Office of Researc h and Development

National Risk Management Research Laboratory 

Methods and Tools for the 

Evaluation of Monitored 

Natural Recovery of 

Contaminated Sediments: 

Lake Hartwell Case Study 
EPA/600/S-10/006/September 2010/www.epa.gov/research 

Methods and Tools for the 

Evaluation of Monitored 

Contaminated Sediments: 

Lake Hartwell Case Study 

Office of Research and Development

National Risk Management Research Laboratory 
Contents
 

Purpose.................................................................................................................................................. 1
 

Introduction........................................................................................................................................... 1
 

Site Background: Lake Hartwell/Twelvemile Creek ............................................................................ 2
 

Research Study Phases.......................................................................................................................... 2
 

Methods to Assess MNR Processes ...................................................................................................... 3
 

Chemical Lines of Evidence ............................................................................................................. 4
 

Biological Lines of Evidence............................................................................................................ 6
 

Physical Lines of Evidence............................................................................................................. 12
 

Conclusions......................................................................................................................................... 17
 

References........................................................................................................................................... 19
 

i 
 

1.0 

Purpose and Introduction
 

Purpose 

The National Risk Management Research Laboratory 

(NRMRL) of the U.S. Environmental Protection 

Agency’s (U.S. EPA’s) Office of Research and 

Development (ORD) has been conducting research 

to develop methods and tools for the evaluation of 

monitored natural recovery (MNR) of sediments 

contaminated with polychlorinated biphenyls 

(PCBs), polycyclic aromatic hydrocarbons (PAHs), 

mercury, and other legacy pollutants. This research is 

supporting a broad, national research program focused 

on contaminated sediments in U.S. waterways. This 

Research Summary provides a synopsis of a multiyear, interdisciplinary research project conducted by 

ORD, specifically NRMRL and the National Exposure 

Research Laboratory (NERL), at the SangamoWeston, Inc./Twelve-Mile Creek/Lake Hartwell PCB 

Contamination Superfund Site in Pickens County, 

South Carolina. The methods and tools described in 

this Research Summary are comprised of quantitative 

approaches for characterizing naturally-occurring 

mechanistic processes that are necessary to manage risk 

using MNR. 

The information developed in this project is expected 

to be used as a reference for site managers and Federal, 

State, and local regulators who may be considering 

MNR as a site remedy or monitoring the progress of 

MNR at a contaminated sediment site(s). The methods 

and techniques evaluated and optimized in this study 

also provide the broader sediment community with an 

approach for characterizing environmental processes 

controlling the risk associated with contaminated 

sediments. 

Introduction 

Sediments are often the ultimate receptors of 

contaminants in aquatic systems. According to an 

estimate by U.S. EPA, approximately 10% (~1.2 billion 

cubic yards) of the sediment underlying the country’s 

surface water is sufficiently contaminated to pose 

potential risks to humans and wildlife (USEPA, 1998). 

To manage the risk imposed by these sediments, it is 

critical to understand the fundamental mechanisms 

responsible for contaminant transport and fate. These 

mechanisms include chemical, biological, and physical 

processes responsible for the risk associated with the 

sediment. Metrics or approaches to characterize the 

effects of sediment remediation are needed to provide 

quantitative measures of success. 

MNR is one approach for managing the risk associated 

with contaminated sediments. With MNR, the 

contaminants are left in place and existing physical, 

chemical, and biological processes that are expected to 

contain, destroy, and/or reduce their bioavailability or 

toxicity are monitored to assure effective management 

is progressing as predicted. These processes must 

be demonstrated to be viable, and occurring at an 

acceptable rate, to attenuate the risk of the contaminant 

remaining in place. Similarly, an effective and 

quantitative monitoring plan must be able to demonstrate 

definitively that these processes are progressing as 

predicted and that risk is being reduced over time at the 

anticipated rate. 

MNR relies on a few primary mechanisms for mitigating 

contaminant risk to human and ecological receptors. 

First and foremost is a reduction in the contaminant’s 

availability to receptors. This reduction can be 

accomplished either through a physical process of 

isolating the contaminated sediment itself or through 

a chemical process of contaminant sequestration or 

sorption. Physical isolation typically results from 

deposition of sediments not contaminated with the 

compound-of-concern (COC). Generally, exposure to 

contaminated sediment is expected to occur either in the 

water column during transport or, more commonly, at the 

water/sediment interface once deposited. Depending on 

the water body, contaminant exposure in the near surface 

sediment would be expected to occur in the biologically 

active zone of the sediment. This biologically active 

layer can vary from as little as a 1-3 cm to as much as 

1-3 m in thickness depending on the habitat (Clarke, 

2001.). The deposition of sediment, defined as 

accretion, is a natural on-going process in most water 

bodies. If the source of the COC has been effectively 

managed (a term often referred to as source control), the 

contaminated sediment will become physically isolated 

from the primary receptors in the aquatic system. 

However, sediment accretion rates vary spatially and 

temporally, are dependