP-10B 393617075380402 4/27/2011 39°36’16.80” 75°38’03.55” 0.37 1.07 3.32 3.02-3.32 Columbia Formation SS 2 PP 06/11-11/12 DP-11S 393617075380406 6/15/2011 39°36’17.45” 75°38’03.64” -0.17 1.13 0.55 0.40-0.55 Wetland sediments PVC 0.75 PP NC DP-11A 393617075380404 11/16/2011 39°36’17.45” 75°38’03.64” -0.06 1.08 0.75 0.61-0.76 Wetland sediments PVC 2 PDB, PP N/A DP-11B 393617075380401 4/27/2011 39°36’17.45” 75°38’03.64” -0.13 1.01 4.30 3.99-4.30 Columbia Formation SS 2 PDB, PP 06/11-11/12 DP-12S 393618075380204 6/15/2011 39°36’17.55” 75°38’02.37” -0.31 1.04 0.55 0.40-0.55 Wetland sediments PVC 0.75 PP NC DP-12AA 393618075380202 11/16/2011 39°36’17.55” 75°38’02.37” -0.33 1.71 0.79 0.64-0.79 Wetland sediments PVC 2 NC 12/11-11/12 DP-12A 393618075380203 12/13/2011 39°36’17.55” 75°38’02.37” -0.26 1.48 0.98 0.82-0.98 Wetland sediments PVC 2 PP, PS NC DP-12B 393618075380201 4/27/2011 39°36’17.55” 75°38’02.37” -0.30 1.47 3.69 3.38-3.69 Columbia Formation SS 2 PP NC DP-13S 393618075380403 6/15/2011 39°36’18.40” 75°38’03.60” -0.34 1.09 0.59 0.43-0.58 Wetland sediments PVC 0.75 PP NC
Methods of Data Collection and Analysis 13

Table 2. Monitoring network for the wetland study area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.—Continued[NAD83, horizontal coordinate information referenced to North American Datum of 1983; NGVD29, vertical elevation referenced to National Geodetic Vertical Datum of 1929; MP, reference elevation of measuring point referenced to land surface; als, above land surface; bls, below land surface; casing material: PVC, polyvinyl chloride; SS, stainless steel; CS, carbon steel; sampler type: PP, peristaltic pump; PDB, passive diffusion bag; PS, pumped syringe from well tubing; NC, not collected; N/A, not applicable; m, meters; in., inches; ° ’ ”, degrees, minutes, seconds]

Site name

Site 

identification 

number

Construction 

date

Latitude 

(NAD83)

Longitude 

(NAD83)

Land 

surface 

(m above 

NGVD29)

MP 

height 

(m als)

Well 

depth 

(m bls)

Screened 

depth 

(m bls)

Geologic unit1 Casing material

Casing 

diameter 

(in.)

Sampler 

type

Continuous 

record

DP-13A 393618075380402 11/16/2011 39°36’18.40” 75°38’03.60” -0.32 1.07 0.77 0.61-0.76 Wetland sediments PVC 2 NC NCDP-13B 393618075380401 5/11/2011 39°36’18.40” 75°38’03.60” -0.31 1.19 5.67 5.36-5.67 Columbia Formation SS 2 PP, PDB 06/11-02/13

DP-14S 393616075380103 6/15/2011 39°36’16.30” 75°38’00.89” 0.37 1.07 0.58 0.43-0.58 Wetland sediments PVC 0.75 PP, PDB NC

DP-14A 393616075380102 11/16/2011 39°36’16.30” 75°38’00.89” 0.37 1.07 0.74 0.58-0.73 Wetland sediments PVC 2 PDB NCDP-14B 393616075380101 6/15/2011 39°36’16.30” 75°38’00.89” 0.43 1.16 3.96 3.66-3.96 Columbia Formation SS 2 NC 11/11-11/12

DP-15S 393617075380004 6/15/2011 39°36’17.12” 75°38’00.08” 0.02 1.09 0.59 0.43-0.58 Wetland sediments PVC 0.75 PP NC

DP-15A 393617075380003 12/13/2011 39°36’17.12” 75°38’00.08” 0.06 1.00 0.80 0.64-0.79 Wetland sediments PVC 2 PDB, PS NC

DP-15AA 393617075380002 11/16/2011 39°36’17.12” 75°38’00.08” 0.07 1.14 0.85 0.70-0.85 Wetland sediments PVC 2 NC 12/11-01/13

DP-15B 393617075380001 6/15/2011 39°36’17.12” 75°38’00.08” 0.04 1.07 4.15 3.84-4.15 Columbia Formation SS 2 PP, PDB 6/11-02/13

DP-16S 393618075375903 6/15/2011 39°36’17.96” 75°37’58.95” -0.37 1.09 0.60 0.43-0.58 Wetland sediments PVC 0.75 PP NC

DP-16AA 393618075375902 11/16/2011 39°36’17.96” 75°37’58.95” -0.32 1.80 0.85 0.70-0.85 Wetland sediments PVC 2 NC 11/11-11/13

DP-16A 393618075375904 1/18/2012 39°36’17.96” 75°37’58.95” -0.36 1.46 0.98 0.82-0.98 Wetland sediments PVC 2 PDB NCDP-16B 393618075375901 6/15/2011 39°36’17.96” 75°37’58.95” -0.31 1.53 10.15 9.85-10.15 Wetland sediments SS 2 PP, PDB 12/11-02/13DP-17A 393620075375502 11/16/2011 39°36’20.87” 75°37’53.38” -0.31 1.12 0.70 0.55-0.70 Wetland sediments PVC 2 PDB, PS NCDP-17B 393620075375501 11/15/2011 39°36’20.87” 75°37’53.38” -0.26 1.43 5.36 5.06-5.36 Columbia Formation SS 2 PDB, PP 12/11-02/13

DP-18AA 393621075375503 5/24/2012 39°36’21.47” 75°37’52.75” 0.23 0.74 0.83 0.59-0.75 Wetland sediments PVC 2 PS NC

DP-18B 393621075375501 12/13/2011 39°36’21.47” 75°37’52.75” 0.23 0.52 2.29 2.06-2.29 Columbia Formation SS 2 PDB NC

DP-104A 393616075380201 5/11/2011 39°36’16.42” 75°38’02.44” 0.39 0.40 0.75 0.46-0.76 Wetland sediments PVC 2 PDB, PS NCDP-107A 393615075380001 5/11/2011 39°36’15.96” 75°38’00.74” 0.42 0.45 0.59 0.27-0.58 Wetland sediments PVC 2 PDB, PS NC

Monitoring wells

MW-21 393616075380502 11/10/1995 39°36’15.86” 75°38’04.58” 2.26 0.67 6.10 3.05-6.10 Columbia Formation CS 2 PDB NCMW-22 393616075380202 11/10/1995 39°36’15.65” 75°38’02.92” 2.04 0.58 5.03 1.98-5.03 Columbia Formation CS 2 PDB NCMW-23 393615075375901 12/27/1995 39°36’15.08” 75°37’59.73” 2.09 0.73 7.92 4.88-7.92 Columbia Formation CS 2 PDB NCMW-24 393614075380901 10/7/1999 39°36’14.91” 75°38’09.76” 4.36 0.83 11.43 8.38-11.43 Columbia Formation CS 4 NC NCMW-25 393616075380601 10/5/1999 39°36’15.86” 75°38’07.02” 3.39 0.70 9.51 6.46-9.51 Columbia Formation CS 4 PDB NCMW-36 393624075375401 10/1/1999 39°36’24.35” 75°37’54.42” 4.62 0.96 16.46 13.41-16.46 Columbia Formation CS 4 NC NCMW-37 393622075375203 5/25/2001 39°36’21.89” 75°37’52.19” 1.25 0.59 15.24 9.14-15.24 Columbia Formation CS 4 NC NCMW-38 393620075374801 5/29/2001 39°36’20.53” 75°37’49.03” 1.62 0.57 15.24 9.14-15.24 Columbia Formation CS 4 NC NC
14 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12 Table 2. Monitoring network for the wetland study area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.—Continued [NAD83, horizontal coordinate information referenced to North American Datum of 1983; NGVD29, vertical elevation referenced to National Geodetic Vertical Datum of 1929; MP, reference elevation of measuring point referenced to land surface; als, above land surface; bls, below land surface; casing material: PVC, polyvinyl chloride; SS, stainless steel; CS, carbon steel; sampler type: PP, peristaltic pump; PDB, passive diffusion bag; PS, pumped syringe from well tubing; NC, not collected; N/A, not applicable; m, meters; in., inches; ° ’ ”, degrees, minutes, seconds] Site name Site identification number Construction date Latitude (NAD83) Longitude (NAD83) Land surface (m above NGVD29) MP height (m als) Well depth (m bls) Screened depth (m bls) Geologic unit1 Casing material Casing diam- eter (in.) Sampler type Continuous record Surface-water sampling sites SW-02 393611075375603 5/7/2008 39°36’21.47” 75°37’56.90” -0.50 1.42 N/A N/A N/A N/A N/A NC 06/09-08/10 SW-06AA 393615075375806 12/13/2011 39°36’14.62” 75°37’58.15” 0.23 1.30 N/A N/A N/A N/A N/A NC NC SW-08A 393617075380702 5/11/2011 39°36’16.66” 75°38’06.9” -0.03 0.50 N/A N/A N/A N/A N/A NC NC SW-104A 393616075380203 5/11/2011 39°36’16.42” 75°38’02.44” 0.39 0.40 N/A N/A N/A N/A N/A NC NC SW-107A 393616075380105 5/11/2011 39°36’15.96” 75°38’00.74” 0.42 0.45 N/A N/A N/A N/A N/A NC NC SW-10B 393617075380408 4/27/2011 39°36’16.8” 75°38’03.55” 0.37 1.07 N/A N/A N/A N/A N/A NC NC SW-11B 393617075380409 4/27/2011 39°36’17.45” 75°38’03.64” -0.13 1.01 N/A N/A N/A N/A N/A NC NC SW-12B 393618075380205 4/27/2011 39°36’17.55” 75°38’02.37” -0.30 1.47 N/A N/A N/A N/A N/A PP NC SW-13B 393618075380404 5/11/2011 39°36’18.40” 75°38’03.60” -0.31 1.19 N/A N/A N/A N/A N/A PP NC SW-14B 393616075380104 6/15/2011 39°36’16.30” 75°38’00.89” 0.43 1.16 N/A N/A N/A N/A N/A NC NC SW-15B 393617075380006 6/15/2011 39°36’17.12” 75°38’00.08” 0.04 1.07 N/A N/A N/A N/A N/A NC NC SW-16B 393618075375905 6/15/2011 39°36’17.96” 75°37’58.95” -0.31 1.53 N/A N/A N/A N/A N/A PP NC SW-17B 393620075375503 11/15/2011 39°36’20.87” 75°37’53.38” -0.26 1.43 N/A N/A N/A N/A N/A PP NC SW-18B 393621075375403 12/13/2011 39°36’21.47” 75°37’52.75” 0.23 0.52 N/A N/A N/A N/A N/A NC NC

1 Wetland sediments: most commonly described as Marsh Deposits (Qm) that are structureless to finely laminated, black to dark gray, organic-rich silty clay with some peat beds.

 Columbia Formation (Qcl): Yellow to reddish-brown, fine to coarse sand, typically cross-bedded, with some gravel at base of unit. Scattered beds of clayey silt common, often overlying medium to coarse sand. From Ramsey, 2005.
Methods of Data Collection and Analysis 15

Table 3. Groundwater sampling and site events that may have influenced sampling results in the wetland study area, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware.

[LDPE, low density polyethylene; dialysis, dialysis membrane; PDBs, passive-diffusion bags; DCRC, Delaware City Refining Company]

Sample type Sample locations or sites Sampling dates 

Barrier wall installed in 2007

Peepers 6, 7, 8, 9 July 2009 

In Situ microcosms 6, 9 October–December 2009 

DCRC limited pumping starting December 2009 and stopped in May 2010

LDPE PDBs (in sediment) 8, 102, 103, 104, 106, 107, 108, 109, 110, 7, 6, 14, 

15, 112, 113, 114, 116, 117, 118, 119, 120, 131, 

121, 122, 123, 125, 126, 127

September 2010

LDPE and dialysis PDBs (in sediment) 8, 102, 103, 104, 107, 110, 6, 114, 14, 119, 121, 15, 

113, 117, 126, 130

October 2010 

LDPE and dialysis PDBs 8, 104, 107, 112, 7, 6, 116, 123, 125; 128, 129, 12, 

131, 132, 133, 134, 135, 136, 11, 138, 139, 140, 

144, 141, 142, 143

April 2011

LDPE and dialysis PDBs (in drive-point piezometers) DP-6A, DP-8A, DP-104A, DP-107A Monthly beginning July 

2011–January 2012

Hurricane Irene, August 28, 2011, and damage to tide control structure

LDPE and dialysis PDBs (in sediment and drive-point 

piezometers)

Existing PDB sites + DP-10B, DP-11B, DP-13B, 

DP-14B, DP-15B, MW-21, MW-22, MW-23, 

MW-25, EM-6D, EM-6S

October 2011

Peepers 10, 12, 14, 15 October 2011

LDPE and dialysis PDBs (in sediment and drive-point 

piezometers)

145*–160*, 6, 8, 17, DP-6A, DP-8A, DP-104A, 

DP-107A, DP-17B, EM-6D, EM-6S

December 2011

Peepers 6, 8 December 2011

DCRC pumping resumes January 2011

LDPE and dialysis PDBs/purge and recover DP-6A, DP-8A, DP-104A, DP-107A Monthly beginning February 

2012–May 2012

LDPE and dialysis PDBs (in sediment for comparison 

to PDBs placed in drive-point piezometers)

DP-6A, DP-8A, DP-104A, DP-107A March 2012

Purge and recover for A piezometers, peristaltic pump 

for B piezometers

DP-6A, DP-8A, DP-104A, DP-107A, DP-5A, 

DP-17A, DP-18AA, DP-10B, DP-15B, SW-13, 

SW-17

May 2012

Purge and recover for A piezometers, peristaltic pump 

for B piezometers

DP-6A, DP-8A, DP-104A, DP-107A, DP-5A, 

DP-17A, DP-18AA, DP-10B, DP-15B, SW-13, 

SW-17, DP-10A, DP-15A, DP-12A, DP-12B

July 2012–August 2012 

*

 Only LDPE PDBs installed.
16 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

analysis. Data collected for hydrologic interpretation included 

water-level measurements for determination of groundwater 

hydraulic head, flow direction and rates, and slug tests to 

approximate hydraulic conductivity. Measured gradients and 

hydraulic conductivities were used to calculate specific discharge and average linear velocity of groundwater.

Water-Level Measurements

Water-level measurements were made in the wetland 

study area and across the SCD in order to determine groundwater-flow directions and rates and to monitor for changes in 

the flow system that could result from seasonal fluctuations, 

pumping in the Potomac regional water-supply aquifer, and 

tidal influence following damage sustained by the tide control 

structure on Red Lion Creek. Water-level measurements were 

used to construct groundwater head maps, hydrographs, and 

cross-sectional flow diagrams. Vertical and lateral flows within 

the Columbia aquifer and surficial wetland sediments were 

examined individually, and vertical flow between the two units 

was analyzed.

Water levels in the drive-point piezometers and monitoring wells were measured during monthly synoptic assessments 

and during sampling events (table 3) to monitor any seasonal 

variations within the aquifer and the porewater in the wetland 

sediments. Monthly synoptic measurements were coordinated 

with HydroGeoLogic, Inc. to obtain wetland and upland 

measurements in the SCD area; USGS and HydroGeoLogic, 

Inc. staff measured water levels in wells outside of the wetland 

area, whereas USGS measured water levels in the piezometers 

within the wetland area. The water levels were measured from 

the measuring point with electronic tapes graduated to 

0.30 cm. The altitude of the measuring points is given relative 

to the National Geodetic Vertical Datum of 1929 (NGVD 29) 

and was determined using a digital level to an accuracy of 

0.61cm for all sites except DP-18AA. Water levels in surface 

water, denoted as “SW,” were measured at a cluster from 

the “B” drive point measuring point at all sites except site 6, 

which referenced DP-6AA. All water-level data were entered 

and stored in the USGS National Water Information System 

(NWIS) database. Once in the database, the water levels were 

normalized to altitudes, allowing them to be compared across 

the study area. 

Continuous water-level data were collected to provide 

information about shallow groundwater response to seasonal 

changes, tidal effects, groundwater withdrawals, and groundwater recharge events. Pressure transducers were installed at 

the Red Lion Creek tide gage and in 12 drive-point piezometers, consisting of 5 shallow “AA” piezometers and 7 deeper 

“B” piezometers (table 2), to provide coverage along both 

transects and at the wetland boundary. The pressure transducers were installed in the piezometers between June 2011 and 

December 2011 with a period of record ranging from 12 to 

19 months. The range of the pressure transducers was 5 to 15 

pounds per square inch gage (psig), and the water levels interpreted from readings of pressure were accurate to 0.1 percent. 

The pressure transducers recorded the water-level altitudes 

every 15 minutes, and the data were stored in NWIS. The 

calibration of the transducers was checked during the monthly 

water-level synoptic assessments and any drift or datum corrections were applied in NWIS. 

Slug Tests

Slug tests were conducted at four drive-point piezometer 

nests in the wetland study area in order to approximate 

hydraulic conductivity of the hydrologic units. Tests were 

performed on piezometers screened in wetland sediments 

(DP-6A, DP-11A, DP-12AA), and in the Columbia aquifer 

(EM-6D, DP-11B, DP-12B, DP-15B), representing both the 

east and west sides of the wetland study area during two 

different time periods. Tests were conducted at the DP-11, 

DP-12, and DP-15 drive-point piezometer nests during the 

winter 2011–12 field season, whereas tests at EM-6D and 

DP-6A were conducted during the winter of 2013. A test 

was initiated at DP-15A but was terminated prior to completion because of an extremely slow response to the artificially 

induced change in hydraulic head. 

Slug tests were performed by inducing an artificial 

change in hydraulic head and measuring the time taken for 

a water level to return to its static condition. A Troll 700 (In 

Situ, Inc.) pressure transducer was placed in the piezometer 

following an initial static water-level measurement, and was 

used to collect water-level data at a logarithmic time scale 

during the test. Induced change in hydraulic head was accomplished in two ways depending on the formation in which 

the drive-point piezometer was screened. For drive-point 

piezometers screened in shallow wetland sediments, bail 

tests and slug tests were accomplished using deionized water 

(Freeze and Cherry, 1979, p. 339). Drive-point piezometers 

were bailed for rising head tests, and a slug of deionized water 

was injected for falling head tests. These tests were done 

in sequence at each drive-point piezometer. At DP-11A and 

DP-12A, a falling head test was conducted first by adding 1 

liter (L) of deionized water to the drive-point piezometer and 

recording the return to static water level. Three subsequent 

rising head tests were completed by removing approximately 

340 mL of water from the drive-point piezometer using a PVC 

bailer, which was 4.1 cm in diameter and 30 cm in length. For 

DP-6A, the bailer (approximately 340 mL) was used in place 

of the initial 1-L slug of deionized water. First, the bailer was 

inserted into the drive point and the water level was allowed 

to equilibrate; then it was removed, bailing approximately 340 

mL of native water from the drive-point piezometer, and the 

return to static conditions was monitored. Once static water 

level was regained, the full bailer was replaced, and the return 

to static conditions was again monitored. Two sequences of 

these tests were conducted at DP-6A. At drive-point piezometers screened in the Columbia aquifer, at least three sequences 

of rising and falling head slug tests were conducted using a 

PVC cylindrical slug, which was 2.86 cm in diameter and 
Methods of Data Collection and Analysis 17

152 cm in length. All data were collected using WinSitu software (In Situ, Inc.).

Slug test data were analyzed and hydraulic conductivities 

were computed using the Bouwer-Rice method (Fetter 2001, 

p. 197–200). These estimates of hydraulic conductivity were 

then used to determine approximate groundwater discharge 

rates for given drive-point piezometers. 

Calculations of Hydraulic Gradient, Specific 

Discharge, and Linear Velocity

Groundwater altitudes collected within the wetland study 

area were used to determine hydraulic gradients (flow directions). Hydraulic conductivities derived by slug test analysis 

for selected points were used in conjunction with these 

gradients to compute specific discharge (Darcian velocity) and 

average linear velocity in the wetland and Columbia aquifer in 

the wetland study area. Hydraulic gradient, specific discharge 

(v), and average linear velocity (V) were calculated in both the 

lateral and vertical directions.

Vertical gradients were calculated for drive-point 

piezometer pairs at sites EM-1, EM-2, EM-5, EM-6, DP-10, 

DP-11, DP-12, DP-13, DP-14, DP-15, and DP-16. At each of 

these pairs, the vertical gradient was calculated between wells 

screened in the Columbia aquifer and wetland sediments, 

except for EM-1, where both wells were screened in wetland 

sediments, and EM-2 and EM-5, where both wells were 

screened in the Columbia aquifer or at the interface of the 

Columbia aquifer and wetland sediments, respectively. Lateral 

hydraulic gradients were calculated by estimating the change 

in hydraulic head (estimated from mapped potentiometric 

surface contours) and dividing this difference by the difference 

between these contours along transects where drive-point 

piezometers were installed. 

Lateral specific discharge (Darcian velocity) was calculated using the following equation (Fetter, 2001, p. 82):

v = -K (dh / dl) (1)

where

K = hydraulic conductivity;

dh = the change in hydraulic head between 

points; and

dl = the lateral distance between points.

Estimates of vertical groundwater discharge were calculated across the interface between wetland sediments and the 

Columbia aquifer, which have differing hydraulic conductivities. In order to represent the thickness of each unit through 

which groundwater discharge is estimated, a weighted hydraulic conductivity was used. Thus, the equation for specific 

discharge was modified: 

- ( / ) w v = K dh dz (2)

where

Kw = weighted hydraulic conductivity;

dh = the change in hydraulic head between 

points; and

dz = the vertical distance between screen 

midpoints.

Weighted hydraulic conductivity was calculated using the 

equation (Freeze and Cherry, 1979, p. 34):

Kw = (d1 + d2 ) /((d1 / K1 ) + (d2 / K2 )) (3)

where

Kw = weighted hydraulic conductivity;

d1 = the distance between the screen midpoint 

of the piezometer screened in wetland 

sediments and the interface between the 

wetland sediments and the Columbia 

aquifer at that piezometer;

d2 = the distance between the midpoint of the 

screen in the Columbia aquifer and the 

wetland sediment interface;

K1 = the hydraulic conductivity of wetland 

sediments at the given piezometer; and

K2 = the hydraulic conductivity of the Columbia 

aquifer at the given piezometer. 

Average linear groundwater velocity was calculated using 

the following equation (Fetter 2001, p. 125):

(- / ) V K e = dh dl ∗n for lateral velocity (4)

and a modified version of Fetter’s average linear groundwater 

velocity for the vertical direction:

(- / ) V Kw e = dh dl ∗n for vertical velocity (5)

where

K = hydraulic conductivity;

Kw = weighted hydraulic conductivity;

dh = the change in hydraulic head between 

points;

dl = the lateral distance between points; and

ne = the effective porosity of the unit through 

which groundwater is traveling 

(dimensionless).

Estimates of effective porosity of the wetland sediments 

and the underlying aquifer sediments at the SCD were drawn 

from previous studies conducted in similar wetland environments. For wetland sediments, an effective porosity of 0.4 

was used based on computations conducted for groundwater 

velocity in a similar environment (Lorah and others, 1997; 

Mitsch and Gosselink, 1993). Soil cores were collected at 

four drive-point piezometer nest locations (EM/DP-6, DP-8, 
18 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

DP-104, DP-107) within the SCD wetland study area and 

analyzed for carbon content resulting in a range of 0.19 to 

24.2 percent carbon. For groundwater velocity estimates for 

the Columbia aquifer, an effective porosity of 0.3 was used 

(Domenico and Schwartz, 1990). This effective porosity also 

was used for calculations of vertical velocity, which were 

made across the wetland sediments interface and included both 

Columbia sands and wetland sediments consisting of clay, silt, 

and organic matter. 

Groundwater and Surface-Water Geochemical 

Data

Groundwater and surface-water samples collected 

between July 2009 and August 2012 are summarized in table 

3. Sample collection began with only four sites in July 2009 

and then expanded spatially and temporally to characterize 

the wetland geochemistry. Sample collection and methods of 

analysis and data quality are discussed below.

Sample Collection

An initial assessment of the wetland groundwater chemistry began in 2009 with a detailed vertical characterization 

of the shallow wetland sediments using peepers at four sites 

(table 3). Large scale synoptic and recurring monthly sampling 

events then were used to determine the spatial extent of 

contamination and capture any seasonal variation in the geochemistry of the wetland and aquifer groundwater (table 3). 

Because of the ease of installation and sampling, peepers and 

PDBs were initially installed in the sediments, allowing more 

sampling sites in the wetland area with minimal infrastructure 

before installation of the piezometers. LDPE PDBs were used 

to obtain samples for VOCs and dissolved gas analyses, and 

dialysis membrane PDBs were used to obtain samples for all 

other constituents and field parameters. Usage of either one 

or both of these diffusion samplers are referred to simply as 

PDBs for the remainder of this report, although table 3 notes 

specific sampler types that were used for each sampling event. 

PDBs were used to obtain samples by direct push into the sediment and also to obtain downhole samples from drive-point 

piezometers and wells. For downhole sampling, PDBs were 

placed inside a polyethylene mesh tubing with stainless steel 

weights and suspended within or slightly above the screened 

interval with braided nylon cord. All PDBs were allowed to 

equilibrate for a minimum 2-week period before removal for 

sample collection. 

In July 2009, peepers at sites 6, 7, 8, and 9 were sampled 

for VOCs, ferrous iron, chloride, nitrate, and ammonia. In 

September 2010, PDBs for the collection of VOC and methane 

samples were installed in the ground along three parallel 

transects: one extending from site 8 to site 6, along the edge 

of the wetlands and two shorter transects in the eastern part of 

the wetland (table 3). The synoptic sampling event in October 

2010 was a resampling of selected sites using in-ground PDBs 

and included samples for VOCs, methane, sulfate, chloride, 

pH, and specific conductance (table 3). Multiple sets of PDBs 

also were installed at some sites (6, 8, 15, 121, 104, and 114) 

to assess spatial geochemical variability and assist in selecting 

sites for in situ microcosm experiments. During the April 2011 

synoptic sampling event, PDBs were installed at existing sites 

and at new sites that extended the two northern transects to 

the west (table 3). In July 2011, PDBs were used to sample 

newly installed piezometers DP-10B, 11B, 12B, 13B, and 

15B by suspending them within or slightly above the screened 

interval. 

The largest sampling event began in October 2011 and 

continued through December 2011, collectively referred to 

as the October–December 2011 sampling event in this report. 

PDBs were installed in the sediment at most pre-established 

sites and at new sites; peepers were installed at six sites; 

and selected drive-point piezometers and monitoring wells 

were sampled by suspending PDBs within or slightly above 

the screened interval (table 3). All PDBs were sampled for 

VOCs, dissolved gases, ferrous iron, total iron, sulfide, anions, 

cations, ammonia, pH, and specific conductance, however 

only VOCs and dissolved gases were sampled from the PDBs 

installed in the sediment at new sites 145 to 158 and site 160 

(fig. 3b). Peepers were sampled for VOCs, dissolved gases, 

ferrous iron, sulfide, anions, cations, and ammonia. 

 Monthly samples collected from July 2011 through 

August 2012 are included in this report (table 3). Starting in 

July 2011, monthly sampling occurred at sites 6, 8, 104, and 

107 in the “A” depth piezometers to determine if groundwater 

and porewater geochemistry varied seasonally. At each site 

and event, samples were typically collected for analysis of 

VOCs, dissolved gases, ferrous iron, total iron, sulfide, anions, 

cations, ammonia, pH, and specific conductance. From July 

2011 to January 2012, monthly sampling was conducted by 

suspending PDBs downhole. Beginning in February 2012, the 

drive-point piezometers screened within the wetland sediments were sampled by the purge and recover method, which 

involved purging with a dedicated bailer and then sampling 

the next day with dedicated Teflon tubing, a three-way stopcock, and a syringe. Sampling was conducted after ensuring 

that the water level had recovered at least 75 percent. In May 

2012, the monthly sampling was expanded to include DP-5A, 

DP-17A, DP-18AA, DP-10B, DP-15B, SW-13, and SW-17 

and expanded again in July 2012 to include DP-10A, DP-15A, 

DP-12A, and DP-12B. 

Nearly all piezometers were fully recovered within 24 

hours except DP-104A, which sometimes required 2 days 

to recover. The tubing was placed near the bottom of the 

piezometer while sampling but within the screened interval 

to decrease the turbidity of the purge water and prevent filter 

loading. Throughout sample collection, a three-way stopcock 

was used to maintain positive pressure within the tubing in 

between syringe volumes. All filtered samples were collected 

with a rinsed Pall Acrodisc® 0.2-μm Supor® membrane syringe 

filter and a 10-mL syringe. 
Methods of Data Collection and Analysis 19

A change in sample collection method (downhole PDBs 

to purge and recover sampling; see table 3) was implemented 

in the shallow wetland piezometers due to results of direct 

comparisons of in-well PDBs, sediment PDBs, and purge and 

recover samples. These comparisons indicated that in-well 

PDBs may generally underestimate VOCs and, to a lesser 

extent, methane, sulfide and ferrous iron concentrations in 

porewater compared to both sediment PDBs and purge and 

recover samples. Particularly at elevated concentrations 

(greater than 10 milligrams per liter [mg/L] of VOCs), the 

disparity between in-well and sediment PDBs was apparent. It 

is possible groundwater passively entering screened intervals 

in shallow drive-point piezometers was dominated more by 

deeper porewater (where results generally showed lower 

VOC concentrations, as discussed later in the Distribution of 

Volatile Organic Compounds section) or were less directly 

impacted by contaminant sources in nearby sediment. 

Nonetheless, direct comparisons of the purge and recover 

sampling method to the PDB methods resulted in an order of 

magnitude concentration difference between the sediment and 

in-well PDBs at some sites, although good agreement was 

observed at others. To eliminate this potential bias, further 

samples were collected using the purge and recover method. 

Drive-point piezometers screened in the Columbia 

aquifer were sampled with a peristaltic pump fitted with 

dedicated Teflon and CFlex© tubing, a three-way stopcock, and 

a syringe. The piezometers were purged and sampled using 

the methods described by the USGS National Field Manual 

for the Collection of Water-Quality Data (U.S. Geological 

Survey, variously dated). The pump intake was placed within 

the screened interval and the piezometer was purged of a 

minimum of three casing volumes at a pump rate that allowed 

the water level to remain stable. Field parameters, including 

pH, water temperature, specific conductance, and sometimes 

turbidity were recorded every 5 minutes. Once the field parameters were stable for 25 minutes, the piezometer was sampled 

with a syringe.

The surface-water samples were collected using a peristaltic pump fitted with dedicated Teflon and C-Flex® tubing, 

a three-way stopcock, and a syringe. The intake tubing was 

placed within the center one-third of the water column and 

held in position with a permanently installed PVC pipe. Use of 

the pipe ensured the samples were consistently collected at the 

same location and prevented suspension of bottom sediments. 

The sample was collected after the tubing had been rinsed with 

native water a minimum of 10 tubing volumes and the field 

parameters were stable. 

Chemical Analysis

Water temperature, specific conductance, and pH were 

measured with a Hach HQ 40d multi-parameter meter with a 

standard conductivity probe and gel-filled pH probe. Turbidity 

was determined using a Hach 2100Q turbidity meter calibrated 

with Hach StablCal® standards. All meters were calibrated 

daily as described in the USGS National Field Manual for the 

Collection of Water-Quality Data (U.S. Geological Survey, 

variously dated).

Methane, ethane, and ethene samples were collected 

by injecting 2 to 5 mL of raw sample water into a weighed, 

sealed, nitrogen purged, serum bottle filled with at least 

.035 grams (g) of trisodium phosphate (TSP), a preservative 

(Majcher and others, 2007). The samples were stored on ice 

or in a refrigerator until the headspace was analyzed on a 

HP 7890 gas chromatograph with a flame ionization detector 

(GC-FID) located at the USGS Maryland-Delaware-DC Water 

Science Center (MD-DE-DC WSC) (Spencer and others, 

2002). 

VOC analyses were conducted at the USGS MD-DE-DC 

WSC Research lab on a purge and trap gas chromatograph 

with a mass spectrometer (GC-MS) and a headspace autosampler. A modified purge and trap method that combines components of EPA methods 524.2 (U.S. Environmental Protection 

Agency, 1995) and 8260C (U.S. Environmental Protection 

Agency, 2006) was implemented on the GC-MS with 

modifications outlined in Majcher and others (2007) and the 

addition of selected chlorinated benzenes to the analyte list. 

The GC-MS is equipped with an OI Analytical 4560 sample 

concentrator and an OI Analytical 4552 headspace autosampler. The modified method enabled low volume samples to be 

analyzed. VOC samples were collected in the field in 8-mL 

borosilicate glass vials with Teflon septa with no headspace 

and stored on ice or in a refrigerator until processed. 

Total and ferrous iron concentrations were determined 

using the colorimetric bipyridine method (Baedecker and 

Cozzarelli, 1992) after filtering the samples using a 2-μm 

membrane syringe filter and diluting them by a factor of 10 

with deionized water. The appropriate reagents were added at 

sample collection to induce the color change and fixed in the 

field with sodium acetate. The samples were stored chilled 

until analyzed on a Hach DR 2800TM spectrophotometer reading at a wave length of 520 nanometers. 

Ammonia, nitrate, and sulfate samples were analyzed 

in the field or mobile lab using Hach TNTplusTM vials and 

analyzed immediately after collection on a Hach DR 2800TM

spectrophotometer (Hach Company, 2007). Sulfide samples 

were collected and analyzed immediately in the field or mobile 

lab using the methylene blue method (Hach Company, 2007) 

and also analyzed on the spectrophotometer. To ensure that the 

sample concentrations fell within the range determined by the 

manufacturer, samples were diluted with deionized water, or 

multiple test kits covering different concentration ranges were 

used if possible. Samples were filtered with a 0.2-μm membrane syringe filter if the sample was visibly turbid, except 

for sulfide samples. Sulfide samples were not filtered because 

of sulfide’s high potential for volatilization and oxidation. 

For each lot of test kit vials or reagents, the calibration of the 

spectrophotometer was confirmed with standards purchased 

from Hach Company. 

Samples for analysis of nonvolatile dissolved organic 

carbon (NVDOC) were filtered with a 0.2-μm membrane 

syringe filter into 4-mL baked amber vials with Teflon septa. 
20 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

The samples were acidified to a pH below 2 with high purity 

hydrochloric acid in the field. The samples were stored chilled 

until analyzed at the USGS Research Laboratory, Reston, 

Virginia. The NVDOC samples were measured using a 

Shimadzu TOC-Vcsn Total Organic Analyzer and the 

680 degrees Celsius (°C)-combustion catalytic oxidation 

method modified from EPA Method 415.3 (U.S. 

Environmental Protection Agency, 2009). 

Before August 2011, chloride concentrations were determined by the mercuric nitrate method using the Hach Digital 

Titrator (Hach Company, 2006) at the USGS MD-DE-DC 

WSC laboratory. The samples were collected raw, stored in 

40-mL amber glass vials with Teflon septa, and chilled until 

processed. After August 2011, anion samples, including chloride, were filtered using a 0.2-μm membrane syringe filter and 

stored in a rinsed 15-mL high density polyethylene bottle. The 

samples were stored chilled and then analyzed at the USGS 

Reston laboratory using a Dionex 120 Ion Chromatograph 

with an ED50 Electrochemical detector. The AS14 anionexchange column was used for the separation of the anions 

chloride, bromide, nitrate, phosphate, and sulfate, meeting the 

performance requirements specified in EPA Method 300.0 (A) 

(U.S. Environmental Protection Agency, 1993). 

Data Quality

The quality of the groundwater data collected was determined based on quality-control data obtained from blanks 

and duplicate samples. Inadvertent cross-contamination of 

water-quality samples was minimized by using either new or 

dedicated tubing, syringes, and other sampling supplies for all 

piezometer samples that were collected by the syringe or peristaltic pump sampling methods, and for withdrawal of samples 

from PDBs and peepers. Equipment blanks were collected 

for about 10 percent of the total number of samples during 

a sampling event by running deionized water through clean 

tubing and syringes and collecting it as a sample to evaluate 

potential contamination from equipment. For each sampling 

event, a set of blank PDBs was sampled after preparing and 

purging with nitrogen gas and before deployment in the field; 

the deionized water used to fill the peepers also was analyzed 

as an equipment blank. In addition, trip blanks for VOC analysis were collected during each sampling event by filling vials 

with deionized water and keeping them with the groundwater 

samples collected at the site to evaluate the potential for crosscontamination during transport and storage until analyses were 

performed. Duplicate samples were collected by sequential or 

concurrent filling of vials during piezometer sampling with 

the syringe or peristaltic pump. Duplicate samples were not 

collected from peepers or PDBs due to the limited amount of 

sample available. 

An analysis of a total of 24 blanks collected during this 

study indicated that concentrations of chlorinated benzenes 

and benzene were generally less than 1 microgram per liter 

(μg/L). One blank sample had sample concentrations of 

CB, 14DCB, and 12DCB between 1 and 2 μg/L, which is 

negligible given the high concentrations observed at the site. 

Sequential duplicate samples were used to estimate 

measurement precision for analyses of chlorinated benzenes 

and benzene. Relative sample difference (RSD) of duplicate 

sample measurements for a given constituent provides an 

indication of general precision. The RSD for each constituent 

was computed as follows:

 ( 2 1 ) /( ( 1 2 )) 100 R i SD = ABS  X − X Mean X + X  ∗   (6)

where

RSDi = the relative sample difference for a given 

constituent X and the ith duplicate pair 

(i = 1, 2, 3 or 4), expressed in percent;

ABS = the absolute value of the entire subsequent 

calculation;

X2–X1 = the difference in measurements between 

first sample collected (X1) and the second 

sample (X2) collected sequentially after X1; 

and

Mean (X1+X2 ) = the value used to normalize their difference 

for comparative purposes.

A total of 18 duplicate samples were collected from 

the following drive-point piezometers: DP-15A, DP-15B, 

DP-12A, DP-12B, DP-10B, DP-107A, and DP-104A. The 

overall average RSD (benzene and chlorinated benzenes) for 

the 18 duplicates included in this analysis was 10 percent. 

Concentrations for samples included in this analysis ranged 

from below detection to over 12,000 μg/L. Average RSD 

values ranged from 6.5 percent for 12DCB to 13.6 percent for 

124TCB. A total of 6 out of 133 individual RSD values were 

excluded from these calculations as extreme outliers, possibly 

due to compound misidentification. 

RSDs of duplicate sample pairs also were calculated for 

field parameters and other geochemical constituents. Specific 

conductance and pH had an average RSD of less than 1 percent. Ammonia, ferrous iron, and sulfide had average RSDs of 

20 percent or lower based on 13, 11, and 14 duplicate pairs, 

respectively. Total iron had an average RSD of 23 percent, 

based on 11 duplicate pairs. Based on 38 duplicate pairs, 

NVDOC, chloride and sulfate measurements had average RSD 

values of 18, 3, and 12 percent, respectively. Methane analyses 

had an average RSD of 10 percent, based on 26 duplicate 

pairs. 

Several analyses (less than 5) for benzene and chlorinated benzenes were completed outside of the recommended 

hold time, which is 2 weeks (monthly samples taken during January, February, April, May, and July of 2012; and the 

October 2011 synoptic event). Typically, the hold time was not 

longer than 3 weeks, however, an additional experiment was 

designed to determine if longer hold times affected the concentrations of contaminants in the sample. The RSD between 

the initial and first week measurements were under 15 percent, 
Methods of Data Collection and Analysis 21

except for benzene, 12DCB and 124TCB, which had RSDs 

of 21, 16, and 20 percent, respectively. However, subsequent 

differences between week one and week five were all under 10 

percent. For VOC analyses, analytical precision of 30 percent 

is considered acceptable. 

The measurement precision among the different constituents was taken into consideration for the interpretation of this 

report. However, with the wide range of concentrations found 

in the study area, the RSDs are within acceptable limits to 

describe patterns in the occurrence and distribution of constituents in the groundwater at the SCD. 

Experiments to Assess Bioremediation

In Situ Microcosms

In situ microcosms allow assessment of bioremediation 

potential under field conditions, including natural groundwater 

flow rates, and with native microbial populations. The in situ

microcosms utilized for field assessment of biodegradation 

of the chlorobenzene compounds and benzene at the SCD 

are Bio-Trap units (Microbial Insights, Rockford, Tennessee) 

contained in a 35.6-cm (14-inch [in.])-long, 3.18-cm 

(1.25-in.)-diameter screened (10 slot) PVC section, along 

with other customized passive diffusion samplers to collect 

groundwater for chemical analyses (fig. 4). Bio-Traps contain 

small beads, called Bio-Sep beads, made of a non-reactive, 

high-surface-area carbon material that provides a surface for 

attachment and subsequent collection of the native subsurface 

microorganisms (White and others, 2003; Peacock and others, 

2004; Geyer and others, 2005; Weiss and Cozzarelli, 2008; 

Burns and others, 2013). The Bio-Sep beads also can be preloaded with amendments, including bioaugmentation cultures 

and 13C-labeled contaminants, to assess specific issues, such 

as the efficiency of natural biodegradation compared to 

enhanced bioremediation and the relative rate of contaminant 

degradation.

The in situ microcosms were pushed directly into the wetland sediment, with each PVC unit containing one treatment 

and separated from other treatments pushed to the same depth 

by a lateral distance of about 25 cm. In October 2009, three 

sets of in situ microcosm devices were deployed for 8 weeks 

at two sites, 6 and 9 (fig. 3), to evaluate natural attenuation 

(unamended device), biostimulation (device amended with 

lactate and chitin as an electron donor), and bioaugmentation 

(device amended with lactate, chitin, and the WBC-2 culture) 

(fig. 4; table 4). Two additional in situ microcosms, a natural 

attenuation treatment and a bioaugmented treatment, were 

constructed with an additional microbial sampler containing 

Bio-Sep beads pre-loaded with 13C-labeled monochlorobenzene and deployed at site EM-6. The 13C-labeled Bio-Traps 

were used to evaluate degradation of the contaminant by loss 

of the 13C-labeled contaminant over the incubation period and 

incorporation of 13C from the labeled monochlorobenzene in 

dissolved inorganic carbon (DIC, reported as carbon dioxide, 

CO2) and in phospholipid fatty acid (PLFA). The presence of 13C in DIC indicates mineralization of the labeled contaminant. 

PLFA is a primary component of the membrane of all living 

cells and decomposes quickly upon death; thus, 13C incorporation in PLFA provides a measurement of contaminant utilization for growth by living biomass. In October 2010, a total 

of 36 in situ microcosms were deployed for 7 weeks at four 

sites, including two sites that were the same or near those used 

for testing in 2009 (sites 6, 8, 104, and 107 in fig. 3; table 4). 

Treatment sets (unamended, biostimulated, and bioaugmented) 

at all sites included control microcosms that were not amended 

with 13C compounds and sets amended with 13C-labeled contaminants (table 4). 13C-14DCB treatments were deployed at 

all four sites, and 13C-CB treatment sets were deployed at three 

sites with a 13C-benzene treatment set deployed instead at the 

fourth site. Screens of the microcosm devices were placed in 

the wetland sediment at a depth of 25 to 56 cm below land 

surface in 2009 and 2010.

Upon retrieval of the in situ microcosms, groundwater 

diffusion samplers for collection of VOCs, anions (sulfate, 

chloride, nitrate, and phosphate), and volatile fatty acids 

were sealed with Teflon-lined caps and placed in coolers 

for transport to the laboratories. VOCs were analyzed in the 

USGS MD-DE-DC WSC laboratory as described above for 

groundwater sampling, whereas anions and volatile fatty 

acids were sent to Microbial Insights for analysis by ion 

chromatography. Groundwater diffusion samplers for the 

Groundwater diffusion sampler–samples for

inorganics and volatile fatty acids analyses.

Groundwater diffusion sampler–samples for

volatile organic compounds and methane.

Microbial sampler with Bio-Sep beads–

unamended Bio-Sep beads for microbial

community attachment. Pre-loaded with

WBC-2 (a dechlorinating consortium of

microorganisms) for bioaugmented

treatments.

Donor diffusion sampler–filled with lactate

for biostimulated treatments. Solid polyvinyl

chloride (PVC) spacer used for treatments

without donor.

Solid donor porous baggie–containing

chitin as an additional slow donor release

for biostimulated treatments. Solid PVC

spacer used for treatments without donor.

Figure 4. Schematic diagram of Bio-Trap (Microbial Insights, 

Rockford, Tennessee) sampler used for in situ microcosms.
22 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

redox-sensitive constituents ammonia, ferrous iron, and sulfide 

were sealed without headspace in VOC vials with Teflonlined caps; samples were analyzed by using colorimetric 

methods (as described in the Geochemical Data section above) 

either immediately in the field or within 24 hours of collection at the USGS MD-DE-DC WSC laboratory. Methane, 

ethane, and ethene analyses by GC/FID were conducted by 

Microbial Insights for the 2009 microcosms and at the USGS 

MD-DE-DC WSC laboratory for the 2010 microcosms. 

The Bio-Traps from the in situ microcosms were immediately placed in zippered bags and sent to Microbial Insights for 

microbial community and 13C analyses using methods detailed 

in Burns and others (2013). Microbial genetic analyses 

included quantitative polymerase chain reaction (qPCR) 

analysis for total eubacteria and for specific target populations 

known to be involved in anaerobic degradation of chlorinated 

solvents. Although these target population tests originally were 

developed specifically for evaluation of reductive dechlorination of chlorinated ethenes and ethanes, several of the target 

populations also have been shown to be involved in reductive 

dechlorination of chlorobenzene compounds. General composition of the microbial communities present in the in situ

microcosms was obtained by analysis of the PLFA structural 

groups (table 5). Microbial populations colonizing the in 

situ microcosms deployed in 2010 also were analyzed using 

denaturing gradient gel electrophoresis (DGGE) and excising 

selected bands for sequencing and identification. During the 

2009 in situ microcosm retrieval, wetland sediment samples 

were collected from depths of 12.7 cm and 27.9 cm below 

land surface at the two microcosm sites and sent to Microbial 

Insights for analysis of microbial communities by qPCR and 

PLFA structure for comparison to the Bio-Trap samplers. 

Isotope analyses included the amount of 13C-labeled contaminant, 13C-CO2

, and 13C-PLFA in the microcosms after incubation compared to quantities measured before the Bio-Traps 

were deployed. 

Laboratory Bioreactors

Three bench scale, static bed bioreactor systems were 

operated in an on-site trailer for the wetland study (a diagram 

of one bioreactor system is shown in figure 5). Two bioreactor systems were seeded with WBC-2 (“D” and “E”), and a 

third bioreactor was seeded only with bacteria derived from 

the site groundwater pumped through the bioreactor (“C,” or 

control bioreactor). The D and E bioreactors were previously 

Table 4. In situ microcosm treatments deployed from October 20–December 16, 2009 and from October 26–December 14, 2010 at 

selected wetland sites, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

[Treatments included MNA, monitored natural attenuation (control); lactate/chitin, donor addition for biostimulation; and WBC-2, West Branch Consortium-2, 

for bioaugmentation; WBC-2 treatment also contained added lactate/chitin; --, no treatment deployed]

Original site 

name

Site 

number

Sample 

date

Control 

treatments

13C-labeled treatments

13C-Chlorobenzene 13C-1,4-Dichlorobenzene 13C-Benzene

PDB-10-28 DP-6 12-16-09 MNA MNA -- --

PDB-10-29 DP-6 12-16-09 Lactate/chitin Lactate/chitin -- --

PDB-10-30 DP-6 12-16-09 WBC-2 WBC-2 -- --

DP-9 12-16-09 MNA -- -- --

DP-9 12-16-09 Lactate/chitin -- -- --

DP-9 12-16-09 WBC-2 -- -- --

PDB-01-01 DP-8 12-14-10 MNA MNA MNA --

PDB-01-02 DP-8 12-14-10 Lactate/chitin Lactate/chitin Lactate/chitin --

PDB-01-03 DP-8 12-14-10 WBC-2 WBC-2 WBC-2 --

PDB-04-10 DP-104 12-14-10 MNA MNA MNA --

PDB-04-11 DP-104 12-14-10 Lactate/chitin Lactate/chitin Lactate/chitin --

PDB-04-12 DP-104 12-14-10 WBC-2 WBC-2 WBC-2 --

PDB-07-19 DP-107 12-14-10 MNA MNA MNA --

PDB-07-20 DP-107 12-14-10 Lactate/chitin Lactate/chitin Lactate/chitin --

PDB-07-21 DP-107 12-14-10 WBC-2 WBC-2 WBC-2 --

PDB-10-28 DP-6 12-14-10 MNA -- MNA MNA

PDB-10-29 DP-6 12-14-10 Lactate/chitin -- Lactate/chitin Lactate/chitin

PDB-10-30 DP-6 12-14-10 WBC-2 -- WBC-2 WBC-2
Methods of Data Collection and Analysis 23

bioaugmented with 1 L of WBC-2 microbial consortium and 

had viable populations of the consortium established during 

operation for a previous USGS-Geosyntec collaborative study 

to assess biodegradation of chlorinated ethanes and ethenes 

(Geosyntec Consultants, 2012).

The bioreactor systems each consisted of a groundwater 

equalization/mixing tank, three 1-L PVC vessels filled with 

a high-surface area polypropylene support matrix for biofilm 

growth, and an effluent (waste) tank (fig. 5). Water was 

pumped to the first reactor vessel with a peristaltic pump, 

equipped with Viton tubing. Effluent from the first vessel was 

transported by gravity to the second and third vessels and 

finally to the effluent tank. Two sample ports were located on 

each vessel, and samples also were collected from the groundwater equalization and effluent tanks. An argon gas blanket 

was maintained in the headspace of each bioreactor vessel by 

delivering a slow, continuous flow of argon to the top of each 

bioreactor vessel from a compressed gas cylinder. Because 

argon has a density greater than air, it would sink and expel air 

from the top of the bioreactor to help keep the system anaerobic. Operation of the three bioreactors began in March–April 

2012, with the bioaugmented D and control C bioreactors 

receiving groundwater inflow from piezometer DP-15B and 

the bioaugmented E bioreactor receiving groundwater from 

piezometer DP-10B. These piezometers, selected on the basis 

of contaminant concentrations and ease of pumping, were 

located within the wetland area but screened at shallow depths 

in the aquifer underlying the wetland sediments (table 2, 

fig. 3). The bioreactors were operated through the end of 

October 2012, with varying amendments during selected 

periods to evaluate factors affecting biodegradation of the 

chlorobenzene compounds and benzene present in the influent 

groundwater. 

Operational maintenance was performed several times 

a week and involved refilling the groundwater equalization 

tank and maintaining the peristaltic pumps and associated 

tubing. The equalization tank was filled with a combination of 

the influent groundwater from DP-10B or DP-15B, nutrients, 

electron donor, and a pH buffer to support microbial growth 

and anaerobic degradation of the contaminants. A combination 

of ammonia chloride and potassium phosphate were used as 

the nutrient stock solution, with optimal concentrations of 3 

mg/L (as N) and 0.3 mg/L (as P), respectively. A mixture of 

lactate and corn syrup was added to provide electron donors; 

donor concentrations were monitored by measuring the chemical oxygen demand (COD) with desired concentrations of 300 

mg/L in the influent and about 50–100 mg/L remaining in the 

waste. Sodium bicarbonate (NaHCO3) was used as a buffer 

to adjust the pH of the influent water to about 7.0. To evaluate the activity of WBC-2 compared to the native microbial 

community in the bioreactors, the bioreactors were amended 

with the main VOC that WBC-2 was developed to degrade, 

1,1,2,2-tetrachloroethane (TeCA), for a period of about 80 

days in addition to the site contaminants. Site contaminants 

were delivered throughout the bioreactor experiment only 

from those naturally present in the site groundwater used to 

fill the inflow tanks, and concentrations varied with changes in 

the groundwater at the piezometer sites during the experiment 

and with VOC losses during groundwater pumping, transport, 

and filling of the influent tanks and during the holding time in 

the tanks. The flow rate for all of the bioreactors ranged from 

Table 5. Description of phospholipid fatty acid (PLFA) structural groups that are used to describe microbial community composition 

on the basis of fatty acid biosynthesis.

PLFA sructural group Community classification Description

Polyenoics Eukaryotes

Found in eukaryotes such as fungi, protozoa, algae, higher plants, and 

animals; includes scavengers that will prey on contaminant-utilizing 

bacteria.

Normal Saturated (Nsats) General Found in all organisms. High proportions indicate a low diversity in 

microbial populations.

Mid-Chain Branched Daturated 

(MidBrSats) Actinomycetes High G+C gram-positive cell membrane structure common in sulfatereducing bacteria and Actinobacteria.

Branched Monoenoic (BrMonos) Anaerobic metal reducers Found in the cell membranes of micro-aerophiles and anaerobes; often 

associated with anaerobic sulfate and iron-reducing bacteria.

Monoenoic (Monos) Proteobacteria 

One of the largest groups of bacteria (gram negative bacteria) and 

represents a wide variety of aerobes and anaerobes, including the 

majority of bacteria that utilize hydrocarbons.

Terminally Branched Saturated 

(TerBrSats) Firmicutes 

Indicative of presence of anaerobic fermenting bacteria (mainly 

Clostridia and Bacteriodes), which are important for breakdown of 

organic substrates to a form utilized by bacteria during reductive 

dechlorination.
24 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

1 and 3 milliliters per minute (mL/min) with a measurement 

precision of 0.1 mL/min. 

Samples were taken from various ports in the vessels and 

from the influent and waste tanks several times a week and 

analyzed for VOCs, methane, ammonia, reactive phosphate, 

sulfate, sulfide, pH, and oxidation-reduction potential (ORP). 

VOCs and methane were analyzed in the USGS MD-DE-DC 

WSC laboratory, as explained for the groundwater sampling. 

Duplicate VOC samples were taken sequentially from one 

sampling port throughout the experiments to obtain 33 pairs of 

duplicates. Average RSDs for the duplicate pairs ranged from 

7.2 to 15 percent for all VOCs analyzed, except for trichloroethene (one of the daughter products of TeCA), which was 

20 percent. COD samples were collected in 8-mL borosilicate 

vials and processed in the USGS MD-DE-DC WSC laboratory 

using Hach TNTplusTM vials and analyzed on a Hach DR 

2800TM spectrophotometer (Hach Company, 2007). Ammonia, 

phosphate, sulfate, and sulfide also were analyzed using Hach 

test kits and a spectrophotometer but were processed on-site 

immediately after sample collection. The pH was measured 

with a Hach HQ 40d multi-parameter meter with a standard 

gel-filled pH probe that was calibrated daily with pH 4 and 

7 buffers. The ORP was measured with a Hach ORP Pocket 

Pal TesterTM in millivolts (mV). The calibration of the ORP 

tester was confirmed monthly with GLI, Internationalʼs +200 

mV ORP Reference Solution (Hach Company, Loveland, 

Colorado). 

Hydrogeologic Characterization

In order to assess the distribution, migration, and potential for degradation of contaminants in the wetland study area 

at the SCD, an understanding of the hydrogeologic context 

for contaminant transport is critical. Prior to this study, the 

understanding of the hydrology of the SCD was limited to the 

uplands. The following section provides this hydrogeologic 

context by presenting the lithology of the wetland study area at 

the SCD, analysis of groundwater-flow directions and rates in 

the wetland and underlying Columbia aquifer, and an estimate 

of groundwater flux to Red Lion Creek. 

Groundwater

Equalization

Tank Not to scale

Bicarbonate

Stock

Solution

Nutrient

Stock

Solution

Electron

Donor

Solution

Sampling

Ports

pH

ORP

ORP

ORP

pH Electrode ORP Electrode

Peristaltic

Pump

Bioreactor

Vessel 1 Bioreactor

Vessel 2 Bioreactor

Vessel 3 Sludge

Separation

Settled

Sludge

Clarified

Effluent

Clarified

Effluent

1

0, Influent

8, Waste

3

5

7

2

4

6

Argon

EXPLANATION

1 Designated sampling locations

ORP Oxidation Reduction Potential

Figure 5. Diagram of the anaerobic bioreactor system and sampling locations.
Hydrogeologic Characterization 25

Lithology of the Study Area

The lithology of the USGS wetland study area at the SCD 

was assessed using core logs from drive-point piezometers that 

were installed by USGS staff as part of this study and from 

previously installed wells along the wetland-upland boundary 

of the wetland study area along Haul Road (fig. 3; Black & 

Veatch, 2007). The upland-wetland boundary occurs at the primary change in slope south of Haul Road and parallel to Red 

Lion Creek (fig. 3b). Boring logs recorded as part of this study 

provide descriptions of wetland sediments and the Columbia 

Formation within the wetland study area, and provide a more 

detailed characterization of wetland sediments in the vicinity 

of Red Lion Creek. Lithology is shown along two transects 

through the wetland study area—transect A–A′, which extends 

to the northeast from the east side of Haul Road in the upland 

area and across Red Lion Creek, and transect B–B′, which 

extends to the northwest from the west side of Haul Road in 

the upland area to the south edge of Red Lion Creek (figs. 3a, 

6, 7). Descriptions of regional stratigraphy from the Delaware 

Geological Survey (Ramsey, 2007) and site-specific stratigraphy from Black & Veatch (2007) also were used to describe 

the Columbia aquifer and the underlying Merchantville and 

Potomac Formations. 

Wetland sediments at the SCD can be grouped into two 

stratigraphic units, an upper organic-rich unit consisting of 

root mass and organic matter in a mineral matrix of silt and 

(or) clay, and a basal unit consisting of clay, silt, and finegrained sand. The upper, organic-rich unit is typically dark 

brown to black with scattered grey clay layers, with a median 

carbon content of 7.7 percent and a maximum carbon content 

of 24.2 percent. Carbon content was measured in sediment 

cores collected from the upper 0.44 to 0.6 m of wetland sediments near the upland-wetland boundary. Organic carbon 

content may be higher in the thick organic units adjacent to 

Red Lion Creek, but was not quantified. The lithology of the 

basal wetland sediment unit varies laterally and includes interfingering lenses of grey to tan and brown clay with fractions 

of silt and (or) sand, and lenses that are predominantly tan to 

brown silt with fractions of clay. Orange or reddish tan clay, 

silty clay, sandy clay, and clayey sand have been noted along 

transect B–B′ at drive-point piezometer nests DP-10, DP-11 

and DP-13 (fig. 7), as well as in sand and clayey sand units 

along transect A–A′ at drive-point piezometer DP-15B, and on 

the north side of the creek at drive-point piezometers DP-17B 

and DP-18B (fig. 6). 

Wetland sediments vary laterally in thickness. Transect 

A–A′ shows a profile of an incised channel, with lenses of clay 

along the creek banks and predominantly silt and organic fill 

in the channel center (fig. 6). Wetland sediments have been 

estimated to extend to greater than 10 m below land surface at 

DP-16B, however, further data collection would help to define 

the interface between wetland sediments and the Columbia 

aquifer in this area (fig. 6). Transect B–B′ shows upper 

organic-rich wetland sediments deepening from the uplandwetland boundary toward the creek to greater than 2 m thick. 

This organic-rich unit is underlain by silty clay, clay and sandy 

clay up to 2 m thick (DP-11B, fig. 7). The combined thickness 

of upper organic rich and basal fine-grained mineral wetland 

sediment units along this transect is as much as 5.2 m at the 

edge of Red Lion Creek at DP-13B (fig. 7). Wetland sediments 

on site are underlain by the Columbia Formation. 

The Columbia Formation has been described as a fine- to 

medium-grained feldspathic quartz sand, which is typically 

yellow to reddish-brown in color, and with varying amounts of 

gravel and intermittent layers of tan to reddish grey, clayey silt 

(Ramsey, 2007). Upper units of the Columbia Formation have 

been reported to contain grey to red-brown silt and very fine 

sand underlain by medium to coarse sand, whereas the base 

of the unit contains quartz and chert gravel beds with cobble 

and boulder-sized clasts of sandstone, siltstone, shale, pegmatite, schist and amphibolite (Ramsey, 2007). This formation is 

thought to have been formed as a result of fluvial deposition of 

glacial outwash filling an eroded, non-planar topographic surface (Ramsey, 2007). Previous investigations have shown that 

this unit ranges from approximately 2.5 to 25 m in thickness 

at SCD, and is capped by wetland sediments in the vicinity 

of Red Lion Creek (Black &Veatch, 2007). Boring logs from 

drive-point piezometers installed in the wetland study area 

describe the Columbia aquifer as a brown to orange-brown, 

medium to coarse, poorly sorted, angular to sub-angular sand 

and gravel with fine to coarse sand. The Columbia sands have 

silty (DP-11B) and clayey (DP-13B) layers along transect 

B–B′ (fig. 7), and a gravel layer that is greater than 1 m thick 

is noted at DP-15B along transect A–A′ (fig. 6). Well logs 

from previously installed wells in the upland part of the study 

area give similar descriptions for the Columbia Formation 

(Black & Veatch, 2007). The Columbia Formation is underlain 

by the low permeability Merchantville Formation, and in areas 

where the Merchantville is absent, is underlain by low permeability beds of the upper Potomac Formation (Black & Veatch, 

2007).

Drive-point piezometers did not extend into the 

Merchantville or Potomac Formations, but they are described 

here to frame the regional geology. The Merchantville in the 

area of the SCD has been described as a grey to green-grey 

glauconitic micaceous clay to silty or sandy clay (Black & 

Veatch, 2007). At the SCD, the unit has been reported to have 

a maximum thickness of 6.7 m and an average thickness of 3.1 

m where present, but has been found to be laterally discontinuous. The Merchantville Formation has been considered to be 

a confining unit regionally, separating the surficial Columbia 

aquifer from the Potomac aquifer. However, numerous studies in the vicinity of the SCD have described absence of the 

Merchantville in boring logs, erosion of the Merchantville 

prior to deposition of the Columbia aquifer, and hydraulic connection between the Columbia and Potomac aquifers (Black 

& Veatch, 2007). Black & Veatch reported in their 2007 

Final Remedial Investigation Report that the Merchantville 

Formation was either missing or could not be distinguished 

from the upper Potomac Formation in nearly half of boring 

logs evaluated from SCD (Black & Veatch, 2007). 
26 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

The Potomac Formation consists of red, grey, pink, and 

white silty clay and clayey silt, interbedded with very fine to 

medium sand units of fluvial deposition (Ramsey, 2005). The 

top of the Potomac at SCD consists of predominantly clayey 

sediments, which may have eroded into paleochannels (Black 

& Veatch, 2007). 

Groundwater Hydraulic-Head Distribution and 

Flow Direction

Potentiometric surface maps for the Columbia aquifer 

and the wetland sediments were constructed for five dates 

between June 22, 2011 and September 30, 2012 in order to 

examine changes in flow direction and hydraulic gradient 

that could result from seasonality or in response to changing 

hydrologic conditions. Potentiometric surface maps for the 

Columbia aquifer and the wetland sediments are shown in 

figures 8 and 9, respectively for August 8, 2012, which is considered representative of “normal” precipitation and hydraulic 

head conditions, with 10 cm total precipitation during the 

preceding month and active photosynthesis. Little to no 

temporal variations in flow direction in the Columbia aquifer 

and minimal variation in flow direction in the shallow wetland 

sediments were observed compared to this period of “normal” 

conditions, although other dates included extreme events 

(fig. 10): 

1.June 22, 2011—relatively dry conditions, preceding 

month had 3.8 cm total precipitation during a period of 

high photosynthetic activity; date also precedes 

Hurricane Irene and Tropical Storm Lee; 

A

(Southwest)

Meters Meters

(Northeast)

A'

5 5

5 5

10 10

15 15

20 20

VERTICAL EXAGGERATION = 10X

EXPLANATION

Sea

Level

Sea

Level

MW-23

Land

DP-14B DP-15B

surfaceDP-16B DP-17B DP-18B

MW-37

Red Lion Creek

?

?

?

?

?

?

?

? ?

?

?

Clay

Gravel

Geologic contact (dashed and

queried where approximate)

Monitoring well (MW) and drive-point

piezometer (DP) locations, screened

interval, and identifier

Clay/sand

Clay/silt

Organics

Organics/clay

Sand

Sand/clay

Sand/gravel

Silt

Silt/clay

?

MW-23

0

0

20 40 60 80 METERS

100 200 FEET

Figure 6. Hydrogeology and locations of monitoring wells and drive-point piezometers along section A–A’, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware.
Hydrogeologic Characterization 27

2. September 14, 2011—very wet conditions, preceding 

month had 41 cm total precipitation; date is immediately following Hurricane Irene and Tropical Storm 

Lee and potential damage to the tide control structure 

on Red Lion Creek;

3. December 8, 2011—high moisture conditions, preceding month had 16 cm total precipitation and this date 

immediately follows a storm during a period of minimal photosynthetic activity; this is a very high point 

in the hydrographs for nearly all of the drive-point 

piezometers in the SCD wetland study area;

4. February 16, 2012—relatively dry period, preceding 

month had a total of 5 cm of total precipitation, during 

a period of low photosynthetic activity.

Hydraulic head distributions measured across the SCD 

showed that lateral flow in the Columbia aquifer is toward Red 

Lion Creek on both its north and south sides (fig. 8), which is 

consistent with previous studies of flow direction at the SCD 

(Black & Veatch, 2007). Although flow generally follows 

topography, the containment barrier installed in 2007 surrounding the developed part of SCD produces an artificially 

low gradient north of the upland-wetland boundary (fig. 8). 

Figure 7. Hydrogeology and locations of monitoring wells and drive-point piezometers along section B–B’, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware.

? ? ?

?

?

?

?

?

?

?

?

?

?

?

?

?

?

?

B

(South)

Meters

(North)

B'

4

2 2

2 2

4 4

6 6

8 8

Sea

Level

Sea

Level

VERTICAL EXAGGERATION = 7X 0

0 20 40 60 FEET

5 10 15 20 25 METERS

EXPLANATION

Meters

4

MW-22

DP-10B

DP-11B DP-13B

Land surface

Red Lion Creek

Clay Geologic contact (dashed and

queried where approximate) Clay/sand

Clay/silt

Organics

Organics/clay

Organics/silt

Sand

Sand/clay

Sand/silt

Silt/clay

MW-22 Monitoring well (MW) and drive-point

piezometer (DP) locations, screened

interval, and identifier

?

Clay/organics
28 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12 STANDARD CHLORINE OF DELAWARE, INC. SUPERFUND SITE STANDARD CHLORINE OF DELAWARE, INC. SUPERFUND SITE 75°38'20" 75°38'10" 75°38'00" 75°37'50" 39°36"20" 39°36"10" Imagery from U.S. Geological Survey, Delaware Valley Regional Planning Commission, and Department of Homeland Security, 2010, NAD 1983 State Plane Delaware

Transverse Mercator Projection

9

Road River 9

Red Lion Creek Red Lion Creek

Haul Road

Haul Road

N

Unnamed

ributary T

00

100 200 METERS

600 FEET 300

MW-10

(3.057)

PMW-48

(3.2)

EM-5D

(0.808)

OR-6C(0.408)

MW-9

MW

(2.664)

-8

MW

(1.591)

-19

(0.856) MW-20

(0.863)

MW-24

(0.515)

MW-40

(0.68)

MW-25

(0.646)

MW-23

(0.674)

MW-21

(0.783) MW-22 (0.805)

DP-13B

(0.235) DP-12B

(0.253)

DP-14B

(0.436)

DP-15B

(0.29)

DP-10B

(0.244)

DP-11B

(0.488)

PZ-15

(1.059)

PMW-47

(1.203)

PZ-12

(0.939)

PZ-7

(2.427)

PZ-5

(2.689)

MW-18

(3.146)

MW-5

(1.881)

CW-11

(1.603)

EM-6D

(0.567)

EM-4D

(1.183)

PZ-18

(2.086)

PZ-17

(2.053)

MW-17

(3.036)

Surface-water gage

(-0.05)

MW-36

(0.643)

MW-37

(0.351)

MW-38

(0.393)

DP-18B

(0.238)

DP-17B

(0.094)

EXPLANATION

Equipotential line; contour interval 0.2 meters(dashed where approximate)Generalized groundwater-flow direction

Drive-point piezometer or monitoring well,

Barrier wall

identifier, and groundwater altitude, in metersSurface-water gage and water elevation,in meters

OR-6C(0.408)

(-0.05)

0.4

0.6

0.4

0.2

0.2

0.4

0.6

0.6

0.8

0.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

2.4

2.6

2.8

3.0

3.2

1.0

1.2

1.4

1.6

1.8

2.0

2.2

2.42.6

2.8

3.0

3.2

0.4

Figure 8. Groundwater altitude (in meters above mean sea level) in the Columbia aquifer, August 8, 2012, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.
Hydrogeologic Characterization 29

On the south side of Red Lion Creek, north of the barrier wall 

and within the wetland study area, lateral groundwater flow 

radiates generally out into the wetland about a flow divide 

between sites 12 and 15. Flow on the east side of this boundary is slightly toward the northeast, and flow on the west side 

of the boundary is slightly to the northwest (fig. 8). 

Wetland sediments at the SCD are present in the lowlying fringe of Red Lion Creek, with less than 1 m of topographic variation from the boundary of the wetland to open 

water on Red Lion Creek. Though topographic variation 

in this area is very shallow, lateral groundwater flow in the 

wetland sediments reflects the topography near the uplandwetland boundary and is generally toward Red Lion Creek 

(fig. 9). Near the creek, where the thickness of wetland sediments increases, hydraulic head in the wetland sediments is 

very close to that of the creek, and the influence of lateral 

movement of groundwater is muted by upward components of 

flow (fig. 9). 

Although groundwater-flow directions in the Columbia 

aquifer and wetland sediments were quite similar, the distribution of lateral hydraulic gradient differed between the units. 

Lateral hydraulic gradients in the Columbia aquifer were 

calculated using the head map constructed from water levels 

measured on August 8, 2012, and were fairly consistent from 

east to west (fig. 8). A gradient of 0.006 meters per meter 

(m/m) to the north/northeast was calculated along transect 

A–A′ on the eastern side of the wetland flow divide, and a 

gradient of 0.0068 m/m to the north/northwest was calculated 

along B–B′ on the west side of the wetland. This is consistent 

with the previously reported average gradient of groundwater 

in the Columbia aquifer of approximately 0.005 feet per feet 

(ft/ft) (or m/m) to the north/northeast (Black & Veatch, 2005). 

75°38'20" 75°38'10" 75°38'00" 75°37'50"

39°36'20"

39°36'10"

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

9

9

N

River Road River Road

Red Lion Creek

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

Unnamed Unnamed

Tributary Tributary

Surfacewater gage

(-0.05)

Haul Road

0

0

100 200 METERS

300 600 FEET

DP-8A

(0.125)

DP-13A

DP-11A (-0.055)

(-0.018)

DP-10AA

(0.381) DP-104A

(0.387)

DP-14A

(0.369)

DP-107A

(0.485)

DP-12AA

(-0.058) DP-16AA

(-0.088)

DP-15AA

(0.098)

DP-6A

(0.354)

DP-18AA

(0.183)

DP-17A

(0.082)

EXPLANATION

Equipotential line; contour interval 0.2 meters

(dashed where approximate)

Generalized groundwater-flow direction

Barrier wall

Drive-point piezometer, identifier, and

groundwater altitude, in meters

Surface-water gage and water elevation,

in meters

DP-8A

(0.125)

(-0.05)

0.4

0.2

0

0

0

0.2

0.2

0.4

0.4

Figure 9. Groundwater altitude (in meters above mean sea level) in the wetland sediments, August 8, 2012, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware.
30 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Lateral gradients were calculated using a groundwater head 

map from the same date for wetland sediments, and these lateral gradients showed greater variation from east to west than 

those calculated for the Columbia aquifer (fig. 9). Gradients 

were steeper in wetland sediments along transect B–B′ 

(0.013 m/m) on the west side of the flow divide than along 

A–A′ (0.007 m/m) on the east side. Lateral gradients calculated from piezometer to piezometer in both the Columbia 

aquifer and wetland sediments varied with proximity to the 

upland boundary with the wetland, with gradients more than 

an order of magnitude greater near the upland-wetland boundary than those observed between piezometers located closer to 

the creek (data on file at the MD-DE-DC WSC). This is due in 

part to the strong component of upward vertical groundwater 

discharge that was observed near the creek. 

Upward vertical flow of groundwater between the 

Columbia aquifer and overlying wetland sediments was apparent at most locations where groundwater heads were measured 

across the wetland study area. Of the drive-point piezometers 

installed in the wetland study area, only two showed variation 

between groundwater recharge and discharge conditions 

(vertically upward and downward gradients, respectively). 

Alternating gradients were observed at the EM-1 and DP-14 

drive-point clusters, whereas all other sites showed consistent vertically upward gradients (table 6). The magnitude of 

upward discharge of groundwater in the wetland showed only 

small variation seasonally and spatially, ranging from -0.01 

m/m to -0.16 m/m for selected dates in 2011 and 2012 

(table 6). Site DP-12, which is located at the boundary 

between the wetland and open water on Red Lion Creek, had 

the largest vertically upward hydraulic gradients site-wide for 

the dates that were evaluated. In contrast, the smallest gradient 

was observed at site EM-1 (-0.01 m/m), which is the southernmost drive-point piezometer and farthest from the creek of the 

sites that were evaluated. 

Overall, the magnitude of lateral and vertical hydraulic 

gradients supports the conceptual model for groundwater flow 

to Red Lion Creek in the wetland study area, which is laterally toward the creek from the north and south and vertically 

upward through the wetland sediments. Cross-sectional flow 

diagrams constructed for transects A–A′ (fig. 11) and B–B′ 

(fig. 12) show that flow is lateral from the south and north 

and becomes increasingly vertical with proximity to the creek 

center.

EXPLANATION

RED LION CREEK

Red Lion Creek

Average following Hurricane Irene

Average prior to Hurricane Irene

Precipitation Dates selected for analysis

EM-2D DP-10AA

Hurricane Irene,

August 2011

DP-10B DP-13B

DRIVE-POINT PIEZOMETERS

1.2

1.0

0.8

0.6

0.4

0.2

0.0

Water-level altitude,

in meters relative to mean sea level

Daily total precipitation,

in centimeters

0.2

0.4

0.6

0.8

Date

7/1/099/1/09

11/2/09

8/1/09

1/3/103/6/10

10/2/0912/3/09

2/3/104/6/105/7/106/7/107/8/108/8/109/8/10

10/9/1011/9/10

12/10/10

1/10/112/10/113/13/114/13/115/14/11

6/14/117/15/118/15/11

9/15/11

10/16/11

11/16/11

12/17/11

1/17/122/17/123/19/124/19/125/20/126/20/127/21/128/21/129/21/12

14

12

10

8

6

4

2

0

Figure 10. Water-level altitudes of Red Lion Creek and selected drive-point piezometers and daily total precipitation, July 2008–

August 2012, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.
Hydrogeologic Characterization 31

Hydraulic Conductivity

Hydraulic conductivity was measured by conducting slug 

tests at four drive-point piezometer nests in the wetland study 

area for use in determining rates of groundwater discharge. 

Hydraulic conductivity was generally higher in the Columbia 

aquifer compared to the wetland sediments, although measurements from both hydrologic units were within two orders 

of magnitude (table 7). Previous studies of hydraulic properties of wetland sediment in a similar Coastal Plain setting in 

Maryland showed hydraulic conductivity estimates ranging 

from 0.0003 to 0.04 feet per day (ft/d) in peat-dominated sediments and 0.0008 to 0.04 ft/d in clay-dominated wetland soils 

(Lorah and others, 1997). Slug tests were performed where 

responses to induced gradient were measureable; therefore, 

these estimates of hydraulic conductivity within the SCD 

wetland study area may be biased slightly high. However, 

estimates of hydraulic conductivity are within the range of 

values reported for non-peat wetland soils to moderately to 

slightly decomposed, northern peatland sediments (Mitsch 

and Gosselink, 1993). A slug test was initiated at DP-15A but 

was abandoned due to slow recovery, which likely would have 

produced an orders-of-magnitude smaller hydraulic conductivity value for surrounding wetland sediments compared to the 

other sites tested. 

Groundwater Discharge and Velocity

Because most drive-point piezometer nests were screened 

in the Columbia aquifer and wetland sediments, rates of vertical groundwater discharge were estimated using weighted values of hydraulic conductivity and vertical gradients calculated 

Table 6. Calculated vertical gradients from selected drive-point piezometer pairs in the Columbia aquifer and wetland sediments, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

[Negative gradient values indicate vertically upward discharge, whereas positive gradients indicate vertically downward movement of groundwater. See table 

2 for piezometer site information; --, no data] 

Site 

identification 

number

Drive-point 

pair Geologic unit

2011 2012

June July Sept. Dec. Jan. Feb. Aug.

EM-1 1S–1D Wetland sediment—Wetland sediment -- 0.03 -0.01 -0.05 -0.00 -0.00 0.05

EM-2 2S–2D Columbia aquifer—Columbia aquifer -0.04 -0.04 -0.05 -0.05 -0.05 -- -0.04

EM-5 5A–5D Columbia aquifer/Wetland sediment interface—

Columbia aquifer

-- -- -- -- -- -- -0.12

EM/

DP-6

6S–6D Columbia aquifer/Wetland sediment interface—

Columbia aquifer

-0.11 -0.10 -0.11 -0.12 -- -0.11 -0.09

6A–6D Wetland sediment—Columbia aquifer -0.12 -0.10 -0.12 -0.15 -- -0.11 -0.09

DP-10

10S–10B Wetland sediment—Columbia aquifer -0.04 -0.04 -- -- -0.07 -0.07 --

10AA–10B Wetland sediment—Columbia aquifer -- -- -- -0.12 -0.08 -0.07 -0.04

DP-11

11S–11B Wetland sediment—Columbia aquifer -0.10 -0.09 -0.09 -0.05 -0.10 -0.10 --

11A–11B Wetland sediment—Columbia aquifer -- -- -- -0.06 -0.11 -0.10 -0.07

DP-12

12S–12B Wetland sediment—Columbia aquifer -0.15 -- -- -0.11 -0.15 -0.13 --

12AA–12B Wetland sediment—Columbia aquifer -- -- -- -0.11 -0.16 -0.16 -0.11

DP-13

13S–13B Wetland sediment—Columbia aquifer -0.10 -0.12 -0.06 -0.04 -0.08 -0.09 --

13A–13B Wetland sediment—Columbia aquifer -- -- -- -0.05 -0.09 -0.07 -0.06

DP-14

14S–14B Wetland sediment—Columbia aquifer -- -- -- -0.08 0.01 0.00 --

14A–14B Wetland sediment—Columbia aquifer -- -- -- -0.11 -- -0.06 -0.02

DP-15

15S–15B Wetland sediment—Columbia aquifer -0.04 -0.05 -- -0.06 -0.05 -0.04 --

15AA–15B Wetland sediment—Columbia aquifer -- -- -- -0.07 -0.07 -0.06 -0.06

DP-16

16S–16B Wetland sediment—Wetland sediment -- -- -- -0.05 -0.05 -0.06 --

16A–16B Wetland sediment—Wetland sediment -- -- -- -- -0.08 -0.07 -0.05

DP-17 17A–17B Wetland sediment—Columbia aquifer -- -- -- -- -0.07 -0.06 -0.04

DP-18 18AA–18B Wetland sediment—Columbia aquifer -- -- -- -- -- -- -0.06
32 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

between these two hydrologic units at drive-point piezometer 

nests for which hydraulic conductivity values were available 

(table 7). Vertical specific discharge (Darcian velocity) was 

calculated using equation 1, and vertical linear velocity was 

calculated using equation 5 (table 8). Both the arithmetic 

and geometric means were calculated and the generally close 

agreement of the two means, for most sites, indicate that the 

data for the different variables (hydraulic gradient and hydraulic conductivity) generally did not contain extreme ranges that 

dominate the weighting. Results of these analyses showed a 

maximum specific discharge rate of 118 meters per year 

(m/yr) from the Columbia aquifer through the wetland 

sediments in the wetland study area (table 8). Groundwater 

discharge is spatially variable in the wetland study area, with 

minimal rates of discharge near the upland-wetland boundary 

at site EM/DP-6 (table 8). This site is on the southeast side 

of the upland-wetland boundary and set farther south of 

open water on Red Lion Creek than the other sites for which 

discharge was calculated. The greatest rates of discharge were 

obtained at site DP-15, where the piezometer screened within 

the Columbia aquifer (DP-15B) is in a gravel dominated part 

of the formation. It should be noted that hydraulic conductivity values for wetland sediments measured at DP-6A were 

used for this estimate of vertical discharge rather than those 

measured at DP-15A because the slug test at DP-15A had to be 

terminated prior to completion. 

A

(Southwest)

Meters Meters

(Northeast)

A'

5 5

5 5

10 10

15 15

20 20

VERTICAL EXAGGERATION = 10X

EXPLANATION

Sea

Level

Sea

Level

MW-23

Land

DP-14 DP-15

surface DP-16 DP-17 DP-18

MW-37

Red Lion Creek

?

?

?

?

?

?

?

? ?

?

?

Line of equal head, in meters

(Contour interval 0.1 meters;

dashed where approximate)

0.5 Generalized groundwater-flow direction

Clay

Gravel

Geologic contact (dashed and

queried where approximate)

Monitoring well (MW) and drive-point

piezometer (DP) locations, screened

interval, and identifier

Clay/sand

Clay/silt

Organics

Organics/clay

Sand

Sand/clay

Sand/gravel

Silt

Silt/clay

MW-23

?

0

0

20 40 60 80 METERS

100 200 FEET

0

0.1

0.2

0.3

0.4

0.5

Figure 11. Head distributions and groundwater-flow direction in section A–A’, August 8, 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware.
Hydrogeologic Characterization 33

Rates of lateral groundwater discharge through the 

Columbia aquifer and wetland sediments along transects 

A–A′ and B–B′ at the SCD also were estimated using slug 

test derived hydraulic conductivity and gradients obtained 

from groundwater head maps (table 9). The maximum lateral 

groundwater velocity in the Columbia aquifer was 151 m/yr 

toward Red Lion Creek along transect A–A’, whereas the highest rate in the wetland sediments was an order of magnitude 

less. Maximum lateral linear velocities of 3.9 m/yr to 

13.8 m/yr were estimated for the wetland sediments, with 

higher rates along transect B–B′ than A–A′, which was 

expected based on the spatial variation in hydraulic gradient. 

Groundwater velocity in the Columbia aquifer, however, was 

greater along A–A′ than B–B′. This variation in velocity is 

mostly due to an elevated hydraulic conductivity measured 

at DP-15B (on transect A–A′) compared to other estimates of 

hydraulic conductivity in the Columbia aquifer (table 7), and 

was not caused by a spatial difference in gradient. 

When comparing lateral and vertical linear groundwater 

velocity, flow in the wetland is predominantly vertical along 

Figure 12. Head distributions and groundwater-flow direction in section B–B’, August 8, 2012, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware.

? ? ?

?

?

?

?

?

?

?

?

?

?

?

?

?

?

?

B

(South)

Meters

(North)

B'

4

2 2

2 2

4 4

6 6

8 8

Sea

Level

Sea

Level

VERTICAL EXAGGERATION = 7X

EXPLANATION

Meters

4

MW-22

DP-10

DP-11 DP-13

Land surface

Red Lion Creek

Clay Geologic contact (dashed and

queried where approximate) Clay/sand

Clay/silt

Organics

Organics/clay

Organics/silt

Sand

Sand/clay

Sand/silt

Silt/clay

MW-22 Monitoring well (MW) and drive-point

piezometer (DP) locations, screened

interval, and identifier

?

Clay/organics

Line of equal head, in meters

(Contour interval 0.1 meters;

dashed where approximate)

0.5 Generalized groundwater-flow direction

0

0 60 FEET

5 10 15 20 25 METERS

20 40

0

0.1

0.2

0.3

0.5

0.4

0.7

0.6

0.8
34 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Table 7. Maximum and minimum estimates of hydraulic conductivity calculated from slug tests conducted at selected drive-point 

piezometers in the Columbia aquifer and wetland sediments, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

[K, hydraulic conductivity; NM, not measured; ft/d, feet per day; m/d, meters per day; max, maximum; min, minimum]

Site 

identification 

number 

Columbia aquifer Wetland sediments

Drive-point 

identification 

number

K 

(ft/d)

K 

(m/d)

Drive-point 

identification 

number

K 

(ft/d)

K 

(m/d)

DP-11 Max DP-11B 7.80 2.38 DP-11A 5.30 1.62

DP-11 Min DP-11B 5.50 1.68 DP-11A 4.20 1.28

DP-12 Max DP-12B 7.30 2.23 DP-12AA 1.40 0.43

DP-12 Min DP-12B 4.40 1.34 DP-12AA 1.00 0.30

EM/DP-6 Max EM-6D 14.00 4.27 DP-6A 1.50 0.46

EM/DP-6 Min EM-6D 10.00 3.05 DP-6A 0.83 0.25

DP-15 Max DP-15B 68.00 20.73 DP-15A NM NM

DP-15 Min DP-15B 40.00 12.19 DP-15A NM NM

Max measured–

Columbia 68.00 20.73 Max measured–wetland 5.30 1.62

Min measured–

Columbia 4.40 1.34 Min measured–wetland 0.88 0.27

Table 8. Variation in vertical specific discharge and linear velocity of groundwater from the Columbia aquifer through wetland 

sediments, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

[Hydraulic gradients calculated from water levels collected in June and September 2011 (except for DP-12 and DP-15), February and August 2012; maximum 

and minimum calculated hydraulic conductivities used; effective porosity of 0.3 used for linear velocity calculation; m/yr, meters per year]

Location 

Maximum 

specific 

discharge 

range 

(m/yr)

Arithmetic 

mean 

specific 

discharge 

(m/yr)

Geometric 

mean 

specific 

discharge 

(m/yr)

Median 

specific 

discharge 

(m/yr)

Maximum 

linear 

velocity 

range 

(m/yr)

Arithmetic 

mean linear 

velocity 

(m/yr)

Geometric 

mean linear 

velocity 

(m/yr)

Median 

linear 

velocity 

(m/yr)

DP-6 15–43 27.2 18.6 25.0 48–143 90.7 86.0 83.2

DP-11 36–78 57.8 56.2 56.5 121–261 193 187 188

DP-12 27–62 41.4 39.9 40.9 91–205 138 134 137

DP-15 12–118 55.9 39.7 42.5 39–393 186 132 142

Table 9. Spatial variation in lateral groundwater specific discharge and linear velocity of groundwater at Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware.

[Calculated using groundwater head gradients measured on August 8, 2012 and maximum, minimum, and mean values of hydraulic conductivity from DP-15B 

and EM-6D for the Columbia aquifer along transect A–Aꞌ, DP-11B and DP-12B for the Columbia aquifer along transect B–Bꞌ, DP-6A for wetland sediments 

along transect B–Bꞌ, and DP-11A and DP-12AA for wetland sediments along transect B–Bꞌ; m/yr, meters per year]

Hydrologic unit Transect

Maximum 

specific 

discharge 

(m/yr)

Minimum 

specific 

discharge 

(m/yr)

Mean 

specific 

discharge 

(m/yr)

Maximum 

linear 

velocity 

(m/yr)

Minimum 

linear 

velocity 

(m/yr)

Mean 

linear 

velocity 

(m/yr)

Columbia 

aquifer

A–Aꞌ 45 7 20 151 22 67

B–Bꞌ 6 3 4 20 11 15

Wetland 

sediments

A–Aꞌ 1.2 0.6 1.0 3.9 2.1 3.2

B–Bꞌ 4.1 0.8 2.3 13.8 2.6 7.8
Hydrogeologic Characterization 35

both transects A–A′ and B–B′ (tables 8 and 9). Maximum 

vertical groundwater velocity through the Columbia aquifer 

and wetland sediments along transect B–B′ was 261 m/yr at 

DP-11 (table 8), whereas the highest lateral velocity along 

this transect was within the Columbia aquifer (20 m/yr, table 

9). Along transect A–A′, maximum vertical groundwater 

velocity was 393 m/yr (at DP-15), whereas lateral groundwater velocities were less than half the vertical estimate, with a 

maximum of 151 m/yr in the Columbia aquifer, and only 

3.9 m/yr in the wetland sediments. Although maximum rates 

of lateral groundwater movement along A–A′ may be biased 

slightly high due to the lack of hydraulic conductivity data 

from DP-15A, estimates of vertical flow are generally biased 

low because they are calculated between the higher conductivity Columbia aquifer and the lower conductivity wetland 

sediments. Overall, comparisons of lateral to vertical rates of 

groundwater velocity show that vertical discharge predominates in the wetland near Red Lion Creek (tables 8 and 9). 

Effect of Tidal Fluctuation in Red Lion Creek

Tidal fluctuation along Red Lion Creek has been minimized since the installation of a tide control structure downstream of the SCD, when it has operated as designed. Previous 

investigations indicated that with the tide gate open, Red Lion 

Creek exhibited a tidal range of 1.29 m, whereas the Delaware 

River showed fluctuations of 0.87 m (Black & Veatch, 2005). 

Hydrographs from Red Lion Creek and drive-point piezometers in the wetland study area indicate that the tide gate may 

have been damaged during Hurricane Irene and Tropical 

Storm Lee in late August and early September of 2011, resulting in consistent inundation of the wetland study area at the 

SCD and increased tidal influence upon both creek stage and 

groundwater altitude at the SCD after these storm events. 

Evaluation of the continuous creek stage data from Red Lion 

Creek between August 28, 2008 and August 28, 2012 showed 

that the average creek stage increased from an elevation of 

-0.27 m above mean sea level to -0.09 m above mean sea level 

following Hurricane Irene (fig. 10). Continuous water-level 

data collected at the gage on Red Lion Creek and at several 

drive-point piezometers installed in the wetland study area at 

the SCD prior to the major storms during fall 2011 showed 

little to no indication of tidal influence (data on file at the 

MD-DE-DC WSC), whereas hydrographs following these 

storms show a distinctly tidal signature at Red Lion Creek in 

both drive-point piezometers screened in the Columbia aquifer 

and in some wells screened in shallow wetland sediments 

(fig. 13). In both the Columbia aquifer and wetland sediments, 

the magnitude of the tidal signature varies with proximity to 

Red Lion Creek and the material in which the given piezometer is screened. Response to tidal fluctuation in the Columbia 

aquifer is the least evident at EM-2D, which is positioned on 

the southeast edge of the upland part of the wetland study area, 

and is the farthest from Red Lion Creek of the drive-point 

piezometers instrumented for continuous water-level data collection (fig. 13). Drive-point piezometers located closer to the 

creek, such as DP-11B, DP-13B, DP-15B and DP-17B more 

closely reflect the fluctuation in elevation of Red Lion Creek 

than those located closer to the upland part of the study area, 

such as DP-10B and DP-14B (fig 13). In the wetland sediments, no response to tidal fluctuation was seen at DP-10AA, 

DP-15AA or DP-16AA, but DP-12AA, which was screened 

in shallow wetland sediments, and DP-16B, which is screened 

in much deeper wetland sediments, showed fluctuations which 

very closely mimic that of the creek (all except DP-12AA 

shown in fig. 13). 

A potential combined effect of the increase in tidal influence on Red Lion Creek and variation in hydraulic properties 

of the material underlying the creek is the cyclical reversal 

of head gradients, which would result in fluctuation between 

groundwater recharge and discharge at points in the wetland 

along Red Lion Creek. This fluctuating condition would 

occur if the range of head fluctuation in a more hydraulically 

conductive unit was large enough to overlap the range of a 

less hydraulically conductive unit, changing which well has 

higher hydraulic head throughout the course of a tidal cycle. 

Paired drive points (one screened in the Columbia aquifer and 

one in the wetland sediments) at three locations were instrumented for continuous water-level data collection in order to 

8/6/12

12:00

8/7/12

0:00

8/7/12

12:00

8/8/12

0:00

8/8/12

12:00

8/9/12

0:00

8/9/12

12:00

8/9/12

0:00

8/9/12

12:00

0.8

0.6

0.4

0.2

0.0

0.2

0.4

Date and time

DP-10AA

DP-15AA

EM-2D

DP-10B

DP-13B

DP-14B

DP-15B

DP-16AA

Red Lion Creek DP-16B

Red Lion

Creek

EM-2D

DP-10B

DP-14B

DP-10AA

DP-15AA

DP-16AA

DP-16B

DP-15B

DP-13B

COLUMBIA

AQUIFER

SHALLOW

WETLAND

SEDIMENTS

DEEP

WETLAND

SEDIMENTS

EXPLANATION

Water-level altitude,

in meters relative to mean sea level

Figure 13. Water-level altitudes at selected wetland-aquifer 

pairs and other drive-point piezometers screened in the Columbia 

aquifer, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware.
36 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

determine whether tidal influence was significant enough to 

reverse head gradients and subsequent recharge and discharge 

conditions. Hydrographs from these pairs showed little to 

no response to tidal fluctuation at points screened in shallow 

wetland sediments (DP-10AA, DP-15AA, and DP-16AA) 

but showed a distinct tidal signature at points screened in the 

Columbia aquifer (DP-10B and DP-15B) and the deepest 

wetland sediments (DP-16B) (fig.13). The range of fluctuation 

in water levels in the more responsive wells, however, is not 

large enough to cause hydraulic head in the wetland sediments 

to be higher than those observed in the Columbia aquifer, and 

thus, despite tidal influence, gradients remain vertically up 

throughout the tidal cycle (fig. 13). This indicates that groundwater discharge to Red Lion Creek is not reversed despite tidal 

influence on Red Lion Creek and groundwater altitudes in the 

wetland study area. Consistent vertically upward discharge 

of groundwater (or, flux, described in the following section) 

from the wetland sediments to Red Lion Creek contributes to 

the transport of contaminants from the groundwater system to 

surface water at the SCD. 

Groundwater Discharge to Red Lion Creek

In order to assess the potential significance of groundwater discharge to tidal Red Lion Creek, a comparison was 

made between vertical groundwater flux estimates (see previous section) integrated over the area of the tidal creek and an 

estimate of total discharge from the watershed draining to tidal 

Red Lion Creek (fig. 14). The goal was to determine, within 

reasonable limits given the absence of a detailed water budget, 

the percentage of the net recharge that may be discharging to 

the tidal creek and wetland.

During this study, no streamgage was active in the watershed that could provide data on discharge, though stage data 

were collected at River Road Road bridge on Red Lion Creek 

near the wetland study area (data available online for stage 

gage 01482320 at http://waterdata.usgs.gov/de/nwis/sw)

(fig. 14). A previously active streamgage (station 01482298, 

Red Lion Creek near Red Lion, Delaware), upstream of the 

study area (fig. 14), provided measurements of annual mean 

discharge for water years 1979–81 of 5.27, 3.71, and 

1.85 cubic feet per second (ft3/s). The average annual discharge for those years, 3.61 ft3

/s (3,220,000 cubic meters per 

year [m3/yr]), was used in this analysis, although the data do 

indicate substantial variation from year to year. The annual 

mean discharge measured at the gage represents the sum of 

surface and groundwater runoff, although groundwater runoff 

(base flow) is likely to be the biggest contributor in the relatively flat and permeable materials of the surficial aquifer in a 

Coastal Plain setting (Bachman and others, 1998). Assuming 

no change in storage within the watershed on an annual basis, 

annual mean discharge should be balanced by net recharge 

(precipitation minus evapotranspiration). Therefore, annual 

mean discharge represents an upper limit on recharge to the 

groundwater system in a watershed.

The drainage-area ratio method was used to estimate 

annual mean discharge for the ungaged subwatershed draining 

to tidal Red Lion Creek, based on the average annual mean 

discharge at the Red Lion Creek streamgage. The drainagearea ratio method is a technique that statistically transfers 

streamflow information from one location to another on the 

basis of the drainage areas of the two locations. The method is 

algebraically simple and is fairly straightforward to implement 

(Asquith and others, 2006). The method is expressed mathematically as:

Y

X

A Y X

A



  =    

(7)

where

Y = the streamflow for the ungaged location;

X = the streamflow at a streamflow-gaging 

station;

AY = the drainage area for the ungaged location; 

and

AX = the drainage area for the streamflow-gaging 

station.

Although the exponent α may take on a range of values 

depending on the particular flow duration (seasonal, monthly, 

storm peaks) being estimated (Doheny and Fisher, 2000; 

Emerson and others, 2005), its value is approximately equal to 

1 at an annual average time scale, and therefore the drainagearea ratio method is a direct proration of flow based on streamflow per unit area.

In order to apply the method, the watershed above the 

streamgage, as well as the watersheds above tidal Red Lion 

Creek (point A in fig. 14) and the River Road bridge below 

tidal Red Lion Creek (point B in fig. 14), were delineated 

using 30-m Digital Elevation Model (DEM) data. The contributing area between points A and B is referred to as the subwatershed draining to tidal Red Lion Creek. The areas for these 

watersheds are listed in table 10. Applying the drainage-area 

ratio method based on the average mean annual discharge at 

the gage and the areas determined for the different watersheds 

results in estimates of annual discharge for each of the watersheds (table 10). The estimate for the subwatershed draining 

to tidal Red Lion Creek was 795,000 m3/yr, and represents the 

expected increase in discharge between points A and B, and an 

upper limit on potential groundwater discharge to the wetland.

The groundwater discharge to tidal Red Lion Creek was 

estimated using calculated vertical specific discharge values 

from the nested piezometer sites with hydraulic conductivity 

values and the area of the tidal creek estimated from aerial 

photography (table 10). Using a minimum value of 12 m/yr 

for the range in vertical specific discharge (table 8) results in a 

calculated groundwater discharge to the tidal creek of 

372,000 m3/yr, or 47 percent of the estimated total discharge 

for the subwatershed. Using a minimum value of 18.6 m/yr 

for the geometric mean vertical specific discharge (table 8) 

results in a calculated groundwater discharge to the tidal creek 
Hydrogeologic Characterization 37

of 576,600 m3/yr, or about 72 percent of the estimated total 

discharge for the subwatershed. Minimum values for specific 

discharge were chosen partly to reflect the possible bias in 

hydraulic conductivity values based on slug tests that were 

more likely to be successful in more permeable locations, and 

partly to provide conservative estimates. 

These estimates, with some limitations, indicate that at 

least half of the net recharge to the subwatershed of tidal Red 

Lion Creek is discharging as groundwater to the tidal creek 

and wetland. The estimates of 47 to 72 percent represent lower 

limits, as minimum vertical specific discharge ranges and 

means were chosen for the tidal creek. These estimates are 

similar, however, to base-flow estimates recently reported for 

Coastal Plain streams in Maryland, Delaware, and Virginia 

using chemical hydrograph separation techniques (Sanford 

and others, 2012). Base-flow estimates by Sanford and others 

(2012) ranged from more than 50 percent to more than 

90 percent of measured streamflow.

The calculated volumetric groundwater fluxes (372,000 

to 576,600 m3/yr) to Red Lion Creek were based on the area 

of the tidal creek only and do not include the surrounding 

wetland that is likely receiving groundwater discharge (based 

on observed vertical hydraulic head gradients). Using a larger 

discharge area or higher value of vertical specific discharge 

would result in an even greater calculated groundwater contribution to the wetland. The large estimated volume of groundwater discharge to Red Lion Creek indicates that groundwater transport of contaminants to Red Lion Creek could be 

significant. 

75°44'00" 75°43'00" 75°42'00" 75°41'00" 75°40'00" 75°39'00" 75°38'00" 75°37'00"

39°38'00"

39°37'00"

39°36'00"

39°35'00"

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

0 1

0 .5

2 KILOMETERS

1 MILE

Red Lion Creek

Delaware

River

Delaware

River

Lea Road Governor

Wrangle

Hill

Road

9

9

71

71

River

Road

River

Road

NN

13

13

40

40

01482298 Stage gage

01482320

End of tidal

Red Lion

Creek

River Road

bridge

A B

72

72

72

Red Lion Road

Pulaski

Pulaski

Highway

Dupont

Dupont

Highway

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

EXPLANATION

Red Lion Creek wetland tidal area

U.S. Geological Survey streamgage and identifier

Watershed boundary above Red Lion Creek gage

Watershed boundary above stage gage 01482320

Watershed boundary above River Road bridge

Standard Chlorine of Delaware, Inc. barrier wall

01482298

Figure 14. Location of Red Lion Creek watershed, U.S. Geological Survey streamgage 01482298 (Red Lion Creek near Red Lion, 

Delaware), U.S. Geological Survey stage gage 01482320 at River Road Bridge, and the watersheds used to estimate discharge to 

tidal Red Lion Creek.
38 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Geochemical Characterization

Geochemical characterization of the groundwater in the 

wetland study area, in conjunction with an understanding of 

the hydrogeology, is critical to evaluate possible contaminant 

source locations and the potential effectiveness of natural 

attenuation and other remedial alternatives. Distributions 

and types of VOCs and redox-sensitive constituents (nitrate, 

ammonia, sulfate, sulfide, iron, and methane) present in 

groundwater in the wetland sediments in the study area are 

discussed here for samples collected between July 2009 and 

August 2012. Other geochemical constituents, including chloride and NVDOC, that help characterize contaminant sources 

and natural attenuation also are presented. Spatial distribution 

of geochemical constituents is described aerially using PDB 

results and vertically using piezometer and peeper results from 

synoptic sampling events; temporal changes in the wetland 

geochemistry are assessed using results from monthly sampling of selected sites from July 2011–August 2012, combined 

with results from earlier samples collected from these sites. 

Distributions of geochemical constituents are detailed in this 

section, but their implications for contaminant sources and 

natural attenuation, respectively, are discussed further in later 

sections on DNAPL extent and assessment of biodegradation. 

Distribution of Volatile Organic Compounds

The major VOCs detected in groundwater in the wetland 

study area were consistent with those reported previously for 

aquifer contamination in the upland area at SCD and for sediment samples from the wetland area, including benzene and 

six chlorobenzene compounds—123TCB, 124TCB, 12DCB, 

13DCB, 14DCB; and CB (Black & Veatch, 2007). The sum 

of these major VOCs are collectively referred to as total 

chlorobenzene compounds and benzene, or CBs+benzene, in 

this report. Other VOCs detected in the wetland groundwater 

relatively frequently (more than 5 samples of the 60 PDB and 

piezometer samples collected in the October-December 2011 

synoptic event) included chloroform, toluene, and 1,2,3-trichloropropane (data on file at USGS MD-DE-DC WSC, 

Baltimore, Maryland). These constituents were detected only 

at sites where chlorinated benzenes and benzene also were 

detected and are not discussed in this report.

Spatial Distribution

The distributions of CBs+benzene contamination varied 

widely throughout the wetland study area south of Red Lion 

Creek, as shown by results from PDBs deployed at a depth of 

approximately 20 to 40 cm below ground surface in March 

2011 and October–December 2011 (fig. 15). Based on differences seen in both geochemical and hydrogeologic observations, results were separated into two geographic areas of the 

wetlands, northwest and northeast. The line between the two 

areas is shown in figure 15 and corresponds approximately 

with the flow divide observed in the wetland groundwater 

(fig. 8). Two sites, DP-17 and DP-18, in the wetland area on 

the northern bank of Red Lion Creek (table 2, fig. 3) had low 

concentrations of CBs+benzene, with maximum concentrations in the shallow wetland groundwater and Columbia 

aquifer, respectively, of 50 and 20 μg/L. There were no detections of CBs+benzene in MW-37, which is screened in the 

Columbia aquifer and is the site farthest north of Red Lion 

Creek that was sampled (fig. 3). Results for sites north of Red 

Lion Creek are not included in the discussion below; additional investigation would be needed to determine the source 

of observed low concentrations (such as groundwater transport 

from sources south of the creek or from contaminated sediments in the creek channel or along the northern bank).

 CBs+benzene concentrations ranged from 36 μg/L at 

site 125 in the northeast wetland area to over 64,000 μg/L at 

site 141 in the northwest area for the March 2011 sampling 

event. CBs+benzene concentration ranges were similar in 

the larger October–December 2011 synoptic sampling event 

Table 10. Area and measured or estimated annual mean discharge for subwatersheds in the Red Lion Creek watershed, and 

estimated groundwater discharge to tidal Red Lion Creek.

[km2, square kilometers; m/yr, meters per year; m3/yr, cubic meters per year]

Area 

(km2)

Discharge 

(m/yr)

Discharge 

(m3/yr)

Red Lion Creek, mean measured discharge, 1979–81 8.13 0.396 3,220,000

Watershed above tidal Red Lion Creek, estimated 21.06 0.396 8,350,000

Watershed above River Road bridge, estimated 23.06 0.396 9,150,000

Subwatershed including tidal Red Lion Creek, estimated 2.01 0.396 795,000

Tidal Red Lion Creek, based on minimum value for 

maximum estimated vertical specific discharge 0.031 12.0 372,000

Tidal Red Lion Creek, based on minimum geometric 

mean estimated vertical specific discharge 0.031 18.6 576,000
Geochemical Characterization 39

compared to March 2011, with a minimum of 1 μg/L at site 

151 in the northeast wetland area to over 44,000 μg/L at site 

8 in the northwest wetland area. In general, the northeast 

wetland area had lower concentrations of CBs+benzene than 

those observed in the northwest wetland area for both of the 

synoptic sampling events. 

The higher concentrations of CBs+benzene in the 

northwest area of the wetland than those in the northeast 

wetland are consistent with the flowpath of the major release, 

400,000 gallons of dichlorobenzenes and 169,000 gallons of 

trichlorobenzenes, documented in 1986 (U.S. Environmental 

Protection Agency, 1995a). During that release, the majority of 

chemicals flowed from the center of the site northwest to the 

unnamed tributary, which flows into Red Lion Creek, resulting 

in DNAPL contamination in the wetland sediment (see section 

on DNAPL extent). Although the majority of sites in the northeast part of the wetland had much lower concentrations, there 

were localized areas of moderate contamination, mainly along 

the boundary between the wetland and upland. 

The wetland sediment lithology may also affect the 

distribution of contamination in the wetland study area. 

Shallow clay lenses are prevalent in the northwest area of the 

wetland, compared to more organic-rich, marsh-like sediments 

in the northeast area (figs. 6 and 7). The shallow clay layers 

may cause areas of more concentrated groundwater flow to 

the surface in the northwest wetland area, yielding higher 

concentrations of CBs+benzene. However, concentrations of 

CBs+benzene were consistently less than 10 μg/L at sites 145 

through 157, which are the sites farthest from the upland and 

from the path of the1986 major release (fig. 15b). Many of the 

sampling locations in the northeast wetland are on a thick matlike upper wetland sediment unit, rich in organic matter and 

roots from Phragmites, with a lithology thought to be similar 

to that observed at site 16 (fig. 7). Diffuse groundwater flow 

through this mat-like wetland sediment may result in lower 

75°38'00"

39°36'20"

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

N

Haul Road

Red Lion Creek

0 50

150

100 METERS

0 300 FEET

A

143

142

141

140

139

133

132 136 131

138

11

135

12 129 128

134

104 116 112

107

6

123

125

8

EXPLANATION

VOLATILE ORGANIC COMPOUNDS,

in micrograms per liter

0-100

100-1,000

1,000-5,000

5,000-10,000 (no data)

10,000-65,000

Wetland dividing line

125

Figure 15. Distribution of volatile organic compounds from passive diffusion bags deployed in the wetland sediments in the study 

area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, (A) March 2011, and (B) October–December 2011.
40 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

contaminant concentrations because of dilution and longer 

residence time for physical, chemical, and biological contaminant attenuation processes. 

The October–December 2011 PDB sampling results 

indicated that the concentrations of CBs+benzene were 

significantly (p<0.001, Mann-Whitney Rank Sum Test, NonParametric) greater in the northwest wetland area than in the 

northeast wetland area, with median concentrations of 

16,600 μg/L and 114 μg/L, respectively (figs. 16c, d). 

However, of the individual compounds measured, CB and 

TCB isomers were the only compounds to have a significantly 

(p<0.01) different molar percentage between the northwest 

and northeast wetland areas in the October 2011 PDBs 

(fig. 16). Note that molar compositions are shown in figure 

16 for all samples with total CBs+benzene that had concentration greater than 10 μg/L; thus, sites 145–157, sampled 

in December 2011, are not shown. CB was the predominant 

VOC in PDB samples from the northeast wetland area with 

a median of 69 molar percent, compared to a median of 35 

molar percent in samples from the northwest wetland area. 

Concentrations of the TCB isomers were relatively low compared to the other chlorinated benzenes, as might be expected 

from their lower solubility (table 1), but the median molar percentage of 0.33 percent was greater in the northwest wetland 

area than in the northeast wetland area (median 0 percent). The 

sum of the three DCB isomers were the predominant VOCs 

in most samples from the northwest wetland area in October 

2011 (figs. 16c, d), with a median molar percentage of 50 

percent compared to 24 percent in samples from the northeast 

wetland area, although the difference was not statistically 

significant (p = 0.178). Statistics were not calculated for the 

March 2011 synoptic sampling event (figs. 16a, b) because 

of the limited number of samples collected in the northeast 

wetland area.

75°38'00"

39°36'20"

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

N

Red Lion Creek

Haul Road

143

142

141

140

8

102

144

139

133

132 131

138

11

135

134

128 122

123

15 125

126

127

145

147

148

146 151

152

153

154

155

156 157

150

149

120

119

118

117

116 112

113

114

106

14

107 108

109

110

6

EXPLANATION

VOLATILE ORGANIC COMPOUNDS,

in micrograms per liter

0-100

100-1,000

1,000-5,000

5,000-10,000

10,000-50,000

Wetland dividing line

125

0 50

150

100 METERS

0 300 FEET

B

Figure 15. Distribution of volatile organic compounds from passive diffusion bags deployed in the wetland sediments in the study 

area, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, (A) March 2011, and (B) October–December 2011.

—Continued
Geochemical Characterization 41

Fine-scale vertical delineation of VOCs from peeper 

sampling in 2009 and 2011 showed that concentrations of 

CBs+benzene were generally highest in the upper 40–50 cm 

of wetland sediment in both the northwest and northeast 

areas (fig. 17). Two exceptions to this pattern occurred in 

the peepers at sites 9 and 10 in the northwest wetland area, 

where CBs+benzene concentrations were nearly uniform 

with depth. In peepers in the northwest area, concentrations 

of CBs+benzene were greater than 10,000 μg/L at practically 

all depths sampled and reached a maximum concentration 

of about 75,000 μg/L (figs. 17a, c). As observed in the PDB 

samples, generally lower concentrations of VOCs occurred in 

the peepers in the northeast area compared to the northwest 

area, although concentrations were higher than 10,000 μg/L 

at sites 7 and 15 in the upper 50 cm. For peeper sites where 

deeper piezometer data were also available, the vertical 

distributions showed that VOC concentrations at greater depth 

in the peepers (more than 50 cm deep) were lower than in the 

underlying groundwater sampled from piezometers, except 

at site 14 (figs. 17b-d). These vertical distributions indicate 

that another source of VOCs is present in the shallow wetland 

sediments in addition to upward groundwater discharge from 

the contaminant plume in the deeper wetland sediments and 

underlying aquifer. The individual VOCs that make up the 

Figure 16. Molar composition of volatile organic compounds (VOCs) and total concentrations of chlorobenzene compounds and 

benzene (CBs+benzene) in passive diffusion bag samples collected from the northwest and northeast wetland areas during 

(A, B) March 2011, and (C, D) October–December 2011, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

Benzene

Chlorobenzene (CB)

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total CBs+benzene

-5,000

5,000

15,000

25,000

35,000

45,000

55,000

65,000

0

0 0

0

10

10 10

10

20

20 20

20

30

30 30

30

40

40 40

40

50

50 50

50

60

60 60

60

70

70 70

70

80

80 80

80

90

90 90

90

100

100 100

100

8

104

128

129

12

131

132

133

134

135

136

11

138

139

140

141

142

143

144

Total concentration, in micrograms per liter Total concentration, in micrograms per liter

Total concentration, in micrograms per liter Total concentration, in micrograms per liter

Molar percent Molar percent

Molar percent

Sampler number Sampler number

Sampler number

Sampler number

Molar percent

0

2,000

4,000

6,000

8,000

10,000

12,000

14,000

16,000

18,000

20,000

107 6 7 112 116 123 125

-5,000

5,000

15,000

25,000

35,000

45,000

55,000

65,000

8 102 131 132 133 134 135 11 138 139 140 141 142 143 144

0

5,000

10,000

15,000

20,000

25,000

30,000

106

107

108

109

110

6

112

113

114

14

116

117

118

119

120

122

123

15

125

126

127

128

EXPLANATION

A. March 2011, Northwest VOCs

C. October–December 2011, Northwest VOCs D. October–December 2011, Northeast VOCs

B. March 2011, Northeast VOCs
42 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

total CBs+benzene observed in the wetland groundwater also 

differed between the northwest and northeast wetland areas, 

although these VOC compositions varied temporally and are 

discussed in the next section on Temporal Distribution.

Temporal Distribution

Sites 8 and 6 that were used for peepers, PDB synoptics, 

and monthly sampling provided the best temporal data from 

the wetland groundwater (figs. 18 and 19). The molar composition of the VOCs varied greatly between the July 2009 

and December 2011 peeper sampling events at site DP8 in the 

northwest area (figs. 18a, b). In the July 2009 peeper at site 

8, CB and benzene comprised over 90 percent of the VOC 

composition in the wetland groundwater along the entire 

approximately 120-cm-long profile. Peeper samples from 

nearby site 9 in July 2009 showed similar molar VOC compositions with about 80 percent comprised of CB and benzene 

(data on file, USGS MD-DE-DC WSC). In contrast, the VOC 

composition in peeper samples from site 8 in December 2011 

showed substantially lower percentages of CB and benzene 

and higher percentages of the DCB isomers (12DCB, 13DCB, 

and 14DCB) than those observed in 2009. Monthly sampling 

results in 2011–12 for DP8A, which is screened at site 8 

deeper in the wetland sediment (370–670 cm) than the zone 

sampled with the peepers, were consistent with the VOC composition observed in the December 2011 peeper and remained 

consistent throughout the year, although the total concentrations of CBs+benzene varied between about 12,000 and 

58,000 μg/L (fig. 18c). The trichlorobenzene isomers 123TCB 

and 124TCB made up an insignificant percentage of the high 

concentrations of total VOCs in the wetland groundwater at 

site 8 during all sampling periods (fig. 18). 

In the northeast area at site 6, the VOC composition 

in July 2009 differed greatly from the VOC composition 

observed at site 8, with a substantially higher percentage of 

12DCB and lower percentages of CB and benzene (figs. 18a 

and 19a). In 2011–12, the VOC compositions in wetland 

groundwater at sites 6 and 8 were similar, however, with the 

three DCB isomers generally making up 50 percent or more 

of the total VOCs (figs. 18b, c and 19b, c). The TCB isomers 

A. July 2009, Northwest Total VOCs

C. October–December 2011, Northwest Total VOCs D. October–December 2011, Northeast Total VOCs

B. July 2009, Northeast Total VOCs

0

0 0

0

20

20 20

20

40

40 40

40

60

60 60

60

80

80 80

Site 8

Site 8

Site 6

Site 9

Site 10

Site 12

Site 7

EM6D (318 centimeters,

14,500 micrograms per liter)

DP10B (315 centimeters,

14,300 micrograms per liter)

EM6D (318 centimeters,

1,270 micrograms per liter)

DP14B (381 centimeters,

5,460 micrograms per liter)

DP15B (400 centimeters,

10,600 micrograms per liter)

80

100

100 100

100

120 120

140

120 120

140

320

320 400

0

0

0

0

20,000

10,000

10,000

5,000

40,000

20,000

20,000

10,000

60,000

30,000

30,000

15,000 20,000

80,000

40,000

40,000

25,000

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Concentration, in micrograms per liter

Concentration, in micrograms per liter

Concentration, in micrograms per liter

Concentration, in micrograms per liter

Site 6

Site 14

Site 15

Figure 17. Vertical distributions of volatile organic compounds (VOCs) from peepers in the northwest and northeast wetland areas 

(A, B) July 2009, and (C, D) October–December 2011, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.
Geochemical Characterization 43

were a significant percentage of the total VOC concentrations 

observed at site 6 when total CBs+benzene concentrations 

were relatively low, such as in the deep peeper samples (about 

68 to 113 cm below land surface) in December 2011 (fig. 19b) 

and in the 2011–12 monthly samples (fig. 19c).

Because 2009–12 temporal data were only collected from 

a few sites in the wetland, it is difficult to determine if the 

changes at sites 8 and 6 are representative of changes across 

the northwest and northeast wetland areas. The synoptic PDB 

sampling in October 2011 indicates that the VOC composition observed at site 8 was consistent with most of the sites 

sampled in the northwest area (fig. 16c). In the northeast area, 

many of the sites had a predominantly higher percentage of 

CB than the percentage observed at site 6, although at least six 

other sites in October 2011 had a similar VOC composition 

to the one observed at site 6 (fig. 16d). The peeper at site 6 in 

December 2011 did have a relatively high percentage of CB 

and benzene in the upper 24 cm of wetland sediment 

(fig. 19b), indicating that variability with depth may account 

for some of the variability observed at different sites across the 

wetland area. 

Distribution of Redox-Sensitive and Other 

Geochemical Constituents

In addition to measuring VOCs, geochemical constituents were measured during the sampling events between 2009 

and 2012, although the suite of constituents varied with the 

amount of sample available from the sampling device (table 

3). Selected constituents are discussed here to describe the 

redox conditions of the groundwater and the general differences observed between the northwest and northeast wetland 

areas. Inorganic constituents, except methane, and NVDOC 

were measured from PDBs constructed of dialysis membrane 

and placed at the same location as the PDBs sampled for 

VOCs and methane. 

Spatial Distribution

The primary inorganic constituents in the wetland 

groundwater that showed distinct differences between the 

northwest and northeast wetland areas during all sampling 

0

10,000 5,000

10,000

15,000

20,000

25,000

30,000

35,000

40,000

20,000

30,000

40,000

50,000

60,000

70,000

80,000

0

10

20

30

40

50

60

70

80

90

100

2.0 2.0

7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1 118.1 7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1

13.1 24.1 13.1 24.1No data

35.2 46.2 57.3 68.4 79.4 90.5 101.5 112.6 35.2 46.2 57.3 68.4 79.4 90.5 101.5

Total CBs+benzene, in micrograms per liter

Total CBs+benzene, in micrograms per liter

Molar percent

Depth below land surface, in centimeters

Sampling date

Depth below land surface, in centimeters

A. July 2009, Peeper DP8

Benzene

Chlorobenzene (CB)

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total CBs+benzene

EXPLANATION

0

10

20

30

40

50

60

70

80

90

100

Molar percent

0

B. December 2011, Peeper DP8

0

10

20

30

40

50

60

70

80

90

100

Total CBs+benzene, in micrograms per liter

Molar percent

0

10,000

20,000

30,000

40,000

50,000

60,000

9/14/2011 12/8/2011 2/15/2012 4/26/2012

10/19/2011 1/17/2012 3/28/2012 8/9/2012

C. September 2011–August 2012, Piezometer DP8A

Figure 18. Temporal changes in molar composition of volatile 

organic compounds and total concentrations of chlorobenzene 

compounds and benzene (CBs+benzene) at site 8 in the northwest 

wetland area from (A) peeper sampling in July 2009, (B) peeper 

sampling in December 2011, and (C) monthly sampling events, 

September 2011–August 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware.
44 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

events were pH, chloride, sulfate and sulfide (figs. 20 and 21). 

The pH of the wetland groundwater measured in the northeast 

area was significantly higher (p = 0.043, March 2011; 

p = 0.003, October 2011) than the pH measured in the northwest area for both synoptic sampling events in 2011. Median 

pH values for the two synoptic sampling events were 5.8 and 

6.1 in the northeast wetland area and 4.5 and 5.2 in the northwest wetland area (fig. 20). The lowest pH (3.6) was measured 

at site 141 in the northwest wetland area, whereas the highest 

pH (6.8) was measured at site 127 in the northeast area.

In addition to chlorinated benzenes and benzene releases, 

the study area was affected by at least two acid spills (Black 

& Veatch, 2007). Hydrochloric acid (HCl) was part of the 

major release of 1986 and primarily affected the northwest 

wetland area, whereas a sulfuric acid spill from the oil 

refinery affected the groundwater flowing from the Columbia 

aquifer to the northeast wetland area. Despite both areas of 

the wetland experiencing strong acid spills, the lower median 

pH and higher concentrations of CBs+benzene in the northwest area indicate a more concentrated flow of contamination 

through the wetland sediments than in the northeast wetland 

area, where the median pH was higher and concentrations of 

CBs+benzene were lower. Reductive dechlorination of the 

chlorinated benzenes (fig. 2) in the more highly contaminated 

northwest wetland area also may have generated additional 

HCl and contributed to the lower pH compared to the pH in 

the northeast wetland area.

The northeast wetland groundwater had significantly 

higher concentrations of sulfate and sulfide (October 2011, 

p<0.01) than the concentrations measured in the northwest 

wetland area. During the October 2011 sampling event, the 

highest concentration of sulfate (1,090 mg/L) was found at site 

6 in the northeast wetland area, and the lowest concentration 

of sulfate (0.42 mg/L) was observed at site 135 in the northwest wetland area. The anion sulfate is the conjugate base of 

sulfuric acid, and sulfide can be produced from sulfate reduction in oxygen-depleted environments. Thus, the higher sulfate 

and sulfide in the northeast wetland groundwater compared to 

the northwest wetland groundwater provides evidence of the 

effect of the sulfuric acid spill. Because the sulfate reduction 

reaction utilizes hydrogen ions and thus can increase the pH 

of a solution, this reaction may be contributing to the higher 

0

500

1,000

1,500

2,000

2,500

1,000

2,000

3,000

4,000

5,000

0

10

20

30

40

50

60

70

80

90

100

2.0 2.0

7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1 7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1

13.1 24.1 13.1 24.1 112.6 No data

35.2 46.2 57.3 68.4 79.4 90.5 101.5 112.6 35.2 46.2 57.3 68.4 79.4 90.5 101.5

Total CBs+benzene, in micrograms per liter

Total CBs+benzene, in micrograms per liter

Molar percent

Depth below land surface, in centimeters

Sampling date

Depth below land surface, in centimeters

A. July 2009, Peeper DP6

Benzene

Chlorobenzene (CB)

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total CBs+benzene

EXPLANATION

0

10

20

30

40

50

60

70

80

90

100

Molar percent

0

B. December 2011, Peeper DP6

0

10

20

30

40

50

60

70

80

90

100

Total CBs+benzene, in micrograms per liter

Molar percent

0

200

400

600

800

1,000

10/19/2011 12/8/2011 2/16/2012 4/26/2012

11/15/2011 1/17/2012 3/28/2012 7/2/2012

8/9/2012

C. October 2011–August 2012, Piezometer DP6A

Figure 19. Temporal changes in molar composition of volatile 

organic compounds and total concentrations of chlorobenzene 

compounds and benzene (CBs+benzene) at site 6 in the northeast 

wetland area from (A) peeper sampling in July 2009, (B) peeper 

sampling in December 2011, and (C) monthly sampling events, 

October 2011–August 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware.
Geochemical Characterization 45

pH observed in the northeast wetland area compared to the 

northwest area. 

Concentrations of chloride were significantly higher 

(p = 0.003) in the northwest wetland area than in the northeast 

wetland area during the October 2011. Site 132 in the northwest wetland area had the highest measured concentration of 

chloride (171 mg/L), whereas site 110 in the northeast area 

had the lowest measured concentration of chloride (27 mg/L). 

The chloride anion is the conjugate base of HCl and could 

have resulted from the HCl spill. Another possible source of 

chloride at this site is from the reductive dechlorination of the 

chlorinated benzenes. 

Other redox-sensitive constituents, in addition to sulfate 

and sulfide, that were predominant in the wetland groundwater 

at many sites included ferrous iron and methane (fig. 21), but 

concentrations of these constituents were not statistically significant different between the northwest and northeast wetland 

areas in the March and October 2011 PDB samples. The presence of ferrous iron indicates moderately reducing conditions, 

whereas the presence of methane indicates highly reducing 

conditions. Wetland sediments are commonly anaerobic, and 

the distribution of these additional redox-sensitive constituents 

does not appear to be related to the groundwater contamination. Ammonia concentrations were typically low or undetectable in the wetland groundwater (fig. 21). During October 

2011, the highest concentration (19.1 mg/L) of aqueous phase 

methane was found at site 142 in the northwest wetland area. 

The lowest concentrations, under 1 mg/L, were observed at 

numerous sites in both the northwest and northeast wetland 

areas. Similar to the methane results in the March and October 

2011 samples, several localized areas within the wetland had 

relatively high concentrations of ferrous and total dissolved 

iron (greater than 5 mg/L), whereas the majority of areas had 

low concentrations (approximately 1 mg/L). On average, 

90 percent of the total dissolved iron measured was ferrous 

iron, which is the reduced form of iron. 

The distribution of NVDOC, measured during the 

October 2011 PDB sampling, was variable and not statistically 

significant based on geographic position (fig. 22). Because the 

analytical method for organic carbon excluded volatile carbon 

compounds, the presence of chlorinated benzenes and benzene contamination or other VOCs should not have affected 

the NVDOC distribution. Instead, the NVDOC distribution reflected the presence of natural organic carbon in the 

groundwater that can support microbial reduction of metals 

and sulfate in addition to the reductive dechlorination of the 

chlorinated benzenes. The northeast wetland area had an average NVDOC concentration of 56 mg/L, whereas the northwest wetland area had an average concentration of 14 mg/L. 

Both the highest (436 mg/L) and lowest (1.6 mg/L) NVDOC 

concentrations were found in the northeast wetland area (sites 

108 and 6, respectively). Site 108, which had the maximum 

NVDOC concentration, also had elevated concentrations of 

ferrous iron, sulfide, and methane compared to other wetland 

sites. 

Vertical distributions of chloride and sulfate in the July 

2009 peepers were consistent with the results of the 2011 

PDB and piezometer sampling, showing substantially higher 

concentrations of chloride and lower concentrations of sulfate 

in the wetland groundwater at all depths in the northwest area 

than those observed in the northeast area (figs. 23 and 24). In 

peepers sampled in October–December 2011, however, chloride concentrations in the peepers in both the northwest and 

northeast areas had similar maximum concentrations, ranging 

from 122–223 mg/L in the northwest area peepers and from 

Figure 20. Boxplots showing distributions of pH in the 

wetland study area (A) March 2011, and (B) October–

December 2011 synoptic sampling results, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware.

A. March 2011 passive diffusion bag (PDB) synoptic

B. October–December 2011 PDB synoptic

6.5

7.0

6.0

6.5

6.0

5.5

5.5

5.0

5.0

pH, in standard units pH, in standard units

4.5

4.5

4.0

4.0

3.5

3.0

Northwest

Northwest

Northeast

Northeast

EXPLANATION

Data value within 1.5 times the IQR

outside the quartile

MAXIMUM OUTLIER (greater than

1.5 times the IQR outside the quartile)

MINIMUM OUTLIER (less than 1.5 times

the IQR outside the quartile)

75th PERCENTILE (75 percent of the

data values are less than this value)

Median

25th PERCENTILE (25 percent of the

data values are less than this value)

INTERQUARTILE

RANGE (IQR)
46 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

150–180 mg/L in the northeast area peepers (figs. 23c, d). 

Minimum chloride concentrations were lower in the northeast 

peepers (40–50 mg/L) than those sampled in the northwest 

area (80–112 mg/L). Groundwater chloride concentrations 

from the only peeper in the northwest wetland where a deeper 

groundwater sample from a piezometer was also available (site 

10) were nearly uniform with depth and similar to the concentration in the underlying groundwater of the Columbia aquifer 

(fig. 23c). In the northeast wetland area, however, the chloride 

concentrations in all three peepers reached concentrations 

two to five times greater than those observed in the deeper 

groundwater from piezometer samples (fig. 23d). Maximum 

chloride concentrations in the peeper samples from both 

wetland areas were more than three times as high as chloride 

concentrations in the surface-water samples. Thus, vertical 

distributions of chloride indicate another source of chloride in 

the shallow wetland porewater besides surface-water infiltration or upward discharge of contaminated groundwater, as was 

also observed from vertical distributions of the VOCs (fig. 17). 

Surface-water infiltration may instead act to dilute shallow 

groundwater concentrations in the wetland at some sites where 

chloride concentrations in the peeper samples decrease near 

land surface. 

Sulfate concentrations at each peeper site in 2009 and 

2011 were relatively uniform with depth, except in the upper 

20 cm below land surface where the greatest change in sulfate 

concentrations occurred (fig. 24). In the northwestern peeper 

sites where sulfate was low, the highest sulfate concentrations 

0

50

100

150

200

250

300

350

0

0 0

0

5

10

5

5

5

10

10 15

10

15

20

15 25

15

20

30

25

20

20

35

30

25 40

25

8

131

132

133

134

135

136

11

138

139

140

141

142

143

128 144

129

12

104

Chloride and sulfate, in milligrams per liter

Chloride and sulfate, in milligrams per liter

Chloride and sulfate, in milligrams per liter

Chloride and sulfate, in milligrams per liter

Ferrous iron, sulfide, and methane,

in milligrams per liter

Ferrous iron, sulfide, and methane,

in milligrams per liter

Ferrous iron, sulfide, methane, and ammonia,

in milligrams per liter

Ferrous iron, sulfide, and methane,

in milligrams per liter

Sampler number Sampler number

Sampler number

Sampler number

0

200

400

600

800

1,000

1,200

107 6 7 112 116 123 125

0

80

60

40

20

100

120

140

160

180

8 102 131 132 133 134 135 11 138 139 140 141 142 143 144

0

200

400

600

800

1,000

1,200

106

107

108

109

110

6

112

113

114

14

116

117

118

119

120

122

123

15

125

126

127

EXPLANATION

A. March 2011, Northwest PDBs

C. October 2011, Northwest PDBs D. October 2011, Northeast PDBs

B. March 2011, Northeast PDBs

Methane

Sulfide

Ferrous iron

Ammonia

Chloride

Sulfate

Figure 21. Distributions of chloride, sulfate, ammonia, sulfide, methane, and ferrous iron in passive diffusion bag (PDB) samples of 

wetland groundwater collected in the nothwest and northeast areas in (A, B) March 2011, and (C, D) October 2011, Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware.
Geochemical Characterization 47

occurred in this upper 20-cm zone and were similar to surfacewater concentrations (available in 2011), indicating surfacewater infiltration. In contrast, wetland groundwater concentrations of sulfate were 10 to 50 times greater than surface-water 

concentrations in peepers in the northeast wetland area. The 

equal or higher sulfate concentrations at piezometers screened 

deeper than the peeper samples at the same sites indicate that 

groundwater discharge from the Columbia aquifer provides a 

source for the high sulfate. Upland monitoring wells MW22 

and MW23, screened in the Columbia aquifer near the wetland 

boundary in the northeast area, had higher sulfate concentrations than monitoring wells in the northwest area (MW21, 

MW25; table 11). In addition, previous samples collected from 

drive-point piezometers along the eastern wetland-upland 

boundary also showed high sulfate concentrations, ranging 

from 180 to 2,490 mg/L, that were attributed to a sulfuric 

acid spill at the oil refinery (table 12). Other redox-sensitive 

constituents in addition to sulfate showed temporal as well as 

spatial variations in the wetland groundwater and are discussed in the next section.

Temporal Distribution

Although limited data collection occurred during the 

early stages of this study, large changes in concentrations of 

some geochemical constituents were apparent in both areas 

of the wetland from July 2009 to March 2011, compared 

to relatively stable concentrations of constituents observed 

during monthly sampling from August 2011–August 2012 

(fig. 25). Chloride and sulfate concentrations in the shallow 

wetland groundwater at two sites—8 in the northwest wetland 

area and 6 in the northeast wetland area—that were consistently sampled during 2009–12 reflect these changes (fig. 25). 

To construct figure 25, concentrations in peeper samples were 

averaged over their 120-cm-depth range to compare with PDB 

samples (depths of 20 to 40 cm) before installation of DP-6A 

and DP-8A and the piezometer samples from DP-6A and 

DP-8A (screens at about 40 to 70 cm below land surface; 

table 2). Chloride concentrations at site 8 in the northwest 

decreased substantially from about 350 to 180 mg/L in July 

2009–March 2011, and then remained relatively low (between 

100–125 mg/L) during August 2011–12. Sulfate concentrations at site 6 increased from about 300 to 1,400 mg/L 

between July 2009 and December 2010 but then decreased 

again, returning to around 300–400 mg/L by the end of the 

2012 sampling period (fig. 25). 

Although sulfate concentrations showed greater temporal variability than chloride concentrations at these sites, 

sulfate concentrations were consistently a factor of 10 or more 

higher at site 6 in the northeast than observed at site 8 in the 

northwest (fig. 25). In contrast, chloride concentrations at the 

two sites were similar in August 2011–12 (fig. 25a), as noted 

previously in the peepers (figs. 23c, d). It is possible that the 

decline in chloride concentrations in the northwest wetland 

area observed between July 2009 and March 2011 is due to the 

installation of the barrier wall in 2007 that removed a source 

of chloride from groundwater flow from the Columbia aquifer 

to wetland sediments in the northwest area. Because the high 

sulfate source is believed to be from the sulfuric acid spill 

at the oil refinery, which is outside this barrier wall, sulfate 

concentrations in groundwater discharging to the northeast 

wetland area would not be affected by the barrier wall; thus, 

sulfate concentrations remained high in the northeast area, 

whereas chloride concentrations decreased in the northwest 

area.

Substantial changes in concentrations of redox-sensitive 

constituents occurred in the wetland porewater in both the 

northwest and northeast areas between 2009–12 (figs. 26 and 

27). Although a predominance of ferrous iron indicated that 

iron reduction was the predominant redox process at site 8 

in the northwest during each sampling period, ferrous iron 

concentrations were about a factor of 10 higher in the July 

2009 peeper samples than peeper samples from other sampling 

periods at this site (fig. 26). In addition, high methane concentrations at site 8 during July 2009, which peaked at nearly 

10 mg/L between depths of 62–79 cm below land surface, 

indicated mixed iron reducing-methanogenic conditions. In 

A. October 2011 Northwest PDBs

B. October 2011 Northeast PDBs

60

60

436

50

50

40

40

30

30

Nonvolatile dissolved organic carbon,

in milligrams per liter

Nonvolatile dissolved organic carbon,

in milligrams per liter

10

10

20

20

0

8

107 108 110 6 114 117 118 119 123 15 127

Site number

Site number

102 14 132 134 135 11 138 141 144

0

Figure 22. Spatial distribution of nonvolatile dissolved 

organic carbon (NVDOC) in passive diffusion bag (PDB) 

samples of the wetland groundwater collected in 

(A) northwest, and (B) northeast wetland areas in October 

2011, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware.
48 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

December 2011, methane concentrations were generally less 

than 0.040 mg/L at all depths in the peeper sampled at site 8, 

and in conjunction with the low ferrous iron concentrations, 

indicate a shift to less reducing conditions (fig. 26b). Although 

temperature decreases in the winter can lead to lower methanogenic activity, ferrous iron and methane concentrations also 

were low or undetectable during monthly sampling events 

from September 2011 to August 2012 at DP-8A (fig. 26c). 

Other northwest sites sampled with peepers in October 2011, 

sites 10 and 12, also showed low or nondetectable methane 

concentrations (data on file, USGS MD-DE-DC WSC). 

In contrast, the peeper samples from site 6 in the northeast had sulfide concentrations that were up to three to four 

times higher in December 2011 than in July 2009, indicating 

a shift to more reducing conditions (figs. 27a, b). Peeper 

samples at site 6 in December 2011 had the highest concentrations of reduced iron and sulfide in the upper 20 cm of wetland 

sediment, whereas the July 2009 samples showed only slightly 

reducing conditions in this upper zone (total concentrations 

of reduced species less than 1 mg/L). Other peepers sampled 

in the northeast wetland area in October 2011 (sites 14 and 

15) had high sulfide concentrations in the upper 60 cm, with 

maximum sulfide concentrations of 12 and 16 mg/L (data on 

file, USGS MD-DE-DC WSC). The sulfide concentrations in 

peeper samples from October-December 2011 were consistent 

with the significantly higher sulfide concentrations observed in 

PDB samples from the northeast wetland area in October 2011 

than those in the northwest area (fig. 21). Methane concentrations also were high (about 8 to 34 mg/L) in the upper 60 cm 

of sediment in the northeast peepers sampled in October 2011 

(data on file, USGS MD-DE-DC WSC). It should be noted 

however, that monthly samples from DP-6A showed low concentrations of reduced constituents in 2011–12. This apparent 

discrepancy may be due to the preferential sampling of deeper 

groundwater with the piezometers (screened 43–73 cm below 

land surface) where conditions were less reducing than at 

shallower depths in the wetland sediment, as indicated by the 

peeper profile at site 6 in December 2011 (fig. 27b). 

A. July 2009, Northwest Chloride

C. October–December 2011, Northwest Chloride D. October–December 2011, Northeast Chloride

B. July 2009, Northeast Chloride

0

0 0

0

20

20 20

20

40

40 40

40

60

60 60

60

80

80 80

Site 8

Site 10

Site 12

SW13, 17

DP10B (315 centimeters,

111 milligrams per liter)

EM6D (318 centimeters,

51 milligrams per liter)

DP14B (381 centimeters,

48.4 milligrams per liter)

DP15B (400 centimeters,

87 milligrams per liter)

80

100

100 100

100

120 120

140

120 120

140

320 400

0

0

0

0

100

100

100

100

200

200

200

200

300

300

300

300

400

400

400

400

500

500

500

500

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Concentration, in milligrams per liter

Concentration, in milligrams per liter

Concentration, in milligrams per liter

Concentration, in milligrams per liter

Site 8

Site 9

Site 6

Site 14

Site 15

Site 6

Site 7

SW13, 17

Figure 23. Vertical distributions of chloride concentrations in peepers from the northwest and northeast wetland areas in 

(A, B) July 2009, and (C, D) October–December 2011, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.
Geochemical Characterization 49

Figure 24. Vertical distributions of sulfate concentrations in peepers from the northwest and northeast wetland areas in (A, B) July 

2009, and (C, D) October–December 2011, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

A. July 2009, Northwest Sulfate

C. October–December 2011, Northwest Sulfate D. October–December 2011, Northeast Sulfate

B. July 2009, Northeast Sulfate

0

0 0

0

20

20 20

20

40

40 40

40

60

60 60

60

80

80 80

Site 8

Site 10

Site 12

SW13, 17

DP10B (315 centimeters,

1.75 milligrams per liter)

EM6D (318 centimeters,

650 milligrams per liter)

DP14B (381 centimeters,

680 milligrams per liter)

DP15B (400 centimeters,

860 milligrams per liter)

80

100

100 100

100

120 120

140

120 120

140

320 400

0

0

0

0

2

5

200

200

4

10

400

400

6

15

600

600

8

20

800

800

10

25

1,000

1,000

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Depth below land surface,

in centimeters

Concentration, in milligrams per liter

Concentration, in milligrams per liter

Concentration, in milligrams per liter

Concentration, in milligrams per liter

Site 8

Site 9 Site 6

Site 7

Site 6

Site 14

Site 15

SW13, 17

Table 11. Concentrations of selected inorganic constituents measured in groundwater samples collected with passive diffusion bags 

from upland monitoring wells, October 2011.

[cm, centimeters; mg/L, milligrams per liter; <, less than]

Well 

name Sample date Depth(cm)

pH 

(standard 

units) 

Chloride

(mg/L)

Sulfate, 

as SO4

(mg/L)

Ammonia, 

as N

(mg/L)

Ferrous iron, 

as Fe

(mg/L)

Total 

dissolved 

iron

(mg/L)

Sulfide, 

as S

(mg/L)

Methane

(mg/L)

MW-25 10/26/2011 736.2 5.05 255 0.25 0.35 5.39 5.53 0.018 <.027

MW-21 10/26/2011 404.6 5.80 83.9 35.2 0.34 5.42 5.53 0.006 <.027

MW-22 10/26/2011 349.8 5.11 76.8 320 0.10 0 0.06 <.005 0.073

MW-23 10/26/2011 626.7 6.92 36.3 122 0.37 15.9 15.59 <.005 <.027
50 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Dense Non-Aqueous Phase Liquid 

(DNAPL) Extent

The chlorinated hydrocarbon contaminants that spilled 

and were detected in the subsurface environment at the SCD 

can be characterized as DNAPLs (denser than water), whereas 

benzene is a light non-aqueous phase liquid (LNAPL) [less 

dense than water]. Previous reports and calculations made 

with data collected for this study showed evidence only of 

DNAPLs, not LNAPLs. Because DNAPLs have a different 

density than water, they behave uniquely in the subsurface 

environment depending on the route of entry, mixture of 

contaminants present in the DNAPL, and subsurface conditions (Cohen and Mercer, 1993). These properties allow them 

to migrate as a separate phase over long distances and to great 

depths away from their point of entry into the subsurface 

(Kueper and Davies, 2009). The presence and transport of 

DNAPLs is critical to the understanding of a contaminated site 

and ultimately the selected remediation, because their behavior 

and transport is different than the behavior of dissolved contaminants. In addition, DNAPLs may persist for centuries as a 

source of groundwater contamination, stored in the aquifer as 

dispersed droplets or pools along contacts of low permeability 

sediment layers and in low permeability sediment from matrix 

diffusion (Cohen and Mercer, 1993; Kueper and Davies, 

2009). Failure to understand the extent of DNAPLs can result 

in the expansion of contamination during either characterization or remediation activities (Cohen and Mercer, 1993). The 

following sections summarize past DNAPL characterization 

efforts at the SCD and provide an updated conceptual model 

for DNAPL extent in the subsurface using data collected during this study, in conjunction with other recent data collected 

in the wetland and the Columbia and Potomac aquifers.

Previous Dense Non-Aqueous Phase Liquid 

(DNAPL) Investigations

The presence of DNAPL has been reported within the 

Columbia aquifer for 25 years, starting with early investigations of the SCD. Quarterly sampling of monitoring wells in 

the late 1980s through the early 1990s consistently showed 

free product in some previous monitoring and recovery wells 

including TW-5, TW-28, TW-32, RW-2, and RW-5 (Roy F. 

Weston, Inc., 1993a) that are all located within the current 

barrier wall area, within and to the west of the main industrial 

area of the site, and near the settling pond (fig. 1). Laboratory 

reports indicate that DNAPL was visible as a thin layer of 

film at the base of the sample (Roy F. Weston, Inc., 1993a) 

collected from these wells. Similarly, early investigations 

conducted by the SCD (Conestoga Rovers Associates, 2000) 

indicated the likely presence of DNAPL (by the use of Sudan 

dye tests with soil samples) in the central industrial part of 

SCD (near previous monitoring well TW-30). On the basis of 

these initial results, possible DNAPL migration and extent was 

Table 12. Concentrations of chloride and sulfate measured in 

2008 in drive-point piezometers screened in the Columbia aquifer 

near the upland-wetland boundary east of the Standard Chlorine 

of Delaware, Inc. Superfund Site, Delaware.

[Samples analyzed by the U.S. Geological Survey National Water Quality 

Laboratory; mg/L, milligrams per liter; dup, duplicate]

Well name Sample date Chloride

(mg/L)

Sulfate

(mg/L)

EM-1D 6/12/2008 102 462

EM-2D 6/12/2008 61.0 2,450

EM-2D (dup) 6/12/2008 61.3 2,490

EM-5D 9/16/2008 55.1 180

EM-6D 9/16/2008 75.3 637

EM-6D (dup) 9/16/2008 74.6 636

400

60 1,600

1,400

1,200

1,000

800

600

400

200

0

350

50

300

250

40

200

30

150

20

100

50

10

0

0

12/18/08

Sampling date

7/6/09 1/22/10 8/10/10 2/26/11 9/14/11 4/1/12 10/18/12

Concentration, in milligrams per liter Concentration in DP-8A and DP-9, in milligrams per liter

Concentration in DP-6A and DP-7, in milligrams per liter

A. Chloride

B. Sulfate

DP-8A

DP-8A

DP-9

DP-9

DP-6A

DP-6A

DP-7

DP-7

Figure 25. Temporal changes in (A) chloride and (B) sulfate for 

selected sites in the wetland study area, July 2009–August 2012, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.
Dense Non-Aqueous Phase Liquid (DNAPL) Extent 51

included in the 1993 feasibility study (Roy F. Weston, Inc., 

1993a) and the 1995 ROD (U.S. Environmental Protection 

Agency, 1995a). The 1993 feasibility study estimated that the 

extent (based on effective solubility estimates with groundwater concentrations) extended beyond the barrier wall slightly to 

the west and included three monitoring points in the wetland 

near Red Lion Creek that were screened in the Columbia 

aquifer (U.S. Environmental Protection Agency, 1995). When 

these reports were published, the extent was estimated to be 

limited to the Columbia aquifer with migration interrupted 

by the underlying confining unit (Merchantville clay). Any 

migration within the Columbia aquifer was estimated to be 

toward the center of the SCD (Roy F. Weston, Inc., 1993a). 

Subsequent investigations have identified dissolved phase 

contaminants in the Upper Potomac aquifer underlying 

this confining layer (Merchantville/Upper Potomac clay) 

(Christopher Wolfe, HydroGeoLogic, Inc., written commun., 

2012), although to date, no evidence of DNAPL migration into 

the Lower Potomac aquifer has been reported. 

In preparation for the design and construction of the 

barrier wall and in keeping with directives outlined in the 

1995 ROD to limit and contain the migration of DNAPL, 

additional work was conducted in 2002 to aid in identifying 

potential DNAPL pools (by use of Membrane Interface Probe 

[MIP]/Electrical Conductivity [EC] screening), the extent of 

contamination, and the continuity of the clay layer underlying 

the contaminated Columbia aquifer (Black & Veatch, 2005). 

Extensive contamination was reported based on the MIP 

screening results (Black & Veatch, 2005); however, no 

additional spatial resolution of DNAPL pools was reported. 

EC results from the MIP/EC screening yielded locations for 

groundwater sampling and possible DNAPL identification 

(using FLUTe liners). Staining of the FLUTe liners installed 

at three locations in the central industrial part of SCD (out of 

the five tested) showed the presence of DNAPL in the range 

of 15.2–18.3 m (5–60 ft) below land surface in the Columbia 

aquifer (Black & Veatch, 2005). 

0

25

25

50

50

75

75

100

100

125

125

150

150

100

200

300

400

500

0

10

20

30

40

50

60

70

80

90

100

2.0 2.0

7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1 118.1 7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1

13.1 24.1 35.2 46.2 57.3 68.4 79.4 90.5 101.5 112.6 13.1 24.1 35.2 46.2 57.3 68.4 79.4 90.5 101.5 Chloride, in milligrams per liter Chloride and sulfate, in milligrams per liter

Chloride, in milligrams per liter

Depth below land surface, in centimeters

Sampling date

Depth below land surface, in centimeters

A. July 2009, Peeper Site 8

0

5

10

15

0

B. December 2011, Peeper Site 8

0

1

2

3

4

5

0

9/14/2011 12/8/2011 2/15/2012 4/26/2012

10/19/2011 1/17/2012 3/28/2012 8/9/2012

C. September 2011–August 2012, Piezometer DP-8A

EXPLANATION

Methane

Sulfide

Ferrous iron

Ammonia

Chloride

Sulfate

Ferrous iron, sulfide, and methane,

in milligrams per liter

Ferrous iron, sulfide, methane, and ammonia,

in milligrams per liter

Ferrous iron, sulfide, methane, and ammonia,

in milligrams per liter

Figure 26. Temporal changes in concentrations of chloride 

and redox-sensitive constituents at site 8 in the northwest 

wetland area from (A) peeper sampling in July 2009, (B) peeper 

sampling in December 2011, and (C) monthly sampling events, 

September 2011–August 2012, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware.
52 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

During this investigation, the EPA installed DNAPL 

recovery systems at recovery wells RW-2 (to the north of the 

settling basin) and RW-5 (at the western boundary of the existing barrier wall near the Air Products facility) in an attempt 

to remove product (Black & Veatch, 2005). Although some 

DNAPL was recovered, recharge was reported to be extremely 

slow; therefore, long-term recovery by this method was not 

considered to be cost-effective (Black & Veatch, 2005). 

Following the installation of the barrier wall, which was 

designed to contain the majority of known DNAPL extent, no 

subsequent DNAPL investigations have been conducted and 

no additional DNAPL areas have been identified. Review of 

the interim action (U.S. Environmental Protection Agency, 

2011) indicates that the barrier wall in combination with 

pumping is performing as designed. 

Updated Dense Non-Aqueous Phase Liquid 

(DNAPL) Site Conceptual Model

Although the barrier wall was designed to hydraulically 

contain most DNAPL from the SCD industrial area in the 

Columbia aquifer, the design report acknowledged that 

DNAPL may extend beyond the wall (Black & Veatch, 2005). 

Possible locations for this DNAPL underlie or are within the 

USGS wetland study area where high tide resulted in the fanning of contaminants into the wetland. Since minimal investigations of the wetland porewater and underlying groundwater 

from the Columbia and Potomac aquifers have been conducted 

until this study, an update to the DNAPL site conceptual 

model is provided. 

0

400

200

600

800

1,000

1,200

1,400

200 200

100 100

400 400

300 300

600 600

500 500

900 900

800 800

700 700

1,000 1,000

0

1

2

3

4

5

6

7

8

9

10

2.0 2.0

7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1 118.1 7.5 18.6 29.7 40.7 51.8 62.8 73.9 85.0 96.0 107.1

13.1 24.1 35.2 46.2 57.3 68.4 79.4 90.5 101.5 112.6 13.1 24.1 35.2 46.2 57.3 68.4 79.4 90.5 101.5 112.6 Chloride and sulfate, in milligrams per liter Chloride and sulfate, in milligrams per liter

Chloride and sulfate, in milligrams per liter

Depth below land surface, in centimeters

Sampling date

Depth below land surface, in centimeters

A. July 2009, Peeper Site 6

0

5

10

15

0

B. December 2011, Peeper Site 6

0

1

2

3

4

5

0

10/19/2011 12/8/2011 2/16/2012 4/26/2012

11/15/2011 1/17/2012 3/28/2012 7/2/2012

8/9/2012

C. September 2011–August 2012, Piezometer DP-6A

EXPLANATION

Methane

Sulfide

Ferrous iron

Ammonia

Chloride

Sulfate

Ferrous iron, sulfide, and methane,

in milligrams per liter

Ferrous iron, sulfide, methane, and ammonia,

in milligrams per liter

Ferrous iron, sulfide, methane, and ammonia,

in milligrams per liter

Figure 27. Temporal changes in concentrations of chloride 

and redox-sensitive constituents at site 6 in the northeast 

wetland area from (A) peeper sampling in July 2009, (B) peeper 

sampling in December 2011, and (C) monthly sampling events, 

September 2011–August 2012, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware.
Dense Non-Aqueous Phase Liquid (DNAPL) Extent 53

Dissolved and DNAPL phase contamination in the 

wetland may originate from various sources: (1) DNAPL 

resulting from overland flow from the 1986 spill that reached 

the wetlands, (2) subsequent product removal efforts that may 

have caused possible migration downward from the surface, 

and (3) long-term upward migration of dissolved contaminants 

into the wetland porewater due to vertical discharge from the 

Columbia aquifer that was contaminated from various waste 

disposal activities, leaks, and spills over the period of industrial operation. In addition, desorption of contaminants from 

highly organic wetland sediments may act as a secondary 

source. 

The likelihood of DNAPL contact with site groundwater 

or the actual presence of DNAPL in groundwater was estimated based on a calculated threshold concentration (C) of 

1 percent of the solubility, Si, (or effective solubility, Ci, in the 

case of 14DCB, 123TCB, and 124TCB) of a given compound 

(table 1). 

(0.01) C > Ci ∗ (8)

Where spill contents were known, the effective solubility, 

Ci, was calculated using Raoult’s Law (Kueper and Davies, 

2009), where mi is the mole fraction of compound in the 1986 

spill and Si is the single-component solubility:

Ci mi i = S (9)

In the case of 14DCB and 123TCB, the super-cooled 

solubilities were used in these calculations because these 

compounds are solids at steady state (VanNoort, 2009). Using 

the super-cooled solubility increases the calculated solubility 

of the compounds by about a factor of 2 (table 1), thus making 

the estimate of DNAPL presence more conservative. 

Columbia and Potomac Aquifers

Following the installation of the barrier wall, total VOC 

concentrations in the Columbia aquifer in the vicinity of the 

wetland (along the upgradient boundary at Haul Road) have 

decreased between 60 and 90 percent, as determined from 

quarterly monitoring reports (fig. 28). Despite this decrease, 

currently much of the Columbia aquifer upgradient of and 

underlying the USGS wetland study area is indicative of 

continued interaction with or the presence of DNAPL, based 

on the percent of solubility of the concentrations present in 

groundwater (fig. 29), and may serve as a possible contaminant source to the wetlands. Whereas 1 percent was used to 

indicate possible DNAPL extent, maximum concentrations 

of individual compounds in 2011–12 indicated a range of 2 

(13DCB at MW-21) to over 40 (123TCB at MW-25) percent 

solubility of individual compounds present in site groundwater. Total VOC concentrations in groundwater wells within the 

barrier wall near its northern extent were some of the highest 

detections currently identified in the Columbia aquifer, ranging 

from 100–145 mg/L (wells PZ29 and PZ11). Wells immediately north-northwest of the barrier wall boundary and just 

upgradient from the wetland study area contained total VOCs 

in the range of 75–90 mg/L (MW-25, PMW-45, and MW-24). 

Moving spatially from west to east toward MW-23, total 

concentrations declined by an order of magnitude and fewer 

compounds had concentrations that are indicative of DNAPL 

presence (fig. 29). As was the case with the Columbia aquifer, 

wells screened in the Potomac aquifer north of the barrier wall 

near the wetland-upland boundary were indicative of DNAPL 

presence or contact for all chlorinated benzenes present at 

the site (wells PW-14 and PW-17; fig. 29). The magnitude of 

concentration in these Potomac wells averaged about 60 mg/L 

(Christopher Wolfe, HydroGeoLogic, Inc., written commun., 

2012).

Groundwater concentrations of VOCs from drivepoint piezometers screened immediately below the wetland 

sediments in the Columbia aquifer differed in the northwest and northeast parts of the USGS wetland study area. 

Concentrations in wells located in the northwest wetland area 

are indicative of DNAPL (mainly for 14DCB, 12TCB, and 

124TCB), and total VOC concentrations ranged from 10 to 14 

mg/L (fig. 29). In the northeast wetland area, the overall magnitude of groundwater concentrations in the Columbia aquifer 

was considerably less, resulting in no exceedances of 1 percent 

of the solubility in this region (fig. 29). The spatial distribution 

of the DNAPL extent in the Columbia aquifer may indicate 

lateral migration from the central part of the SCD downgradient towards the wetland (fig. 8). Contaminant transport from 

the western drainage gully (fig. 1) may have had additional 

impact on the aquifer, particularly along the western boundary. 

Figure 28. Reduction in total volatile organic compound 

concentrations in monitoring wells along the wetland-upland 

boundary after the installation of the barrier wall, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware.

200,000

150,000

Before installation of barrier wall

After installation of barrier wall

(median 2010–12)

100,000

60-percent

reduction

80-percent

reduction 70-percent

reduction

50-percent

reduction

50,000

MW-25 MW-21 MW-22 MW-23

0

Total volatile organic compounds,

in micrograms per liter

Site number
54 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Wetland Porewater 

Results from the current study showed that groundwater 

in the overlying wetland sediments also had concentrations 

in the range of DNAPL contact or presence (fig. 30). The 

impacted wetland area is not spatially consistent with the 

footprint of potential DNAPL in the aquifer, and differences 

between the northwest and northeast areas are not as apparent. 

Possible DNAPL presence or contact extends farther east 

compared to the footprint in the Columbia aquifer (figs. 29, 

30). The area of impact is more spatially extensive, and in 

some cases, concentrations in shallow wetland groundwater 

are higher than concentrations in the immediately underlying 

aquifer. This increase may be a result of overland contamination in the wetland due to the 1986 spill, as was also noted 

from the vertical distributions of contaminants observed in 

the wetland groundwater in the peepers (fig. 17). Although 

the possibility of redistribution of contaminants due to the 

damaged tide control gate and current increased flooding of 

the wetland sediments cannot be ruled out, peeper samples 

collected in 2009 before the storm damage to the tide control 

structure (August 2011) showed the same general increase in 

VOC concentrations in the wetland groundwater near land 

surface (upper 40–50 cm) as observed in peeper samples collected in October–December 2011 (fig. 17). 

EM-1D

EM-1S

EM-2D

EM-2D

EM-6D

EM-6S

EM-5D

EM-5S

75°38'20" 75°38'10" 75°38'00" 75°37'50"

39°36'20"

39°36'10"

EXPLANATION

OR-6C

Barrier wall

DP-11B Drive-point piezometer Columbia monitoring well

Not indicative of DNAPL

Indicative of DNAPL contact or presence for 1–2 compounds (non-specific)

Indicative of DNAPL contact or presence for DCB + TCB

Indicative of DNAPL contact or presence for CB, 12DCB, 13DCB,14DCB,

123TCB, 124TCB

Potential DNAPL extent

Wetland dividing line

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

9

9

N

River Road River Road

Red Lion Creek

Haul Road

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

Unnamed Unnamed

Tributary Tributary

0

0

100 200 METERS

300 600 FEET

PZ-6

PZ-5

PZ-7

PZ-8

PZ-28 PZ-30

PZ-9

DP-7

PZ-11 PZ-13

MW-8

EW-41C

EW-61C

PZ-18

PZ-19

MW-9

PZ-10

MW-19

MW-20

PMW-45

MW-24

MW-24

MW-40

OP-1

MW-25

PZ-12 PZ-15 PMW-47

DP-11B

DP-10B

DP-13B

DP-12B

DP-9

DP-8

PZ-14

EW-51C PZ-16

PZ-29

PMW-46

PW-14

PW-17

MW-21

MW-22

EM-4D

DP-16B

DP-15B

DP-14B

DP-18B

DP-17B

PZ-17

MW-5 CW-11

MW-23

MW-37

MW-38

MW-36

MW-35

Figure 29. Potential current (2012) distribution of dense non-aqueous phase liquids (DNAPLs) in the Columbia aquifer, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware. (DCB, dichlorobenzene; TCB, trichlorobenzene; CB, chlorobenzene; 

12DCB, 1,2-dichlorobenzene; 13DCB, 1,3-dichlorobenzene; 1,4-DCB, 1,4-dichlorobenzene; 123TCB, 1,2,3-trichlorobenzene; 

124TCB, 1,2,4-trichlorobenzene.)
Dense Non-Aqueous Phase Liquid (DNAPL) Extent 55

Source Assessment for Wetland Study Area

An attempt was made to discern more about the contribution of possible sources of contamination to the USGS wetland 

study area (formerly referred to as the eastern wetlands in 

HydroGeoLogic, Inc., 2009) because this is critical in selecting and implementing a final remedy at the wetland area of 

the SCD. Molar composition, molar ratios, and total molar 

concentrations in groundwater samples collected from the 

aquifers and wetland groundwater were compared over time 

and in space to infer more about the transport of the contaminants and their possible origins. 

Columbia Aquifer and Overland Spills as Sources 

of Contamination 

In order to assess possible contributions of the contaminants from the aquifer to the wetland, an analysis of contaminant transport in the aquifer was conducted. Monitoring wells 

north of the spill areas (TW-49, MW-6, and MW-7) spaced 

approximately 61 m apart along the flowpath towards the 

wetland study area (MW-25) were used to compare VOC composition over time and distance (fig. 31a). Molar composition 

of these wells in 1993 and 2004 were dominated by CB and 

benzene, which was unexpected because the largest spill 

Figure 30. Potential current (2012) distribution of dense non-aqueous phase liquids (DNAPLs) in the wetland porewater, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware, on the basis of computed concentrations that exceed 1 percent of the 

compound’s solubility. (DCB, dichlorobenzene; TCB, trichlorobenzene; CB, chlorobenzene; 12DCB, 1,2-dichlorobenzene; 

13DCB, 1,3-dichlorobenzene; 1,4-DCB, 1,4-dichlorobenzene; 123TCB, 1,2,3-trichlorobenzene; 124TCB, 1,2,4-trichlorobenzene.)

75°38'20" 75°38'10" 75°38'00" 75°37'50"

39°36'20"

39°36'10"

EXPLANATION

Barrier wall DP-8A Drive-point piezometer PDB sampler

Not indicative of DNAPL

Indicative of DNAPL contact or presence for 1–2 compounds (non-specific)

Indicative of DNAPL contact or presence for DCB + TCB

Indicative of DNAPL contact or presence for CB, 12DCB, 13DCB,14DCB,

123TCB, 124TCB

Potential DNAPL extent

Wetland dividing line

Imagery from U.S. Geological Survey, Delaware Valley Regional Planning

Commission, and Department of Homeland Security, 2010, NAD 1983

State Plane Delaware Transverse Mercator Projection

9

9

N

River Road River Road

Red Lion Creek

Haul Road

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

STANDARD CHLORINE

OF DELAWARE, INC.

SUPERFUND SITE

Unnamed Unnamed

Tributary Tributary

0

0

100 200 METERS

300 600 FEET

EM-1S

DP-6A

122

DP-11A

143

142 141

140 144

137

135

139

102

133 132

131 136

134

118

117

106

112 114

113

108

160

158 156

157

110 109

119 116

129 128

121

122

123 125

126 127

147

146

148150

155

154

151

152

153

149

DP-10A

DP-104A

DP-107A

DP-13A DP-12A

DP-8A

DP-16A

DP-15A

DP-14A

DP-18A

DP-17A
56 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

(569,000 gallons in 1986) mostly contained 14DCB and 

TCBs. CB and benzene were components of the relatively 

small 1981 spill (5,000 gallons) and are the most soluble of the 

VOCs at the site. Similar total VOC concentrations and molar 

profiles were observed in the wells in 1993 and 2004 

(fig. 31a). There was little downgradient change over time 

to indicate that, prior to the installation of the barrier wall, 

contaminant mass may have migrated in groundwater to Haul 

Road at the upland boundary, particularly groundwater in 

contact with CB DNAPL from the 1981 spill. To evaluate this 

possible migration to the upland-wetland boundary, traveltimes of specific compounds were estimated using assumptions of bulk sediment properties and chemical properties to 

calculate a retardation factor, R, 

1 ( / ) R b e d = + p n ∗K (10)

where

pb = average bulk density of medium sand or 

1.59 grams per cubic centimeter (g/cm3);

ne = effective porosity of 0.3 (Domenico and 

Schwartz, 1990); and

Kd = the distribution coefficient (f

oc*Koc), where 

f

oc is the fraction of organic carbon in the 

sediment, and Koc is the organic carbon 

partitioning coefficient (table 1).

Organic carbon content analyses of sediment core samples 

collected at a depth of 6 m near the wetland-upland boundary 

gave a median f

oc of 0.003.

To estimate traveltime, T, over a given distance, L, 61 to 

122 m 

/ T = L Vc (11)

and contaminant (such as CB) velocity, Vc, 

/ Vc = v R (12)

where

v = average linear velocity in the central part of 

the site, 255 meters per year (836 feet per 

year [ft/yr], Black & Veatch, 2007).

Calculated traveltimes of CB and 14DCB in groundwater 

between wells TW-49 and MW-7 and MW-25 were less than 

the sampling time difference of 11 years shown in figure 31, 

whereas the traveltime of 124TCB exceeded 11 years 

(table 13). These traveltimes indicate that CB (and therefore 

benzene as well) and 14DCB could theoretically have traveled 

in groundwater to the wetland boundary (and beyond) in this 

11-year timeframe. However, given the more than an order 

of magnitude greater estimated traveltime for 124TCB than 

for CB and 14DCB, it is unlikely that 124TCB (and therefore 

123TCB, since these compounds were spilled together in 

1986) traveled in groundwater to the wetland boundary. 

Instead, these compounds may originate from sources nearer 

to the wetland as opposed to flow-through sources in the main 

part of the site. This analysis is consistent with previous discussions of possible sources north of the barrier wall 

(Black & Veatch, 2005). 

Concentrations in relatively shallow monitoring wells 

along Haul Road were compared before and after the barrier 

wall was installed (fig. 31, 2004 data from Black & Veatch, 

Benzene

Chlorobenzene

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total VOCs

EXPLANATION

0

500

1,000

1,500

2,000

2,500

0

10

20

30

40

50

60

70

80

90

100

Total volatile organic compounds (VOCs),

in micromolar

Total volatile organic compounds (VOCs),

in micromolar

Molar percent

Monitoring well number and year

Monitoring well number and year

A. 1993 and 2004

0

10

20

30

40

50

60

70

80

90

100

Molar percent

0

400

200

600

1,400

800

1,000

1,800

1,200

1,600

2,000

B. 2004 and 2012

TW-49

1993

MW-25

2004

MW-25

2012

MW-21

2004

MW-22

2004

MW-23

2004

MW-21

2012

MW-22

2012

MW-23

2012

MW-6

2004

MW-7

2004

MW-25

2004

Figure 31. Comparison of molar composition (in percent) of 

water samples from aquifer monitoring wells in (A) 1993 and 

2004, and (B) 2004 and 2012, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware.
Dense Non-Aqueous Phase Liquid (DNAPL) Extent 57

2007; Christopher Wolfe, HydroGeoLogic, Inc., written 

commun., 2012). As previously discussed (see figure 28), 

the barrier wall installation resulted in a reduction in overall 

concentrations in these wells. In 2012, there were clear differences between wells upgradient (south) of the northwest and 

northeast wetland areas (fig. 31). Upgradient of the northwest 

area (wells MW-25 and MW-21), there is a decreased predominance of CB and an increased predominance of 12DCB in 

the VOC molar composition between 2004 and 2012 (fig. 31). 

This pattern is less apparent moving from west to east where 

overall VOC concentrations declined, but the easternmost well 

MW-23 instead had an increase in the predominance of CB 

and benzene in 2012 compared to 2001 (fig. 31b). An increase 

in the molar percent of TCBs between 2004 and 2012 also was 

apparent in the western wells MW-25 and MW-21, whereas 

the opposite pattern was observed again in the easternmost 

well (MW-23, fig. 31b). In the northwest area, the decrease in 

the percentage of CB may be due to the presence of the barrier 

wall, installed immediately upgradient in 2007, interrupting 

a primary source of CB flowing toward the wetland in the 

aquifer. The simultaneous increase in the molar percent of 

other less soluble, non-spill compounds, such as 12DCB and 

the TCBs, may simply reflect their relative increased contribution to the overall VOC composition with the decline in CB, or 

they may reflect the relative increased importance of DNAPL 

sources outside the barrier wall on the VOC composition in 

the aquifer. The decline in chloride concentrations observed in 

the groundwater in the northwest wetland area between 2009 

and 2011 (fig. 25) supports the possibility of the barrier wall at 

least partly interrupting transport from a source upgradient of 

the wall. The increased molar percentages of CB and benzene 

in the northeast wetland (MW-23) also may be attributable to 

contaminant degradation, given the simultaneous reduction 

in other higher chlorinated benzene compounds and overall 

concentrations (fig. 31b). The possible effects of degradation 

on VOC composition is discussed in more detail in the Natural 

and Enhanced Biodegradation section.

Both similarities and differences between the northwest 

and northeast wetland areas were apparent in the piezometer 

pairs screened in the Columbia aquifer and wetland sediments (fig. 32). For 2012, data in the Columbia aquifer in the 

northwest wetlands and parts of the northeast wetland were 

consistent and only averaged concentrations from representative sites are shown (sites 10 and 12 in the northwest area 

and site 15 in the northeast area, fig. 32). Site 6 was included 

in this analysis, but due to large variability in concentrations 

between sampling events, data were not averaged and are not 

shown. Compared to the upland boundary wells in 2012 

(fig. 31), total VOC concentrations in groundwater from 

piezometers screened in the aquifer immediately below the 

wetland sediments were slightly lower (fig. 32). Molar composition of groundwater from piezometers closer to Red Lion 

Creek, DP-12B and DP-15B, differed from the upland boundary wells and from DP-10B, which is slightly upgradient of 

DP-12B, with decreased proportions of 12DCB and 124TCB 

and increased proportions of 13DCB and 123TCB (fig. 32). 

With decreasing depth, concentrations of VOCs increased 

in the wetland sediments nearest to the creek (DP-12B and 

DP-15B), and another shift in molar composition was 

observed (fig. 32). To better evaluate contaminant sources to 

the aquifer and the overlying wetland sediments, molar ratios, 

Rs, reflective of the 1986 spill compounds were calculated:

(14 ) /(123 124 ) Rs = DCB TCB + TCB (13)

Table 13. Traveltime of selected site contaminants in 

groundwater along a flowpath.

[CB, chlorobenzene; 14DCB, 1,4-dichlorobenzene; 124TCB, 1,2,4-trichlorobenzene]

Site 

compound Calculated R1

Calculated T2

(years) to travel 

between TW-49 

and MW-7 

Calculated T2

(years) to travel 

between TW-49 

and MW-25

CB 3.8 1 2

14DCB 7.9 2 4

124TCB 251 60 120

1 R = 1+ (pb / ne

)*Kd, where R is retardation factor, pb is dry bulk density, 

ne is effective porosity, and Kd is the retardation or distribution coefficient.

2 T = L/Vc

, where Vc = v/R, T is time, L is distance traveled, Vc is contaminant velocity, and v is linear horizontal velocity.

Figure 32. Comparison of molar composition (in percent) in 

the Columbia aquifer immediately below the wetland sediments 

and wetland porewater, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware, 2012.

0

50

100

150

200

250

0

10

20

30

40

50

60

70

80

90

100

Total volatile organic compounds (VOCs),

Molar per

in micromolar

cent

Monitoring well number

Columbia aquifer Wetland porewater

DP-10B DP-12B DP-15B DP-10A DP-12A DP-15A

Benzene

Chlorobenzene

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total VOCs

EXPLANATION
58 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

The Rs calculated from the reported composition of the 

1986 spill is 2.9. The Rs values calculated for the shallow 

Columbia aquifer were slightly higher than the 1986 spill 

composition and were fairly consistent, ranging from 4.0 at a 

site located immediately upgradient of the wetland boundary 

(MW-22) and between 4.5 to 5.8 at sites within the wetland 

area (fig. 33). In contrast, the Rs values of groundwater from 

the overlying wetland sediment were highly variable, ranging 

from about 2.1 to 8.8 (fig. 33). The 1986 spill molar ratio indicates that the deep (Columbia aquifer) and shallow wetland 

groundwater contamination differ from one another, likely 

due to mixing of different sources in the wetland groundwater 

and the effects of biodegradation and other natural attenuation 

processes (fig. 33). The greater variability in Rs is consistent 

with the change in molar compositions of the VOCs (fig. 32) 

in the wetland groundwater compared to the Columbia aquifer 

and indicates the complexity of processes that could affect 

the contaminant distribution in the wetland. A similar, limited 

analysis was used to evaluate possible common sources of 

contamination in the Potomac aquifer near the wetland compared to the Columbia aquifer or wetland groundwater. The 

total VOC concentrations in groundwater from the Potomac 

aquifer wells were similar to those detected in the Columbia 

aquifer at MW-25 and only slightly lower than concentrations 

measured inside the barrier wall (PZ-29) (fig. 34a). 

Groundwater samples from both the Potomac and Columbia 

aquifers indicated that about half of the molar mass is made 

up of CB and benzene (similar to the older profile from inside 

the barrier wall). However, some differences in the molar 

composition of the VOCs were found between the Columbia 

and Potomac aquifers. 123TCB and 13DCB were more dominant in the groundwater in the Potomac aquifer than in the 

Columbia aquifer, whereas 124TCB and 12DCB were more 

dominant in the Columbia aquifer (fig. 34a). Nonetheless, 

calculated source ratios (based on the 1986 spill, equation 13) 

between the Potomac aquifer and Columbia aquifer wells were 

similar (fig. 34b). It is possible that contamination in both 

aquifers resulted from the same contaminant source containing 

a mixture of 123TCB and 124TCB, but microbial degradation of the contaminants over time occurred through different 

pathways, leading to differing predominances in the TCB and 

DCB isomers.

Sediment as a Source of Contamination

In addition to the possible presence of contaminants as 

DNAPL or in contact with DNAPL, a long-term, secondary 

contaminant source to groundwater in the wetland may be the 

wetland sediments themselves due to desorption. The relatively low solubility and high affinity of the chlorinated benzenes for the solid phase (table 1) indicate that sorption and 

desorption are dominant processes in the wetland sediment. 

Simple theoretical calculations were conducted to estimate 

the equilibrium desorption of contaminants into groundwater. 

These calculated groundwater concentrations were compared to measured groundwater concentrations from USGS 

PDB sampling sites in the wetland sediments. Theoretical or 

predicted concentrations (Cpred) in groundwater were calculated using the f

oc of the sediments (USGS determined), 

average organic carbon partitioning coefficient (Koc, table 1), 

and sediment concentrations (Csed) from the shallow sampling 

depth detailed in the previous Wetland Remedial Approach 

study (HydroGeoLogic, Inc., 2009) according to the following 

formula. 

( ) / Cpred Koc oc sed = ∗ f C (14)

The predicted concentrations compare fairly well to 

measured concentrations in the wetland porewater, particularly 

for the higher chlorinated compounds (table 14). In the case 

of CB and 14DCB (both of which were spill compounds and 

also possible anaerobic degradation products of higher chlorinated benzenes) the predicted concentrations were greater 

than the actual measured concentrations by a factor of about 

2. Reasonable agreement of predicted and actual measured 

concentrations may indicate a large amount of sorbed mass in 

the sediments (near capacity). Gomez-Hermosillo and others 

(2006) reported that at very high concentrations of DCBs and 

TCBs in wetland porewater when the sorption capacity of 

the sediment is reached, desorption is largely linear for the 

chlorinated benzenes. In cases where concentrations are not 

0

50

100

150

200

250

0

1

2

3

4

5

6

7

8

9

10

Total volatile organic compounds (VOCs),

in micromolar

Molar ratio, micromolar/micromolar

Monitoring well number

MW-22 DP-10B DP-12B DP-15B DP-10A DP-12A DP-15A

Columbia aquifer Wetland sediment

EXPLANATION

14DCB/(124TCB+123TCB) Total VOCs

Figure 33. Comparison of volatile organic compound molar 

ratios (micromolar/micromolar) in the Columbia aquifer 

and wetland porewater for the 1986 spill compounds 

(1,4-dichlorobenzene/1,2,4-trichlorobenzene+

1,2,3-trichlorobenzene), Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware, 2012. (14DCB, 1,4-dichlorobenzene; 

124TCB, 1,2,4-trichlorobenzene; 123TCB, 1,2,3-trichlorobenzene.)
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 59

as significant, previous studies have reported between 30 and 

62 percent of CB to be relatively resistant to desorption (Shin, 

2000) and up to 95 percent of DCBs and TCBs present in a 

“slow desorption fraction” (ten Hulscher and others, 1997). 

This desorption resistance can result in the over-prediction 

of porewater concentrations using equilibrium partitioning. 

The relatively good agreement for high concentrations of 

chlorinated benzenes at SCD may indicate that equilibrium 

partitioning could be used to estimate sediment concentrations 

using measured groundwater concentrations, particularly in 

the western wetland in an effort to target additional sediment 

sample collection. 

Natural and Enhanced Biodegradation 

of Chlorinated Benzenes and Benzene

An understanding of biodegradation processes of site 

contaminants is critical for evaluating the feasibility of 

monitored natural attenuation or engineered bioremediation 

as part of the site remedial action. Defining biodegradation 

processes at SCD is complicated by the fact that contaminants 

used, stored, and released at the site include metabolites of 

the anaerobic degradation pathways; thus, for example, the 

presence of CB does not necessarily indicate that reductive 

dechlorination has occurred (fig. 2). In addition, aerobic degradation pathways produce only the ubiquitous end products 

of CO2, chloride, and water, or short-lived metabolites that 

could not be analyzed. Spatial and temporal variability in 

contaminant concentrations and the effect of possible abiotic 

processes, such as volatilization or evapotranspiration, also 

affect interpretation of field concentration data for biodegradation processes. To compensate for these complexities, a dual 

field and experimental approach was utilized and is discussed 

below, including evaluation of field geochemical data, field 

microcosms experiments under native (untreated) and stimulated conditions, and bench-scale experiments with bioreactors 

using site groundwater.

Evidence from Field Sampling Results

To evaluate biodegradation on the basis of field results, 

the molar VOC compositions of the water samples collected 

between 2009–12 were used in conjunction with knowledge of 

the redox conditions in the water, as indicated by the relative 

concentrations of the redox-sensitive constituents (figs. 16–19 

and 26–27). The molar compositions, rather than the individual VOC concentrations, can indicate production or removal of 

compounds despite the complication of possible metabolites, 

such as DCBs and CB, also being source (parent) contaminants at this site. Comparison of the VOC molar compositions 

in the wetland groundwater to the upland and wetland sediment sources assists in determining the occurrence and extent 

of biodegradation. Molar compositions of water samples 

from wells screened in the Columbia aquifer in upland source 

areas did indicate a change before and after installation of the 

barrier wall in the northwest compared to the northeast areas; 

however, the greater variability in the VOC compositions and 

calculated molar ratios, Rs (reflective of the 1986 spill compounds), in the wetland groundwater indicates that processes 

such as reductive dechlorination may be occurring to a greater 

extent in the wetland sediments (fig. 33). 

Compared to source VOC compositions, anaerobic 

degradation by reductive dechlorination would be expected 

to produce a higher molar percentage of the less chlorinated 

compounds, CB and benzene (fig. 2). Highly reducing conditions, such as methanogenesis compared to nitrate-reducing 

conditions, are most favorable for complete reductive 

Figure 34. Comparison of (A) molar composition (in percent) 

and (B) molar ratios (micromolar/micromolar) in the Columbia and 

Potomac aquifers, using average concentrations from 2011–12, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware.

EXPLANATION

Benzene

Chlorobenzene

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene (14DCB)

1,2,3-Trichlorobenzene (123TCB)

1,2,4-Trichlorobenzene (124TCB)

Total VOCs

0

200

400

600

800

1,000

0

10

20

30

40

50

60

70

80

90

100

Total volatile organic compounds (VOCs),

Molar per

in micromolar

cent

Monitoring well number

A

B

PZ-29 MW-25 PW-14 PW-17

Total volatile organic compounds

(VOCs), in micromolar

Molar ratio, micromolar/micromolar

Monitoring well number

0

1

2

3

4

5

6

7

0

800

600

400

200

1,000

PZ-29 MW-25 PW-14 PW-17

Columbia aquifer

Columbia aquifer

Potomac aquifer

Potomac aquifer

14DCB/(124TCB+123TCB)
60 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12 Table 14. Comparison of calculated and measured porewater volatile organic compound concentrations. [foc, fraction organic carbon; Csed, concentration in sediment; Cgw pred, predicted concentration in groundwater; μg/kg, micrograms per kilogram; μg/L, micrograms per liter; B, benzene; CB, chloroben- zene; 12DCB, 1,2-dichlorobenzene; 13DCB, 1,3-dichlorobenzene; 14DCB, 1,4-dichlorobenzene; 123TCB, 1,2,3-trichlorobenzene; 124TCB, 1,2,4-trichlorobenzene; --, no data; <, less than] Sediment identifier1 Location foc Csed (μg/kg)

B

Cgw pred

(μg/L)

B

Cgw 

actual

(μg/L)

B

Csed

(μg/kg)

CB

Cgw pred

(μg/L)

CB

Cgw 

actual

(μg/L)

CB

Csed

(μg/kg)

12DCB

Cgw pred

(μg/L)

12DCB

Cgw 

actual

(μg/L)

12DCB

Csed

(μg/kg)

13DCB

Cgw pred

(μg/L)

13DCB

Cgw 

actual

(μg/L)

13DCB

EW-07A

In vicinity 

of DP12, 

site 128

0.071 0 0 1,382 230,000 18,646 9,509 630,000 7,145 7,007 240,000 2,203 2,922 

EW-04A South of Site 6 0.071 23 5 <5 15,000 1,216 1,203 11,000 125 45 2,200 20 75 EW-05A In vicinity of site 158 0.071 0 0 5 5,100 413 1,227 2,700 31 6,999 1,400 13 198 

Sediment 

identifier1 Location foc

Csed

(μg/kg)

14DCB

Cgw pred

(μg/L)

14DCB

Cgw 

actual

(μg/L)

14DCB

Csed

(μg/kg)

123TCB

Cgw pred

(μg/L)

123TCB

Cgw 

actual

(μg/L)

123TCB

Csed

(μg/kg)

124TCB

Cgw pred

(μg/L)

124TCB

Cgw 

actual

(μg/L)

124TCB

EW-07A

In vicinity 

of DP12, 

site 128

0.071 560,000 18,089 9,604 130,000 566 263 340,000 305 553 

EW-04A South of site 6 0.071 12,000 388 226 51 0 <5 73 0 <5 EW-05A In vicinity of site 158 0.071 4,900 158 2,621 -- -- 4 -- -- 62 

1 Identification from HydroGeoLogic, Inc., 2009.
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 61

dechlorination. When the peeper profiles from July 2009 

showed highly reducing methanogenic conditions in the 

wetland groundwater at northwest site 8 compared to weaker 

reducing conditions (no detected methane and factor of 10 

lower ferrous iron concentration) at northeast site 6 (figs. 26a, 

27a), the VOC compositions at northwest site 8 showed a correspondingly higher percentage of CB and benzene compared 

to the northeast site (figs. 18a, 19a). The molar percentages of 

greater than 90 percent for CB and benzene combined indicate 

that TCBs and DCBs were being almost completely reduced 

through reductive dechlorination in the northwest wetland 

groundwater at site 8 in July 2009. Similar differences in 

redox conditions and VOC compositions at other peeper sites 

in July 2009 (site 9 in the northwest and site 7 in the northeast) 

gave evidence of similar extensive reductive dechlorination in 

the northwest compared to the northeast wetland study areas 

(data on file, USGS MD-DE-DC WSC). Chloride concentrations, which increase with chlorine released during each reductive dechlorination reaction, can provide another indication of 

the occurrence of this degradation process. Although chloride 

was high throughout the peeper profile at site 8 in July 2009 

(and at other northwest sites, probably from the HCl spill), a 

distinct increase in chloride was observed between depths of 

about 46 and 90 cm below land surface, corresponding to the 

depth over which the highest methane concentrations and lowest total CBs+benzene were observed (figs. 18a, 26a). 

Peeper results in 2011 showed a change in reducing 

conditions and associated change in the extent of reductive 

dechlorination at the northwest sites compared to the northeast 

sites. With a change to more moderate reducing conditions 

at site 8 in 2011, less benzene and more DCBs made up the 

molar VOC composition than those observed at this northwest 

site in 2009, indicating that reductive dechlorination was 

less complete in 2011. Molar VOC compositions across the 

wetland area from the PDB samples collected in 2011 also 

indicated that reductive dechlorination was less extensive in 

the northwest area than than in the northeast area. Temporal 

monitoring showed a continuation of the redox and VOC 

molar compositions observed in 2011 through August 2012. 

Thus, the change to more moderate reducing conditions and 

less extensive reductive dechlorination that was observed in 

the December 2011 peeper at site 8 compared to July 2009 

does not appear to be due to a short-term seasonal effect of 

lower temperatures on microbial activity. The opposite trend 

was observed in the peeper at site 6 in 2011 compared to 2009, 

with sulfide concentrations increasing and more extensive 

reductive dechlorination indicated by the higher percentage 

of CB and benzene in the upper approximately 60 cm of 

wetland sediment (figs. 18b, 26b). These peeper results were 

consistent with results for the northwest and northeast wetland 

areas collectively from the PDB sampling in October 2011. 

Overall, the significantly higher sulfide concentrations in the 

northeast area indicated more favorable reducing conditions 

for reductive dechlorination in the wetland groundwater in 

the northeast area than in the northwest area in October 2011, 

and CB was the predominant VOC in PDB samples from the 

northeast wetland area (median of 69 molar percent) compared 

to samples from the northwest wetland area (median of 35 

molar percent).

Volatilization/evapotranspiration and aerobic biodegradation could also affect the molar VOC composition in the 

wetland groundwater but both of these processes would be 

expected to preferentially remove the lighter (less chlorinated) 

compounds (Field and Sierra-Alvarez, 2008). Thus, volatilization and aerobic biodegradation would result in a decrease in 

the molar percentages of benzene and the lower chlorinated 

compounds, such as CB, in contrast to the effect of anaerobic reductive dechlorination. Some of the PDB samples and 

near-surface peeper samples with very low total VOC concentrations and high percentages of TCBs and DCBs could 

be indicative of these processes, such as samples from the 

2.0- and 7.5-cm depth in the December 2011 peeper at site 8 

(fig. 18b).

There are several possible causes of the changes observed 

between 2009 and 2011, including natural temporal changes in 

groundwater-flow conditions and geochemistry in the wetland 

area or induced changes from the installation of the barrier 

wall, industrial pumping effects, or the inoperative tide control 

structure. Because longer term data are not available for the 

wetland area, it is not possible to fully evaluate the most likely 

cause. However, the decrease in chloride concentrations in the 

northwest wetland area (fig. 25) and results of the 2009 and 

2010 in situ microcosms (see the following section) indicate 

that the change occurred before October 2010, which is before 

the problem with the tide control structure began (August 

2011). Industrial pumping for the DCRC was stopped from 

May 2010 to January 2011 and possibly affected groundwaterflow conditions and wetland geochemistry between the July 

2009 and October 2010 sampling in the wetland, but monthly 

groundwater sampling after January 2011 when pumping 

resumed indicated little change in wetland geochemistry 

(figs. 18c, 25, 26c). If the decreased chloride concentrations 

in the northwest wetland area were caused by the barrier 

wall decreasing or eliminating the contribution from a source 

behind the wall, it is possible that other constituents associated 

with this source acted to generate more highly reducing conditions in the Columbia aquifer discharging to the northwest 

wetland area, and contributed to the onset of highly methanogenic conditions observed in the 2009 wetland groundwater. 

The addition of selected redox constituents and chloride to the 

current monitoring plan for the upland area near the wetland 

boundary, and the addition of piezometers within the wetland 

area to this monitoring network would assist in determining 

the effect of the barrier wall on the wetland geochemistry, 

contaminant sources, and biodegradation. 

Except for limited near-surface decreases that could be 

due to volatilization or aerobic degradation, it is important to 

note that total VOC concentrations did not decrease substantially along upward groundwater-flow gradients in the peeper 

profiles, even when the molar compositions indicate substantial reductive dechlorination was occurring. This is probably 

due to the additional source of VOCs in the upper wetland 
62 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

sediments, present as DNAPL or sorbed to the sediment, that 

was evident throughout much of the northwest wetland area 

and part of the northeast wetland area (fig. 30). The presence 

of this source and its continued dissolution/desorption is 

indicated by the increased concentrations between depths of 

approximately 15 to 40 cm below land surface in many of the 

peepers (fig. 17). At site 8 in July 2009 and site 6 in December 

2011, reductive dechlorination apparently occurred at a sufficient rate to degrade contaminant input from this additional 

source, however, because the percentages of higher chlorinated compounds—DCBs and TCBs—did not increase along 

upward groundwater-flow gradients in the peeper profiles. In 

parts of the northeast wetland area where DNAPL presence 

was not evident (fig. 30) and concentrations of VOCs in PDBs 

were low or undetectable (fig. 16), natural attenuation may be 

occurring to completion within the shallow wetland groundwater. However, groundwater samples from deeper than about 

50 cm in the wetland sediments or aquifer were not collected 

and would be needed to determine whether contamination is 

present at these sites. Achieving more complete dechlorination 

of the chlorinated benzenes and removal of benzene, either 

by slowing groundwater-flow rates or enhancing degradation 

rates, would be necessary for a bioremediation remedy where 

DNAPL presence is still evident. 

In Situ Microcosms

In situ microcosms provided evidence under field conditions of the potential for biodegradation by the native microbial community in the wetland, with or without stimulation 

by addition of electron donors to support reductive dechlorination, and of the potential for biodegradation by a bioaugmented microbial consortium, WBC-2. Three microcosm 

treatment types were deployed at each site—a monitored natural attenuation treatment (referred to as the MNA treatment) 

that had no added amendments, a biostimulated treatment that 

was amended with lactate and chitin as donors for reductive 

dechlorination (lactate/chitin treatment), and a bioaugmented 

treatment amended with lactate/chitin and WBC-2 (WBC-2 

treatment). The molar compositions of VOCs and redox 

conditions in the three standard (no 13C label) treatments and 

the results of stable isotope probing for the 13C-labeled treatments are discussed for in situ microcosms incubated during 

October–December 2009 and October-December 2010 

(figs. 35–38; tables 15–16). Although in situ microcosms were 

deployed at sites 6 and 9 in October–December 2009, only site 

6 results are presented here because VOC concentrations and 

compositions in the three standard treatments at site 9 gave 

inconclusive evidence of biodegradation and 13C-labeled treatments were not performed at this site in 2009. 

The MNA microcosm at site 6 in October–December 

2009 had a relatively high percentage of combined TCBs and 

DCBs (greater than 60 percent), indicating a relatively low 

extent of naturally occurring reductive dechlorination of the 

VOCs (fig. 35). The aerobic or mildly reducing conditions in 

the MNA treatment, indicated by the high sulfate concentrations and low or undetectable concentrations of ammonium, 

ferrous iron, sulfide, and methane, were not conducive for 

reductive dechlorination of the chlorinated benzenes at site 

6 in October–November 2009. The MNA microcosm results 

(screen depth of 26–50 cm below land surface) were consistent 

with the results from the peeper sampled at site 6 in July 2009 

(figs. 19a, 27a). The high sulfide concentrations in the lactate/

chitin treatment compared to the MNA treatment showed that 

sulfate-reducing conditions were stimulated by the added 

donor, although sulfate concentrations and the molar composition of the VOCs showed little change. Addition of WBC-2 

and lactate/chitin, however, generated more highly reducing 

conditions than those observed in the MNA or lactate/chitin 

treatments, as shown by the complete removal of sulfate and 

increased sulfide production. A concomitant increase in the 

molar percentage of CB and benzene (collectively about 90 

percent of the total VOCs) occurred in the WBC-2 treatment 

compared to the MNA or lactate/chitin treatments, indicating 

an increase in reductive dechlorination of the higher chlorinated benzenes and CB with bioaugmentation (fig. 35a). 

The decrease in total VOC concentrations indicates complete 

degradation of VOCs in the bioaugmented treatment compared 

to the MNA or lactate/chitin treatments (fig. 35a).

The standard (no 13C label) in situ microcosms deployed 

in 2010 were variable for the four sites (fig. 36). Site 6 showed 

very little difference between the three treatments, unlike the 

2009 in situ microcosm results. The MNA treatment at site 6, 

however, indicated that reductive dechlorination was naturally 

more extensive at this site in 2010 than in 2009, even though 

the aerobic to mildly reducing conditions observed in the 

MNA treatment in 2010 were similar to the conditions in the 

2009 MNA treatment (table 15). The molar percentage of CB 

and benzene together was about 60 percent in the MNA treatment in 2010, compared to about 30 percent in 2009. Thus, 

the in situ microcosm results for VOCs at site 6 in 2010 are 

consistent with the VOC molar compositions observed in the 

peeper installed at this site in 2011, when more highly reducing conditions were observed in the peeper compared to 2009 

(figs. 19b, c; 27b, c). 

Of the four sites where microcosms were deployed in 

2010, in situ microcosms at northwestern site 8 showed the 

greatest difference between the MNA and biostimulated 

treatments, where total VOC concentrations decreased by 

more than 50 percent and the percentage of CB and benzene 

approximately doubled in the lactate/chitin biostimulated 

treatment. These changes in VOC concentration and composition are consistent with stimulation of more extensive 

reductive dechlorination due to the stimulation of more highly 

reducing conditions, as evidenced by high methane concentration in the lactate/chitin treatment at DP-8 (table 15). The 

addition of WBC-2 at site 8 did not substantially stimulate 

reductive dechlorination compared to the lactate/chitin treatment, indicating that the native microbial population at this 

site contained dechlorinators. These experimental results are 

consistent with the fact that reductive dechlorination appeared 
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 63

to be naturally extensive at site 9 and nearby site 8 in 2009, 

when highly reducing methanogenic conditions were naturally 

present (figs. 18a, 26a). The in situ microcosm results for sites 

6, 8, and 9 also indicate that the changes observed in the peepers between July 2009 and December 2011 may have occurred 

by October–November 2010. 

Although the standard in situ microcosms indicated that 

reductive dechlorination was producing CB and benzene at 

some sites under natural or stimulated conditions, the ultimate 

fate of these intermediate compounds could not be determined 

from these tests and instead required the use of the 13C-labeled 

treatments. In situ microcosms amended with 13C-CB provided 

direct, quantitative evidence of CB removal in the MNA and 

WBC-2 bioaugmented treatments at site 6 in 2009 (fig. 37a). 

Compared to predeployment measurements of the amount of 

13C-CB placed on the Bio-Sep beads, CB decreased by 66–69 

percent in the MNA and WBC-2 treatments over the 57-day 

incubation period in the wetland sediments, giving calculated 

first-order rate constants of 0.02 per day for both treatments. 

Although CB degradation rates were similar for the two treatments, the different redox conditions and distribution of 13C in 

microbial biomass (PLFA) and DIC, which would primarily be 

CO2, indicated that different degradation pathways occurred in 

the MNA and WBC-2 treatments (fig. 37). Not all redox-sensitive constituents were measured in the 13C-CB treatments, but 

sulfate concentrations of 426 and 11 mg/L, respectively in the 

MNA and WBC-2 treatments with 13C-CB at site 6 indicated 

similar redox conditions as those observed in the respective 

Sulfide

Ferrous iron

Ammonium

Sulfate

EXPLANATION

Benzene

Chlorobenzene

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total VOCs

EXPLANATION

0

100

50

150

200

250

300 12 600

500

400

300

200

100

10

8

6

4

2

0 0 0

10

20

30

40

50

60

70

80

90

100

Total VOCs + benzene, in micrograms per liter

Sulfate, in milligrams per liter

Concentration, in milligrams per liter

Molar VOC composition, in percent

A. 2009, VOCs, in situ microcosms B. 2009, Redox, in situ microcosms

MNA Lactate MNA

DP-6, 2009

Lactate

DP-6, 2009

WBC-2 WBC-2

Figure 35. (A) Molar composition of volatile organic compounds (VOCs), and (B) concentrations of redox-sensitive constituents in 

standard in situ microcosms (MNA, monitored natural attenuation treatment; lactate, lactate/chitin biostimulated treatment; WBC2, bioaugmented treatment) deployed in the wetland sediment at site DP-6, Standard Chlorine of Delaware, Inc. Superfund Site, 

Delaware, October 20–December 16, 2009. [Note that methane was not detected.]

EXPLANATION

Benzene

Chlorobenzene

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene

1,2,4-Trichlorobenzene

Total VOCs

0

20,000

10,000

30,000

40,000

50,000

60,000

0

10

20

30

40

50

60

70

80

90

100

Total VOCs + benzene, in micrograms per liter

Molar VOC composition, in percent

2010, In situ microcosms MNA MNA

MNA

MNA

Lactate

WBC-2

Lactate

WBC-2

Lactate

WBC-2

Lactate

WBC-2

DP-8, 2010 DP-104, 2010 DP-107, 2010 DP-6, 2010

Figure 36. Molar composition of volatile organic compounds 

(VOCs) in standard in situ microcosms (MNA, monitored natural 

attenuation treatment; lactate, lactate/chitin biostimulated 

treatment; WBC-2, bioaugmented treatment) deployed in the 

wetland sediment at four sites, Standard Chlorine of Delaware, 

Inc. Superfund Site, Delaware, October 26–December 14, 2010.
64 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

standard microcosm treatments (fig. 35b). Both the MNA 

and WBC-2 treatments showed incorporation of 13C into DIC 

(or CO2), which can be produced as an end product through 

aerobic or anaerobic degradation pathways of the CB and 

indicates use of the contaminant for cellular energy. Only the 

MNA treatment with 13C-CB showed incorporation of 13C in 

the microbial biomass measured on the beads, indicating that 

some microorganisms were utilizing the carbon as a source for 

cellular material through oxidative degradation pathways 

(figs. 37b, c). In reductive dechlorination, the chlorinated 

contaminant is utilized as an energy source (electron acceptor) 

but not as a carbon source for cellular growth (Vogel and others, 1987). Previous studies have shown that CB is susceptible 

to aerobic degradation through oxidation, and the high sulfate 

concentrations and low or undetectable concentrations of the 

reduced species ammonium, ferrous iron, sulfide, and methane 

support the possibility of dissolved oxygen availability in the 

wetland sediment for aerobic oxidation to occur in the MNA 

treatment (fig. 35b). Anaerobic (rather than aerobic) oxidation 

of CB is not known to occur; although anaerobic oxidation 

has been reported previously for other chlorinated VOCs. 

Recent studies indicate that oxidation under apparent anaerobic conditions may actually occur by aerobic oxidation under 

microaerophilic conditions, and not by oxidation by anaerobic 

bacteria (Bradley and Chapelle, 2011).

The high percent removal of 13C-CB and production of 13C-DIC in the WBC-2 treatment at site 6 under sulfate-reducing conditions (figs. 35 and 37a, b) is significant because few 

field studies have reported evidence of anaerobic degradation 

of monochlorobenzene to either benzene or completely to CO2

(Fung and others, 2009). In the MNA treatment, it also is possible that some of the CB was initially reduced to benzene and 

the benzene was oxidized to CO2, although redox conditions 

were not conducive to reductive dechlorination in this treatment. All three of the 13C-CB in situ microcosms deployed in 

2010 also showed 13C incorporation in DIC, indicating that the 

native microbial population in both the northwest and northeast wetland areas are capable of complete degradation of CB 

(fig. 38a). The average δ13DIC for the the MNA treatments 

at the 2010 sites was 1,420 per mil, which was similar to the 

δ13DIC of 1,190 per mil measured for the MNA treatment in 

2009. Incorporation of 13C in the PLFA of the biomass also 

was observed in the MNA treatments with 13C-CB at all three 

sites in 2010, although the δ13-PLFA values (8 to 176 per mil) 

were substantially lower than those observed for δ13DIC. The 

amount of PLFA biomass enriched with 13C was two orders of 

magnitude lower than the total biomass measured (fig. 38b). 

It is significant, however, that a low amount of oxidation of 

CB (or the anaerobic metabolite, benzene) and incorporation 

into the biomass was possible in the wetland area even where 

methane concentrations indicated highly reducing conditions, 

especially at site 104 (table 15). Macropores, especially along 

root channels, in the wetland sediment may allow low amounts 

of oxygen into the wetland sediment to support oxidation 

of the CB even where bulk water concentrations of redoxsensitive constituents indicate nominally reducing conditions. 

1.00E+04

0.08

1.00E+03

0.06

1.00E+02

0.04

1,186

4.2E+05

1.6E+03

7.8E+05

54

2,170

1.00E+01

0.02

MNA

Predeployment

MNA

DP-6, 2009

DP-6, 2009

WBC-2

MNA WBC-2

WBC-2

1.00E+00

0.00

1.00E+07

1.00E+06

1.00E+05

1.00E+04

1.00E+03

1.00E+02

1.00E+01

1.00E+00

 C PLFA or DIC, in per mil 13 C-chlorobenzene enriched PLFA,

in cells per bead

13 C-chlorobenzene concentration,

in milligrams per bead

13

13 13 B. C incorporation in DIC and PLFA, C-chlorobenzene

A. 13 DP-6, 2009: C-chlorobenzene loss

C. 13 13 C enriched PLFA biomass, C-chlorobenzene

13

2

C utilized

for CO (DIC)

13C utilized

for PLFA

Total biomass

13C enriched

biomass

Figure 37. (A) Removal of 13C-chlorobenzene, (B) incorporation 

of 13C in dissolved inorganic carbon (DIC) and phospholipid 

fatty acid (PLFA) biomass, and (C) enrichment of 13C in the PLFA 

biomass in the 13C-chlorobenzene amended in situ microcosms 

(MNA, monitored natural attenuation; WBC-2, bioaugmented 

treatment) deployed in the wetland sediment at site DP-6, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, 

October 20–December 16, 2009.
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 65

Oxygen transport to the wetland sediments would be expected 

to be greater during the growing season, when the wetland 

plants actively transport oxygen to their roots, compared to the 

December–October period of microcosm deployment. For the 

lactate/chitin and WBC-2 treatments amended with 13C-CB in 

2010, incorporation into 13DIC was observed at some sites, but 

not into the biomass (fig. 38a). 

Natural, complete breakdown of 14DCB to CO2 was 

shown in the MNA treatments with 13C-14DCB microcosms at 

two of the four sites tested in 2010, sites 107 and 104, which 

are in or near the northeast wetland area (table 16). Aerobic 

to mildly reducing conditions that were conducive for aerobic 

oxidation of the 14DCB were naturally present (MNA treatment) at site 6 during 2010, similar to the conditions observed 

in 2009. However, site 107 showed more highly reducing conditions than site 6 in 2010, with low sulfate concentrations and 

relatively high sulfide and methane (table 15). The δ13DIC values in the 13C-14DCB microcosms were an order of magnitude 

lower than those observed for the MNA treatments amended 

with 13C-CB, which may indicate a slower complete degradation rate for 14DCB to DIC than observed for CB under these 

site conditions (table 16, fig. 38a). Oxidation rates of chlorinated VOCs generally decrease with an increasing number 

of chlorines (Field and Sierra-Alvarez, 2008). No 13DIC was 

measured in the lactate/chitin or WBC-2 treatments amended 

with 13C-14DCB, although this does not preclude the possibility that 14DCB was degraded to benzene through reductive 

dechlorination, because carbon isotope ratios of intermediate 

metabolites could not be analyzed. 

In situ microcosms amended with 13C-benzene were 

only deployed at site 6 in 2010, and these showed complete 

mineralization to 13DIC in the MNA, lactate/chitin, and 

WBC-2 treatments (table 16). The δ13DIC measured in the 

MNA treatment amended with 13C-benzene was substantially 

greater than the δ13DIC measured in the other 13C-benzene 

treatments, or in the 14DCB or CB 13C-labeled microcosms, 

indicating a relatively rapid complete breakdown of benzene 

under natural conditions at the site. Incorporation of 13C into 

Figure 38. (A) Incorporation of 13C in dissolved inorganic carbon (DIC) and phospholipid fatty acid (PLFA) biomass, 

and (B) enrichment of 13C in the PLFA biomass in the 13C-chlorobenzene amended in situ microcosms (MNA, 

monitored natural attenuation; lactate, lactate/chitin biostimulated treatment; WBC-2, bioaugmented treatment) 

deployed in the wetland sediment at three sites, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, 

October 26–December 14, 2010.

MNA

MNA

MNA

MNA

MNA

MNA

WBC-2

WBC-2

WBC-2

WBC-2

WBC-2

WBC-2

Not detected

Not detected

Not detected

Not detected

Not detected

Not detected

Not detected

Not detected

Not detected

Not detected

Not detected

Lactate

DP-8,

2010

Lactate

DP-8,

2010

Lactate

DP-104,

2010

Lactate

DP-104,

2010

Lactate

DP-107,

2010

Lactate

DP-107,

2010

1.00E+04

1.00E+03

1.00E+02

1.00E+01

1.00E+00

1.00E+07

1.00E+06

1.00E+05

1.00E+04

1.00E+03

1.00E+02

1.00E+01

1.00E+00

13 13 A. C incorporation in DIC and PLFA, C-chlorobenzene

 C PLFA or DIC, in per mil 13

C-chlorobenzene enriched PLFA,

in cells per bead

13

13 13 B. C enriched PLFA biomass, C-chlorobenzene

13

2

C utilized

for CO (DIC)

13C utilized

for PLFA

Total

biomass

13C enriched

biomass

597 8.3E+06

7.8E+06

3.0E+06

1.6E+06

1.3E+04

5.0E+06

9.0E+05

5.3E+05

4.7E+03

3.1E+06

3.1E+06

1.5E+04 8

10

1,179

43

2,496

66

176

2,354

Not detected
66 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

biomass was not observed for the 13C-benzene treatments at 

site 6 (table 16). This is surprising because benzene is known 

to be easily oxidized by many microorganisms under aerobic 

conditions, and the natural (MNA) conditions at this site in 

December 2010 indicated fairly oxidizing conditions (table 

15). Anaerobic oxidation of benzene can occur under a wide 

range of anaerobic conditions but generally at a slower rate 

than aerobic oxidation.

Microbial Communities in the Wetland 

Sediment 

Microbial communities in the wetland sediment were 

determined by direct analysis of sediment samples collected 

from sites 6 and 9 when the in situ microcosms were removed 

during December 2009 (figs. 39, 40). Cores were collected 

about 0.5 m laterally from the in situ microcosms and at two 

depths, one of which was near the top of the screened interval 

of the microcosms, to allow comparison to the microbial samples collected from the microcosms. Microbial communities 

were analyzed only in the in situ microcosm samples from site 

6 at this time. Microorganisms that colonize Bio-Sep beads 

from groundwater flowing through the in situ microcosms 

over the incubation period are generally considered representative of the community in the surrounding sediment (Peacock 

and others, 2004; North and others, 2012). Because many 

microorganisms prefer to attach to surfaces, the microcosm 

samplers can detect microorganisms that may not be detected 

by sampling the groundwater and may provide more representative community analyses than those obtained by sampling 

the groundwater (Peacock and others, 2004; North and others, 

2012). Microbial analyses performed in 2009 included qPCR 

determination of total eubacteria, which includes all bacteria

except for methanogens and other archaebacteria, and of specific genera of bacteria that are known to have species capable 

Table 15. Concentrations of redox-sensitive constituents and volatile fatty acids measured in samples from in situ microcosms 

deployed from October 20–December 16, 2009 and from October 26–December 14, 2010 at selected wetland sites, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware.

[Units for redox-sensitive constituents are milligrams per liter, and units for volatile fatty acids are micrograms per liter. Ferrous iron, sulfide, and methane 

analyzed by U.S. Geological Survey (Baltimore, Md.), and all other constituents analyzed by Microbial Insights (Rockford, Tenn.). Treatments included monitored natural attenuation (MNA), (control); lactate/chitin, donor addition for biostimulation; and WBC-2, West Branch Consortium-2, for bioaugmentation; 

WBC-2 treatment also contained added lactate/chitin; <, less than, NS, not sampled (broken bottle)]

Wetland 

area

Site 

number

Sample 

date Treatment

Redox-sensitive constituents Volatile fatty acids

Nitrate, 

as 

Nitrogen

Sulfate, 

as SO4

Ferrous 

Iron, 

as Fe

Sulfide, 

as S Methane Lactate Acetate Propionate

Pyruvate

Northwest 

DP-8 12-14-10 MNA <0.10 3.1 1.8 0.00 0.031 <0.39 40 7.8 5.1

DP-8 12-14-10 Lactate/

chitin <0.10 <0.73 31 0.09 7.95 1.6 660 910 <0.69

DP-8 12-14-10 WBC-2 NS NS NS NS 1.11 NS NS NS NS

Northwest 

DP-104 12-14-10 MNA <0.10 5.3 7.4 0.00 8.28 4.9 380 2.3 <0.69

DP-104 12-14-10 Lactate/

chitin <0.10 4.5 2.4 0.00 6.85 4,800 360 26 <6.9

DP-104 12-14-10 WBC-2 2.7 118 40 0.86 4.44 340 5,950 11,600 11

Northeast

DP-107 12-14-10 MNA <0.10 8.6 0.87 0.4 0.82 2.7 130 0.8 <0.69

DP-107 12-14-10 Lactate/

chitin <0.10 1.5 1.4 1.2 0.57 <0.78 850 1,360 1.1

DP-107 12-14-10 WBC-2 <0.10 <0.73 0.58 0.65 0.45 <0.78 650 1,030 <0.14

Northeast

DP-6 12-14-10 MNA <0.10 1,460 1.1 0.00 0.024 0.47 1.70 4.7 6.2

DP-6 12-14-10 Lactate/

chitin <0.10 248 2.6 5.2 0.50 1,300 1,210 1,530 <0.14

DP-6 12-14-10 WBC-2 <0.10 129 <0.50 6.6 0.97 1.2 250 280 <0.69

Northeast

DP-6 12-16-09 MNA <0.10 448 0.28 0.49 <0.040 <0.39 <0.54 <0.31 5.0

DP-6 12-16-09 Lactate/

chitin <0.10 531 0.69 7.5 <0.040 <0.39 27 21 6.0

DP-6 12-16-09 WBC-2 <0.10 7.5 0.37 10 <0.040 <0.39 <0.54 <0.31 <0.69
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 67

of reductive dechlorination of chlorinated VOCs, including 

Dehalococcoides, Dehalobacter, and Desulfitobacterium species. Only total eubacteria and Dehalococcoides species were 

determined by qPCR for the in situ microcosm samples collected in 2010 (fig. 41), but other microbial species, including 

Geobacter, that also are capable of reductive dechlorination of 

chlorinated VOCs were identified using DGGE analysis 

(table 17). In addition, the general microbial community composition was evaluated from PLFA structural groups during 

2009 and 2010 (table 5; fig. 40). 

The population density of total eubacteria in the in situ

microcosm MNA treatment at site 6 in 2009 was about 10 

times lower than the population densities measured at two 

depths in the wetland sediment at the same site (fig. 39). 

Populations of Dehalococcoides and Desulfitobacterium species also were lower in the in situ microcosm MNA treatment 

than in the wetland sediment at 6, although Dehalobacter

populations were similar (fig. 39). At both sites, the populations of total eubacteria and dechlorinating species were very 

similar in sediment samples from the two different depths, 

indicating uniformity with depth. Vertical variability cannot 

be discounted as a factor in the lower populations observed in 

the in situ microcosm sample, however, because the screened 

interval of the in situ microcosm extends about 20 cm below 

the deepest sediment sample collected. With the greater number of MNA in situ microcosms deployed in 2010 (including 

standard and 13C-labeled treatments), spatial variability in the 

native microbial populations was apparent at each site 

(fig. 41). Total eubacteria population levels commonly differed 

by a factor of 10 among the three MNA treatments deployed at 

each of the sites in 2010, and total Dehalococcoides populations varied from less than 1,000 cells per bead to to slightly 

more than 10,000 cells per bead at two of the sites. The PLFA 

structural analysis of the sediment and MNA microcosm sample in 2009 indicated similar microbial community compositions, with Firmicutes (fermenting bacteria) and Proteobacteria 

(gram negative bacteria) the most dominant groups. Overall, 

comparison of the sediment and in situ microcosm samples 

indicates that the microcosms provided microbial samples 

Table 16. Amount of 13C incorporation into phospholipid fatty acid (PLFA) biomass and dissolved inorganic carbon (DIC) measured in 

in situ microcosms that were preloaded with 13C-labeled 1,4-dichlorobenzene or benzene and deployed from October 26–December 14, 

2010 at selected wetland sites.

[All constituents analyzed by Microbial Insights (Rockford, Tenn.). Treatments included monitored natural attenuation (MNA), (control); lactate/chitin, donor 

addition for biostimulation; and WBC-2, West Branch Consortium-2, for bioaugmentation; WBC-2 treatment also contained added lactate/chitin; ND, not 

detected; total biomass is given as a cell equivalent based on the total amount of phospholipid fatty acids (PLFAs) extracted from a sample and includes 

bacterial and eukaryotic biomass]

Wetland 

area

Site 

number

Sample 

date Treatment

Total biomass 

(cells per 

bead)

13C enriched 

biomass PLFA 

(cells per 

bead)

13C in biomass, 

as average 

δ 13C-PLFA 

(per mil)

13C in DIC, 

δ 13C-DIC 

(per mil)

13C-Dichlorobenzene

Northwest 

DP-8 12-14-10 MNA 4.54E+6 ND ND -11

DP-8 12-14-10 Lactate/chitin 5.58E+6 ND ND -11

DP-8 12-14-10 WBC-2 5.43E+6 ND ND -10

Northwest 

DP-104 12-14-10 MNA 2.00E+6 ND ND -17

DP-104 12-14-10 Lactate/chitin 1.56E+6 ND ND -11

DP-104 12-14-10 WBC-2 1.84E+6 ND ND -14

Northeast

DP-107 12-14-10 MNA 3.75E+6 ND ND 129

DP-107 12-14-10 Lactate/chitin 3.73E+6 ND ND -11

DP-107 12-14-10 WBC-2 4.58E+6 ND ND -15

Northeast

DP-6 12-14-10 MNA 7.21E+6 ND ND 286

DP-6 12-14-10 Lactate/chitin 7.59E+6 ND ND -12

DP-6 12-14-10 WBC-2 5.30E+6 ND ND -18

13C-Benzene

Northeast

DP-6 12-14-10 MNA 4.10E+6 ND ND 10,110

DP-6 12-14-10 Lactate/chitin 6.01E+6 ND ND 61

DP-6 12-14-10 WBC-2 4.53E+6 ND ND 244
68 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

that are representative of the native populations in the wetland 

sediment.

The relatively high cell numbers and diversity of populations (Dehalococcoides, Dehalobacter, Desulfitobacterium, 

and Geobacter) known to contain species capable of reductive 

dechlorination of VOCs indicates a high natural dechlorination potential in the wetland sediment at the SCD, confirming 

the geochemical evidence from peeper and other groundwater 

samplers and from the in situ microcosms. Relatively high 

total eubacteria population densities were found even at sites 

where high VOC concentrations were indicative of DNAPL 

presence (sites 8, 104, 107 in fig. 30). VOC concentrations at 

site 8 were substantially higher than those at the other sites 

but showed similar eubacteria population levels (figs. 36, 

41a). Dehalococcoides population levels were lower in the 

MNA microcosms from two of the sites, 8 and 104, that had 

VOC concentrations indicative of DNAPL presence, although 

stimulation with lactate/chitin resulted in measurable increases 

in Dehalococcoides levels (fig. 41b). Both sites 8 and 104 are 

located in the northwest wetland area where the low pH may 

have inhibited bacterial growth and dechlorination activity 

(Zhuang and Pavlostathis, 1995; Schaefer and others, 2010). 

Besides providing donor for reductive dechlorination, the 

added lactate/chitin in the biostimulated microcosms provided 

pH buffering that could have stimulated growth of the dechlorinating species. The biostimulated treatments also had higher 

Dehalococcoides population levels than the MNA treatments 

in several of the in situ microcosms placed at the northeast 

sites 107 and 6 . Although the pH in the northeast wetland area 

was higher than the pH in the northwest wetland area, pH levels across the wetland were lower than the near neutral (7.0) 

optimum pH range for growth of several known dechlorinating 

Dehalococcoides and Desulfitobacterium species (Suyama and 

others, 2001; Tas and others, 2010). All in situ microcosms 

bioaugmented with WBC-2 showed similar Dehalococcoides

population levels, indicating similar survival in the wetland 

porewater with varying VOC contaminant concentrations, 

including the areas where DNAPL presence was indicated 

(figs. 30, 41b).

Total Eubacteria

Dehalococcoides spp.

Dehalobacter spp.

Desulfitobacterium spp.

December 2009, Microbial populations from qPCR:

Sediment and MNA

1.00E+09

1.00E+08

1.00E+07

1.00E+06

1.00E+05

1.00E+04

1.00E+03

Sediment

12.7 cm

(centimeters)

Sediment

27.9 cm

Sediment

12.7 cm

Sediment

27.9 cm

DP-9 DP-9

EXPLANATION

DP-6 DP-6 DP-6

MNA- C

26–50 cm

13

Microbial populations, in cells per bead

Figure 39. Microbial populations determined by quantitative 

polymerase chain reaction (qPCR) analysis of wetland 

sediment collected at sites DP-9 and DP-6 and in the control 

(MNA, monitored natural attenuation) in situ microcosm 

deployed in the wetland sediment at site DP-6, Standard 

Chlorine of Delaware, Inc. Superfund Site, Delaware, 

October 20–December 16, 2009.

Figure 40. Microbial community composition determined 

from analysis of the chemical structures of phospholipid 

fatty acid (PLFA) in wetland sediment collected at sites 

DP-9 and DP-6 and in the control (MNA, monitored natural 

attenuation) in situ microcosm deployed in the wetland 

sediment at site DP-6, Standard Chlorine of Delaware, Inc. 

Superfund Site, Delaware, October 20–December 16, 2009. 

(See table 5 for descriptions of PLFA structural groups.)

Eukaryotes (polyenoics)

General (Nsats)

Actinomycetes (MidBrSats)

Proteobacteria (Monos)

Firmacutes (TerBrSats)

Anaerobic metal reducers (BrMonos)

December 2009, Microbial populations from qPCR:

Sediment and MNA

100

90

80

70

60

50

40

30

20

10

0

Sediment

12.7 cm

(centimeters)

Sediment

27.9 cm

Sediment

12.7 cm

Sediment

27.9 cm

DP-9 DP-9

EXPLANATION

DP-6 DP-6 DP-6

MNA- C

26–50 cm

13

Percent of total community
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 69

Laboratory Bioreactor Experiments

Testing conducted with three flow-through bioreactors from March–October 2012 provided further evidence 

that both the natural wetland microbial community and the 

WBC-2 culture are capable of biodegrading the chlorinated 

benzenes and benzene at the SCD. A substantial decrease in 

concentrations between the inflow (“tank” sample) and the 

outflow (“waste” sample) throughout the tests was observed in 

the three bioreactors (fig. 42). Near-neutral pH and anaerobic 

conditions, characterized by negative ORPs, were initially 

established in all three bioreactors (figs. 43 and 44). The “C,” 

or control bioreactor, which was seeded only with bacteria 

that colonized the bioreactor matrix from the inflow water 

from well DP-15B, can be compared to bioreactor “D,” 

which received the same groundwater from well DP-15B but 

contained a well-established biofilm of WBC-2 on the matrix. 

The “E” bioreactor also was pre-established with WBC-2 but 

received groundwater from well DP-10B, which had higher 

total concentrations of CBs+benzene and lower concentrations 

of sulfate than the input water from well DP-15B (figs. 42 

and 45). Amendments to bioreactor C, including the amounts 

of bicarbonate as a buffer, nutrients, and carbon donor, were 

altered during the testing period to evaluate the factors controlling biodegradation, whereas these amendments were kept 

about the same until near the end of testing in the D bioreactor 

and throughout the entire testing in the E bioreactor 

(fig. 42). TeCA, the primary VOC that WBC-2 was developed 

to degrade, was added as a co-contaminant to all bioreactors 

beginning in August 2012 to assist in evaluating the activity 

of WBC-2 and its importance relative to the native microbial 

communities in degrading the chlorinated benzenes in the D 

and E bioreactors (fig. 42). The geochemical changes resulting 

from these amendments and their effects on biodegradation are 

discussed below. 

Microbial species that are capable of reductive dechlorination of chorinated VOCs generally are strict anaerobes, and 

WBC-2 performs most efficiently under anaerobic conditions, 

although it tolerates some oxygen exposure. The chlorinated 

benzenes and benzene in the study area also are highly susceptible to oxidative degradation pathways by aerobic bacteria. 

Sequential anaerobic and aerobic conditions were tested in the 

C bioreactor because results of the in situ microcosms indicated that degradation of the chlorinated benzenes, especially 

CB, could be occurring by both anaerobic and aerobic bacteria 

in the wetland sediment. Amendment with excess amounts of 

soluble organic donors (corn syrup and lactate) was used to 

generate and maintain anaerobic conditions in the bioreactors, 

and the donor amendments were stopped to try to generate aerobic conditions. From April to mid-June 2012 when the three 

bioreactors received the same donor amendments, ORPs were 

around -300 mV in all bioreactor cells and outflows, indicating 

that similar reducing conditions that are favorable for reductive dechlorination of the chlorinated benzenes had been 

established. Concentrations of sulfate, sulfide, and methane, 

however, indicated that conditions were less reducing in the C 

bioreactor than in the D and E bioreactors during this period 

(figs. 45 and 46). Sulfate concentrations only decreased about 

10 to 15 percent in the outflow (waste) from the C bioreactor 

compared to the input concentrations in the tank. Sulfide 

Table 17. Summary of microbial species identified using denaturing gradient gel electrophoresis (DGGE) analysis of samplers from 

the monitored natural attenuation (MNA) and lactate/chitin biostimulated treatments of in situ microcosms deployed from October–

December 2010.

Site Treatment

Clostridium

species

(Obligately 

anaerobic fermentors; some species 

reduce metals)

Caldiscericum

species

(Obligately anaerobic, 

chemoheterotrophic 

thermophiles that can 

reduce sulfur 

compounds)

Geobacter 

species

(Anaerobic ironreducing bacteria; 

some capable of 

reductive dechlorination of chlorinated 

volatile organic 

compounds)

Pelobacter

species

(Strict anaerobes that 

can oxidize organic 

substrates with ferric 

ion and sulfur as 

electron acceptors)

Desulfitobacterium

species

(Anaerobic sulfatereducing bacteria; 

may reduce heavy 

metals)

DP-8

MNA X X X

Lactate X X

DP-104

MNA X X

Lactate

DP-107

MNA X X

Lactate X

DP-6

MNA X

Lactate X
70 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

concentrations increased to 25–35 mg/L in the C bioreactor 

during this initial anaerobic period but increased to more than 

80 mg/L in the D bioreactor over the same period (figs. 46a, 

b). Methane concentrations were below detection levels in 

the C bioreactor in the few samples that were analyzed before 

donor amendments were stopped on June 14, 2012; however, 

concentrations were about 9 mg/L in the D bioreactor. Thus, 

the well-established WBC-2 biofilm in the D bioreactor apparently had a greater population density of sulfate-reducing and 

methanogenic bacteria than in the C bioreactor and rapidly 

generated highly reducing conditions, despite the high inflow 

sulfate concentrations. The similarly high sulfide and methane production in the E and D bioreactors is indicative of the 

similar WBC-2 community in each (figs. 46b, c). 

When the organic donor and other amendments to the 

C reactor were stopped between June 14 and August 24, 

2012, ORP increased to +200 mV in the C bioreactor cells by 

June 20, 2012, and then remained between + 100 and -100 

mV through August 25, 2012, indicating aerobic to mildly 

reducing conditions (fig. 44a). Sulfide and methane were not 

detectable during this period in the C bioreactor (fig. 46a). The 

D and E bioreactors maintained anaerobic conditions with the 

same organic donor amendments during the June 20–August 

24 period, but donor amendments were stopped in the D bioreactor, as well as in the C bioreactor again, for a 2-week period 

in October 2012. When donor amendments were stopped in 

October 2012, ORPs quickly increased to about +100 to +200 

mV in the C bioreactor and to about -100 to +50 mV in the D 

bioreactor. Sulfate concentrations did not decrease and sulfide 

was not produced during groundwater flow through either 

the C or D bioreactors during this 2-week aerobic period in 

October 2012.

Standard

Standard

Standard

Standard

Standard

Standard

Standard

Standard

C-CB 13

C-CB 13

DP-8

DP-8

EXPLANATION

DP-104

DP-104

DP-107

DP-107

DP-6

DP-6

C-CB 13

C-CB 13

C-CB 13

C-CB 13

C-CB 13

C-CB 13

1.00E+10

1.00E+09

1.00E+08

1.00E+07

1.00E+06

1.00E+05

1.00E+04

1.00E+03

1.00E+07

1.00E+06

1.00E+05

1.00E+04

1.00E+03

A. 2010 in situ microcosms, Total Eubacteria

Total Eubacteria, in cells per bead

Dehalococcoides species,

in cells per bead

B. 2010 in situ microcosms, Dehalococcoides

MNA Lactate WBC-2

C-DCB 13

C-DCB 13

C-DCB 13

C-DCB 13

C-DCB 13

C-DCB 13

C-DCB 13

C-DCB 13

Figure 41. Amount of (A) total Eubacteria, and (B) Dehalococcoides species determined by quantitative polymerase 

chain reaction (qPCR) analysis of samples from standard and monochlorobenzene (13C-CB) or 1,4-dichlorobenzene 

(

13C-DCB) in situ microcosms (MNA, monitored natural attenuation treatment; lactate, lactate/chitin biostimulated 

treatment; WBC-2, bioaugmented treatment) deployed in the wetland sediment at four sites, Standard Chlorine of 

Delaware, Inc. Superfund Site, Delaware, October 26–December 14, 2010.
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 71

Figure 42. Concentrations of total chlorobenzene compounds and benzene (total CBs+benzene), naturally 

input from the wetland groundwater, and amended 1,1,2,2-tetrachloroethane (TeCA) and its anaerobic daughter 

products in (A) bioreactor C, which was seeded only with native microorganisms from the site groundwater, 

Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, and in bioreactors (B) D, and (C) E, which were 

pre-established with the WBC-2 anaerobic culture.

A. Bioreactor C

B. Bioreactor D

C. Bioreactor E

3,000

3,000

5,000

7,000

6,000

6,000

5,000

4,000

3,000

2,000

1,000

5,000

4,000

3,000

2,000

1,000

4,500

4,000

3,500

3,000

2,500

2,000

1,500

1,000

500

0

0

0

2,500

2,500

2,000

2,000

1,500

1,500

1,000

1,000

14,000

12,000

10,000

8,000

6,000

4,000

2,000

0

500

500

0

0

Total CBs+benzene concentration,

in micrograms per liter

Total CBs+benzene concentration,

in micrograms per liter

Total CBs+benzene concentration,

in micrograms per liter

TeCA and TeCA daughter products

concentration, in micrograms per liter

TeCA and TeCA daughter products

concentration, in micrograms per liter

TeCA and TeCA daughter products

concentration, in micrograms per liter

stopped all

amendments

6/14/12

added

buffer

7/12/12

added buffer

+ nutrients

7/27/12

added donor

+ buffer +

nutrients

8/24/12

stopped all

donor

addition

10/5/12

stopped all

donor addition

10/5/12

restarted

donor

addition

10/19/12

restarted

donor

addition

10/19/12

3/12

3/22

2012

Tank (CBs+benzene) Tank (TeCA) Waste (TeCA daughters)

Waste (CBs+benzene) Waste (TeCA)

EXPLANATION

4/1

4/11

4/21

5/1

5/11

5/21

5/31

6/10

6/20

6/30

7/10

7/20

7/30

8/9

8/19

8/29

9/8

9/18

9/28

10/8

10/18

10/28

11/7

aerobic 6/20–8/25 aerobic

10/10–10/20

aerobic

10/10–10/20
72 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

A. Bioreactor C

B. Bioreactor D

C. Bioreactor E

9

9

9

8

8

8

7

7

7

6

6

6

5

5

5

4

4

4

3

3

3

2

2

2

pH, in standard units pH, in standard units pH, in standard units

stopped all

amendments

6/14/12

added

buffer

7/12/12

added buffer

+ nutrients

7/27/12

added donor

+ buffer +

nutrients

8/24/12

stopped all

donor

addition

10/5/12

stopped all

donor addition

10/5/12

restarted

donor

addition

10/19/12

restarted

donor addition

10/19/12

3/12

3/22

2012

CTank

DTank

ETank

C6

D6

E6

CWaste

DWaste

EWaste

C4

D4

E4

C2

D2

E2

4/1

4/11

4/21

5/1

5/11

5/21

5/31

6/10

6/20

6/30

7/10

7/20

7/30

8/9

8/19

8/29

9/8

9/18

9/28

10/8

10/18

10/28

11/7

aerobic 6/20–8/25 aerobic

10/10–10/20

aerobic

10/10–10/20

Figure 43. pH measured in (A) bioreactor C, which was seeded only with native microorganisms 

from the site groundwater, Standard Chlorine of Delaware, Inc. Superfund Site, Delaware, and in 

bioreactors (B) D, and (C) E, which were pre-established with the WBC-2 anaerobic culture.
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 73

Figure 44. Oxidation-reduction potentials (ORPs) in (A) bioreactor C, which was seeded only with 

native microorganisms from the site groundwater, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, and in bioreactors (B) D, and (C) E, which were pre-established with the WBC-2 

anaerobic culture.

A. Bioreactor C

B. Bioreactor D

C. Bioreactor E

300

300

300

200

200

200

100

100

100

0

0

0

-100

-100

-100

-200

-200

-200

-300

-300

-300

-400

-400

-400

-500

-500

-500

-600

Oxidation-reduction potential (ORP), in millivolts Oxidation-reduction potential (ORP), in millivolts Oxidation-reduction potential (ORP), in millivolts

stopped all

amendments

6/14/12

added

buffer

7/12/12

added buffer

+ nutrients

7/27/12

added donor

+ buffer +

nutrients

8/24/12

stopped all

donor addition

10/5/12

stopped all

donor addition

10/5/12

restarted

donor

addition

10/19/12

restarted

donor addition

10/19/12

3/12

3/22

2012

4/1

4/11

4/21

5/1

5/11

5/21

5/31

6/10

6/20

6/30

7/10

7/20

7/30

8/9

8/19

8/29

9/8

9/18

9/28

10/8

10/18

10/28

11/7

aerobic 6/20–8/25 aerobic

10/10–10/20

aerobic

10/10–10/20

CTank

ETank

C6

E6

CWaste

EWaste

C4

E4

C2

E2

DTank

D6

DWaste

D4

D2
74 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

A. Bioreactor C

B. Bioreactor D

C. Bioreactor E

700

700

350

600

600

300

500

500

250

400

400

200

300

300

150

200

200

100

100

100

0

0

50

0

Sulfate concentration, in milligrams per liter Sulfate concentration, in milligrams per liter Sulfate concentration, in milligrams per liter

stopped all

amendments

6/14/12 added

buffer

7/12/12

added buffer

+ nutrients

7/27/12

added donor

+ buffer +

nutrients

8/24/12

stopped all

donor addition

10/5/12

stopped all

donor addition

10/5/12

restarted

donor

addition

10/19/12

restarted

donor

addition

10/19/12

3/12

3/22

2012

4/1

4/11

4/21

5/1

5/11

5/21

5/31

6/10

6/20

6/30

7/10

7/20

7/30

8/9

8/19

8/29

9/8

9/18

9/28

10/8

10/18

10/28

11/7

aerobic 6/20–8/25 aerobic

10/10–10/20

aerobic

10/10–10/20

CTank

CWaste

C4

ETank

EWaste

E4

DTank

DWaste

D4

Figure 45. Sulfate concentrations measured in (A) bioreactor C, which was seeded only with 

native microorganisms from the site groundwater, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, and in bioreactors (B) D, and (C) E, which were pre-established with the WBC-2 

anaerobic culture.
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 75

Figure 46. Sulfate and methane concentrations measured in (A) bioreactor C, which was seeded only 

with native microorganisms from the site groundwater, Standard Chlorine of Delaware, Inc. Superfund 

Site, Delaware, and in bioreactors (B) D, and (C) E, which were pre-established with the WBC-2 

anaerobic culture.

A. Bioreactor C

B. Bioreactor D

C. Bioreactor E

40

35

30

25

20

15

10

5

160

140

120

100

80

60

40

20

14

14

14

12

12

12

10

10

10

8

8

8

6

6

6

4

4

4

2

2

2

0

0

0

90

80

70

60

50

40

30

20

10

0

0

0

Sulfide concentration, in milligrams per liter Sulfide concentration, in milligrams per liter Sulfide concentration, in milligrams per liter

Methane concentration, in milligrams per liter Methane concentration, in milligrams per liter Methane concentration, in milligrams per liter

stopped all

amendments

6/14/12

added

buffer

7/12/12

added buffer

+ nutrients

7/27/12

added donor

+ buffer +

nutrients

8/24/12

stopped all

donor

addition

10/5/12

stopped all

donor addition

10/5/12

restarted

donor

addition

10/19/12

restarted

donor

addition

10/19/12

3/12

3/22

2012

4/1

4/11

4/21

5/1

5/11

5/21

5/31

6/10

6/20

6/30

7/10

7/20

7/30

8/9

8/19

8/29

9/8

9/18

9/28

10/8

10/18

10/28

11/7

aerobic 6/20–8/25 aerobic

10/10–10/20

aerobic

10/10–10/20

ETank, Sulfide

CTank, Sulfide

DTank, Sulfide

E6, Methane

C6, Methane

D6, Methane

EWaste, Sulfide

CWaste, Sulfide

DWaste, Sulfide

E4, Sulfide

C4, Sulfide

D4, Sulfide

E2, Methane

C2, Methane

D2, Methane
76 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

The percent removals of each chlorinated benzene 

compound and benzene during transport through the bioreactors were calculated using the difference in concentrations 

measured in the tank and waste on each sampling date (fig. 42) 

and compared for the three bioreactors, separating the anaerobic (donor amended) and aerobic (donor unamended) periods 

for the C and D bioreactors (fig. 47). The calculated percent 

removals from the inflow and outflow concentrations indicated 

the amount of each contaminant degraded during the time of 

transport (residence time) through each bioreactor. Residence 

times were maintained within a relatively narrow range during 

the March–October 2012 period of operation, with median 

residence times between 37 and 41 hours in the bioreactors 

(fig. 48). The percent removals under anaerobic conditions for 

the three bioreactors showed a general pattern of decreasing 

removal rates with a decreasing number of chlorine atoms in 

the compounds, which is consistent with the relative reductive 

dechlorination rates reported for chlorinated benzenes (Field 

and Sierra-Alvarez, 2008). During the anaerobic periods, 

median percent removals of 123TCB, 124TCB, 14DCB, and 

13DCB were equally high for the native (C) bioreactor and 

WBC-2 bioaugmented (D and E) bioreactors, with median 

percent removals ranging from 86 to 94 percent (figs. 47a, b, 

c, d). The percent removal for the dichlorobenzene isomer 

12DCB, however, was substantially lower in the bioaugmented bioreactors (medians of 76 percent in bioreactors D 

and E) than in the native C bioreactor (median of 88 percent) 

(fig. 47e). This differs from previous studies of chlorinated 

benzenes that have shown 13DCB anaerobic degradation 

rates to be the slowest of the DCB isomers (Field and SierraAlvarez, 2008; Nelson and others, 2011). CB removal in the C 

and D bioreactors was similar (medians of 79 and 80 percent, 

respectively) and higher than observed in the E bioaugmented 

bioreactor (median of 73 percent) (fig. 47f). Of all the site 

contaminants in the groundwater input to the bioreactors, only 

benzene removal was substantially higher in the bioaugmented 

bioreactors (medians of 79 and 72 percent in D and E, respectively) than in the native C bioreactor (median 65 percent) 

under anaerobic conditions (fig. 47g). 

Under the aerobic/mildly reducing conditions in the C 

and D bioreactors when the organic donor amendments were 

stopped, the most substantial differences compared to strongly 

anaerobic conditions in the respective bioreactors were the 

higher percent removals in the native C bioreactor for CB 

(median 96 percent) and benzene (median 99 percent) 

(fig. 47). Slightly lower percent removals in the bioaugmented 

D bioreactor for 123TCB (median 82 percent), 124TCB 

(median 82 percent), 14DCB (median 75 percent), and 13DCB 

(median 74 percent) also were measured during this aerobic/

mildly reducing period rather than during strongly anaerobic 

conditions (fig. 47). The increased removal of CB and benzene 

from the C bioreactor under aerobic conditions actually began 

in the tank inflow concentrations, which commonly were 

nearly the same as the outflow concentrations from the bioreactor during the longer aerobic period from June 20–August 

25, 2012 (fig. 44a). The rapid removal of chlorinated benzenes 

and benzene from the tank occurred under the natural acidic 

conditions of the wetland groundwater, which was observed 

early during this period because pH buffer amendment also 

was stopped, and was also observed under the near neutral pH 

conditions that resulted when pH buffer amendment resumed 

(figs. 42, 43). These results have important implications for 

both natural attenuation in the wetland and the potential application of bioremediation—the native aerobic microorganisms 

that can degrade the chlorinated benzenes could be active in 

the wetland under the range of pH conditions observed in the 

groundwater (fig. 20), and stimulation of aerobic degradation, 

either using an in situ or ex situ engineered bioremediation 

method, would not require the addition of pH buffer. The percent removal calculations for the aerobic period were conservative because only sampling dates when detectable concentrations of VOCs remained in the tank could be used. These 

results support the evidence from the in situ microcosms and 

incorporation of 13C-CB in microbial biomass that bacteria 

capable of oxidation of CB are present in the wetland groundwater (figs. 37, 38). Biodegradation of benzene and and DCBs 

under aerobic or mildly reducing conditions also is supported 

by the bioreactor and in situ microcosm results. 

 Resumption of organic donor addition to the C bioreactor on August 24, 2012 resulted in a rapid (within a week) 

return to -300 mV ORP and production of sulfide (figs. 44a, 

46a). The prolonged aerobic period did not negatively impact 

the resumption of anaerobic degradation of the chlorinated 

benzenes and benzene by the native bacteria in the C bioreactor, indicating that an active, anaerobic degrading microbial 

community had become established in the bioreactor cells and 

was able to become active again with the addition of organic 

donor (fig. 42a). 

Degradation of the chlorinated benzenes and benzene in 

the C, D, and E bioreactors commenced without any lag and 

then improved over time under anaerobic conditions in the D 

bioreactor (figs. 42 and 49). Concentrations measured at different sampling ports in the bioreactors for selected sampling 

dates are shown in figure 49. The concentration profiles show 

that degradation shifted over time in the D bioreactor to 

principally occur in the first bioreactor cell (before sampling 

port D2 in figs. 49e-h); about 50 percent or more of each 

compound was left by port D2 on May 14, 2012, whereas 

only about 2 to 30 percent of each compound was left by port 

D2 on June 5, 2012 and later sampling dates. This improvement in biodegradation indicates that a substantial population 

of microorganisms capable of degradation of the chlorinated 

benzenes and benzene was present in the WBC-2 biofilm and 

grew over time. Similarly, degradation of chlorinated benzenes 

and benzene in the E bioreactor commenced without a lag and 

improved over time under anaerobic conditions until nearly 

the end of its operation (figs. 49i-l); the decline in removal 

near the end may have been caused by toxicity effects from 

the higher VOC concentrations in this bioreactor. Although the 

long period of aerobic conditions in the C bioreactor between 

July–August 2012 most likely slowed the development of an 

anaerobic degrading community, concentration profiles in the 
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 77

EXPLANATION

Data value within 1.5 times the IQR

outside the quartile

MAXIMUM OUTLIER (greater than

1.5 times the IQR outside the quartile)

MINIMUM OUTLIER (less than 1.5 times

the IQR outside the quartile)

75th PERCENTILE (75 percent of the

data values are less than this value)

Median

25th PERCENTILE (25 percent of the

data values are less than this value)

INTERQUARTILE

RANGE (IQR)

A

C

E

G

B

D

F

Anaerobic C

Anaerobic C

Anaerobic C

Anaerobic C

Anaerobic C

Anaerobic C

Anaerobic C

Anaerobic D

Anaerobic D

Anaerobic D

Anaerobic D

Anaerobic D

Anaerobic D

Anaerobic D

Anaerobic E

Anaerobic E

Anaerobic E

Anaerobic E

Anaerobic E

Anaerobic E

Anaerobic E

Aerobic C

Aerobic C

Aerobic C

Aerobic C

Aerobic C

Aerobic C

Aerobic C

Aerobic D

Aerobic D

Aerobic D

Aerobic D

Aerobic D

Aerobic D

Aerobic D

140

140

140

140

140

140

140

120

120

120

120

120

120

120

100

100

100

100

100

100

100

80

80

80

80

80

80

80

60

60

60

60

60

60

60

40

40

40

40

40

40

40

20

20

20

20

20

20

20

0

0

0

0

0

0

0

123TCB percent removal 14DCB percent removal 12DCB percent removal Benzene percent removal

124TCB percent removal 13DCB percent removal Chlorobenzene percent removal

Figure 47. Boxplots showing percent removals of (A) 1,2,3-trichlorobenzene (123TCB), (B) 1,2,4-trichlorobenzene (124TCB), 

(C) 1,4-dichlorobenzene (14DCB), (D) 1,3-dichlorobenzene (13DCB), (E) 1,2-dichlorobenzene (12DCB), (F) chlorobenzene, and 

(G) benzene that were calculated from inflow and outflow concentrations under anaerobic and aerobic conditions in 

bioreactors C, D, and E.
78 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

C bioreactor under the initial anaerobic conditions from May–

June 2012 indicate a slower development of the anaerobic 

degrading community than that observed in the D bioreactor 

(figs. 49a, b, e, f). About 40 to 60 percent of the chlorinated 

benzenes and benzene remained at port C2 on the June 5, 2012 

sampling date compared to the 2 to 30 percent observed in the 

D bioreactor. Thus, although native bacteria, introduced with 

the site groundwater, may have grown over time in the WBC-2 

bioaugmented bioreactors, it does not appear that they were 

primarily responsible for the observed degradation of chlorinated benzenes and benzene in the D and E bioreactors. 

The different responses of the C bioreactor to aerobic 

conditions and to the addition of TeCA compared to the 

responses of the D and E bioreactors also indicate that the 

native C bioreactor and the WBC-2 bioaugmented bioreactors 

continued to have distinct microbial populations throughout 

their operation (fig. 42). Because the well-established WBC-2 

populations in the D and E bioreactors would not be expected 

to contain microorganisms capable of degradation of chlorinated benzenes and benzene under aerobic conditions, a 

decline in the degradation of these compounds would be 

expected, and was observed. The negative impact on percent 

removals of chlorinated benzenes in the D bioreactor under 

aerobic conditions was minimal, however, probably because 

reducing conditions could be maintained in deeper layers of 

the WBC-2 biofilm during the relatively short (2-week) period 

that donor addition was stopped and ORPs increased. During 

this 2-week aerobic/mildly reducing period in October, TeCA 

removal did not change significantly in the WBC-2 bioaugmented D reactor, but it did stop rapidly in the native C bioreactor, as shown by the nearly equal concentrations of TeCA 

in the bioreactor inflow and outflow (fig. 42a). Thus, although 

degradation of the chlorinated benzenes and benzene occurred 

under both anaerobic and aerobic conditions by the native site 

bacteria in the C bioreactor, TeCA degradation required anaerobic conditions. The bioreactor results are consistent with the 

microbial community analyses that showed the presence of the 

same dechlorinating populations in the wetland sediment and 

groundwater from the site as those in WBC-2, but with lower 

population densities.

The generally greater than 70 percent removal of all 

chlorinated benzenes and benzene in the WBC-2 bioreactors 

was similar to the percent removals achieved for TeCA and 

its anaerobic degradation products, which included both 

chlorinated ethanes and chlorinated ethenes (Lorah and Olsen 

1999a; Lorah and others, 2008). The relatively high percent 

removals without the need for an acclimation period for chlorinated aromatic compounds that WBC-2 had not previously 

been exposed to during culturing or fed during maintenance 

indicates that the same dechlorinators are involved as those 

that are involved in the degradation of the chlorinated ethanes 

and ethenes. The complete anaerobic degradation of tri-, di-, 

and monochlorobenzene to benzene and further degradation of 

the benzene demonstrated in the WBC-2 bioreactors has been 

reported for only one other bioaugmentation laboratory test, in 

which a previously enriched methanogenic, benzene-degrading culture was combined with an anaerobic CB-degrading 

culture (Liang and others, 2013). The ability of the WBC-2 

culture to reduce the chlorinated VOCs in the presence of 

high sulfate concentrations (figs. 45b, c), which typically 

inhibits reductive dechlorination by acting as a competing 

electron acceptor with the chlorinated contaminants, and in the 

presence of high sulfide concentrations (figs. 46b, c), which 

typically have a toxicity effect on a wide range of bacteria 

(Maillacheruvu and Parkin, 1996; Parsons Infrastructure & 

Technology Group, Inc., 2010), is unique. The availability of 

a culture that has already been developed to a sediment-free 

form with dense populations of chlorinated aromatic dechlorinators, as well as the ability of the native wetland microbial 

community to degrade the site contaminants, provides flexibility in considering bioremediation options for the SCD wetland 

areas. The laboratory experiments for this study, however, 

were designed only to provide evidence of the occurrence 

of natural and enhanced biodegradation of the chlorinated 

benzenes at the site and to provide preliminary information on 

the degradation pathways and controlling factors. Additional 

studies would be needed to determine the best methods to 

utilize these degradation pathways in a remediation design and 

to optimize the degradation.

EXPLANATION

Data value within 1.5 times the IQR

outside the quartile

MAXIMUM OUTLIER (greater than

1.5 times the IQR outside the quartile)

75th PERCENTILE (75 percent of the

data values are less than this value)

Median

25th PERCENTILE (25 percent of the

data values are less than this value)

INTERQUARTILE

RANGE (IQR)

Bioreactor C Bioreactor D Bioreactor E

100

80

60

40

20

0

Hydraulic residence time, in hours

Figure 48. Boxplots showing hydraulic residence time in 

bioreactors C, D, and E.
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 79

Figure 49. Concentrations of chlorobenzene compounds and benzene measured under anaerobic conditions (except in 

bioreactor C on August 13, 2012) at sampling ports along flowpaths through bioreactors C, D, and E on selected sampling dates.

A. Bioreactor C, 5/14/12

C. Bioreactor C, 8/13/12 – Aerobic

B. Bioreactor C, 6/5/12

D. Bioreactor C, 9/24/12

100

100

100

100

90

90

90

90

80

80

80

80

70

70

70

70

60

60

60

60

50

50

50

50

40

40

40

40

30

30

30

30

20

20

20

20

10

10

10

10

0

0

0

0

CTank

CTank

CTank

CTank

C2

C2

C2

C2

C4

C4

C4

C4

C6

C6

C6

C6

CWaste

CWaste

CWaste

CWaste

Percent remaining Percent remaining

Percent remaining Percent remaining

EXPLANATION

1,2,4-Trichlorobenzene

1,4-Dichlorobenzene

1,3-Dichlorobenzene

1,2-Dichlorobenzene

Chlorobenzene

Benzene
80 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Figure 49. Concentrations of chlorobenzene compounds and benzene measured under anaerobic conditions (except in 

bioreactor C on August 13, 2012) at sampling ports along flowpaths through bioreactors C, D, and E on selected sampling 

dates.—Continued

E. Bioreactor D, 5/14/12

G. Bioreactor D, 8/10/12

F. Bioreactor D, 6/5/12

H. Bioreactor D, 9/4/12

100

100

100

100

90

90

90

90

80

80

80

80

70

70

70

70

60

60

60

60

50

50

50

50

40

40

40

40

30

30

30

30

20

20

20

20

10

10

10

10

0

0

0

0

DTank

DTank

DTank

DTank

D2

D2

D2

D2

D4

D4

D4

D4

D6

D6

D6

D6

DWaste

DWaste

DWaste

DWaste

Percent remaining Percent remaining

Percent remaining Percent remaining

EXPLANATION

1,2,4-Trichlorobenzene

1,4-Dichlorobenzene

1,3-Dichlorobenzene

1,2-Dichlorobenzene

Chlorobenzene

Benzene
Natural and Enhanced Biodegradation of Chlorinated Benzenes and Benzene 81

Figure 49. Concentrations of chlorobenzene compounds and benzene measured under anaerobic conditions (except in 

bioreactor C on August 13, 2012) at sampling ports along flowpaths through bioreactors C, D, and E on selected sampling 

dates.—Continued

I. Bioreactor E, 5/14/12

K. Bioreactor E, 8/10/12

J. Bioreactor E, 6/5/12

L. Bioreactor E, 9/4/12

100

100

100

100

90

90

90

90

80

80

80

80

70

70

70

70

60

60

60

60

50

50

50

50

40

40

40

40

30

30

30

30

20

20

20

20

10

10

10

10

0

0

0

0

ETank

ETank

ETank

ETank

E2

E2

E2

E2

E4

E4

E4

E4

E6

E6

E6

E6

EWaste

EWaste

EWaste

EWaste

Percent remaining Percent remaining

Percent remaining Percent remaining

EXPLANATION

1,2,4-Trichlorobenzene

1,4-Dichlorobenzene

1,3-Dichlorobenzene

1,2-Dichlorobenzene

Chlorobenzene

Benzene
82 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

Summary and Conclusions

Wetland areas north of an industrial facility that operated 

from 1966–2002 at the Standard Chlorine of Delaware, Inc. 

Superfund Site (SCD) in New Castle County, Delaware, 

are affected by contamination with chlorobenzenes and 

benzene from past waste storage and disposal, spills, leaks, 

and contaminated groundwater discharge. Prior to this study, 

the understanding of the hydrology of the SCD was limited 

to the uplands. In cooperation with the U.S. Environmental 

Protection Agency (EPA), the U.S. Geological Survey 

(USGS) began an investigation in June 2009 to provide 

a hydrogeologic and geochemical characterization of the 

wetland study area and to assess the feasibility of monitored 

natural attenuation or of combined natural attenuation and 

enhanced bioremediation as remedial strategies for the contaminated wetland areas. This assessment included a field and 

experimental approach, with measurement of water levels, 

collection of groundwater samples using passive diffusion 

samplers and piezometers, in situ microcosms deployed 

directly in the wetland sediment, microbial community analyses, and tests with flow-through anaerobic bioreactors seeded 

with either the native microbes or a bioaugmented dechlorinating culture, WBC-2.

Wetland sediments at SCD are present in the low-lying 

fringe of Red Lion Creek, with less than 1 meter (m) of topographic variation from the boundary of the wetland to open 

water on Red Lion Creek. Wetland sediments at SCD can be 

grouped into two stratigraphic units—an upper organic rich 

unit consisting of root mass and organic matter in a mineral 

matrix of silt and (or) clay, and a basal unit consisting of clay, 

silt and fine-grained sand. Wetland sediments generally are 

between 5 and 10 m thick in the study area but are estimated 

to extend to greater than 10 m at one site. Further data collection would help to define the interface between wetland 

sediments and the underlying Columbia aquifer. Groundwater 

flow in the wetland area is predominantly vertically upward in 

the wetland sediments and the Columbia aquifer, with lateral 

velocities less than half the estimated vertical flow velocities. 

Groundwater discharge accounts for an estimated minimum 

of 47 percent of the total discharge for the subwatershed of 

tidal Red Lion Creek at the site. Thus, groundwater transport 

to shallow creek sediments and surface water is most likely a 

significant transport mechanism for contaminants to ecological 

receptors, requiring treatment of both sediments and groundwater in the wetland and Red Lion Creek.

 The major contaminants detected in groundwater in 

the wetland study area included benzene and six chlorinated 

benzenes—1,2,3- and 1,2,4-trichlorobenzene (123TCB and 

124TCB); 1,2-, 1,3-, and 1,4-dichlorobenzene (12DCB, 

13DCB, and 14DCB); and chlorobenzene (CB). Shallow 

wetland groundwater in the northwest area was characterized 

by high concentrations of total chlorinated benzenes and 

benzene (with a maximum concentration of about 75,000 

micrograms per liter or μg/L), low pH (median 4.5 to 5.2), 

high chloride concentrations, and low or undetectable sulfate 

concentrations. In the northeast area, wetland groundwater had 

low to moderate concentrations of total chlorinated benzenes 

and benzene (generally not greater than 10,000 μg/L), moderate pH (median 5.8 to 6.1), and high sulfate concentrations 

(200–1,000 milligrams per liter or mg/L) as a result of a 

sulfuric acid spill. Concentrations in excess of 1 percent of the 

solubility of the individual chlorinated benzenes remain in the 

groundwater in the wetland sediments and aquifer, indicating 

that sources of dense nonaqueous phase liquids (DNAPLs) 

remain outside the subsurface barrier constructed around the 

main industrial facility. The consistently higher contaminant 

concentrations detected in the shallow wetland groundwater 

than in groundwater deeper in the wetland sediments or the 

Columbia aquifer also indicate a continued source in the 

wetland sediments, either as DNAPLs or compounds sorbed 

to the sediments, or both. Decreasing chloride concentrations 

in the shallow groundwater in the northwest wetland area from 

2009–11 may have resulted from an interruption of a contaminant source from behind the barrier wall after its construction 

in 2007; however, volatile organic compound (VOC) concentrations indicative of DNAPL presence remain outside the 

barrier wall. 

When highly reducing conditions (methanogenic or sulfate-reducing) existed in the wetland groundwater, molar ratios 

of the VOCs showed that CB and benzene were predominant, 

compared to higher percentages of TCBs and DCBs under 

weaker reducing conditions, and indicated that biodegradation 

of the chlorinated benzenes was occurring through reductive 

dechlorination pathways. In July 2009, reductive dechlorination was evident in the groundwater in the northwest wetland 

area but not the northeast area; however, a change in redox 

conditions in the wetland reversed this pattern, with groundwater in the northeast wetland area showing greater evidence 

of reductive dechlorination in 2011–12. Total VOC concentrations did not decrease substantially along upward groundwater flowpaths in the wetland sediment, even when the molar 

compositions indicated substantial reductive dechlorination, 

most likely due to the additional source of VOCs in the upper 

wetland sediments. Natural attenuation alone, therefore, is not 

sufficient to remove the VOCs from the wetland groundwater 

where these sources are present in the wetland sediment.

The changes in the wetland geochemistry observed 

between 2009 and 2011 were most likely due to alterations 

in groundwater flow and geochemistry from the Columbia 

aquifer after the installation of the barrier wall in 2007, 

although induced changes from industrial pumping and tidal 

fluctuations with an inoperative tide control structure also may 

have been contributing factors. The addition of selected redox 

constituents and chloride to the current monitoring plan for 

the upland area near the wetland boundary, and the inclusion 

of piezometers within the wetland area to this monitoring 

plan would assist in determining the effect of the barrier 

wall on the wetland geochemistry, contaminant sources, and 

biodegradation. 
Acknowledgments 83

Unamended and WBC-2 bioaugmented microcosms 

(deployed during 2009 and 2010 at two to four sites in the 

wetland sediments) confirmed that the native microbial community and WBC-2, respectively, were able to biodegrade the 

higher chlorinated benzenes through reductive dechlorination. 

Both the unamended and WBC-2 treatments showed incorporation of 13C-labeled CB into dissolved inorganic carbon 

(predominantly carbon dioxide), which can be produced as 

an end product through aerobic or anaerobic degradation 

pathways and indicates use of the contaminant for cellular 

energy. Only the unamended treatments with 13C-labeled 

CB, however, showed incorporation of 13C into the microbial 

biomass (measured as the amount of phospholipid fatty acids), 

indicating that microorganisms were utilizing the carbon as 

a source for cellular material through oxidative degradation 

pathways. This apparently aerobic degradation pathway for 

CB was evident in all the unamended microcosms under a 

wide range of measured redox conditions. Natural, complete 

breakdown of 13C-labeled 14DCB and benzene to carbon 

dioxide was also measured in unamended microcosms. The 

relative amounts of 13C incorporation in carbon dioxide in 

these in situ microcosms indicated that dichlorobenzene 

degradation was slowest and benzene degradation was fastest 

by the native microbial community, which is consistent with 

the expected order of degradation rates by aerobic oxidation 

pathways. Microbial community analyses of samples from the 

in situ microcosms and from the wetland sediment, however, 

showed relatively high cell numbers (103 to 105 cells per bead) 

and diversity of populations (Dehalococcoides, Dehalobacter, 

Desulfitobacterium, and Geobacter) that are known to contain 

species capable of reductive dechlorination of VOCs. Thus, a 

high natural reductive dechlorination potential at the site also 

was apparent, consistent with evidence from molar composition of the contaminants observed under highly reducing 

conditions in the wetland groundwater. 

Experiments with a continuous flow, fixed-film bioreactor 

seeded with native microorganisms in groundwater from the 

wetland area showed both aerobic and anaerobic biodegradation of dichlorobenzenes, monochlorobenzene, and benzene, 

although monochlorobenzene and benzene degradation rates 

decreased under anaerobic conditions compared to aerobic 

conditions. In two bioreactors with established biofilms of 

WBC-2 that also operated with inflow of site groundwater, 

generally greater than 70 percent removal of all chlorinated 

benzene compounds and benzene were removed under mixed 

sulfate-reducing and mixed sulfate-reducing/methanogenic 

conditions. During anaerobic operation, median percent 

removals of all chlorinated benzene compounds were equally 

high for the native and WBC-2 bioaugmented bioreactors, 

with median percent removals ranging from 86 to 94 percent. 

Benzene removal was greater in the WBC-2 bioaugmented 

bioreactors than in the bioreactor seeded only with site 

groundwater. The relatively high percent removals without the 

need for an acclimation period for the chlorinated aromatic 

compounds that WBC-2 had not previously been exposed 

to during culturing or maintenance indicates that the same 

dechlorinators are involved as those in the degradation of the 

chlorinated ethanes and ethenes that the culture was developed 

to degrade. The ability of the WBC-2 culture to completely 

reduce the chlorinated benzenes and benzene, even in the presence of high sulfate concentrations and high sulfide concentrations, is unique for known bioaugmentation cultures. 

Both field and laboratory investigations provide evidence 

that biodegradation is active in the wetland and can be 

enhanced under anaerobic conditions by bioaugmentation with 

the dechlorinating culture WBC-2. The availability of a culture 

that has already been developed to a sediment-free form with 

dense populations of dechlorinators, as well as the ability of 

the native wetland microbial community to degrade the site 

contaminants, provides a range of possible bioremediation 

options for the wetland areas at the SCD. Further investigation and testing would be needed, however, to evaluate the 

most appropriate bioremediation methods, or combination of 

methods, for implementation in the different wetland areas. 

The laboratory and field tests for the current study were 

designed only to provide preliminary information on degradation pathways and controlling factors. Additional studies 

would be needed to determine the best methods to utilize these 

degradation pathways in a remediation design and to optimize 

the degradation, taking into consideration the continued source 

of contaminants in areas of the wetland and aquifer in their 

DNAPL form.

Acknowledgments

The authors thank the following agencies and people for 

their contributions to this study. EPA, Region III, provided 

management and logistical support; the efforts of Hilary 

Thornton and Brad White are gratefully acknowledged. The 

authors thank Duane Graves with Geosyntec Consultants 

(Knoxville, Tennessee) and SiREM Laboratory (Guelph, 

Ontario, Canada) for their continued maintenance of WBC-2 

and support under our USGS-Geosyntec Cooperative Research 

and Development Agreement (CRADA). WBC-2 culture 

and anaerobic medium used in all laboratory and field tests 

reported here were provided by SiREM with the assistance 

of Jennifer Webb. The following USGS personnel also are 

thanked for laboratory and field assistance throughout this 

study: Mastin Mount and Melody Flinchbaugh for VOC and 

methane analyses; Roberto Cruz, Luke Myers, and Jessica 

Thompson for hydrogeologic measurements, field and 

experiment sampling, and database support. We also thank 

Andrew Reid (currently at the University of Rhode Island) and 

Elizabeth Shipley for assistance with sampling preparation, 

laboratory analyses, and data management. Special thanks are 

extended to Jeanne Jaeschke (USGS) for providing chemical 

analyses of major ions and dissolved organic carbon. 
84 Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12

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Edited by Valerie M. Gaine.

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Lorah, M.M., and others—Hydrogeologic Characterization and Bioremediation of Chlorinated Benzenes in Wetland Areas, SCD Superfund Site, Del., 2009–12—Scientific Investigations Report 2014–5140

ISSN 2328-031X (print)

ISSN 2328-0328 (online)

http://dx.doi.org/10.3133/sir20145140 Printed on recycled paper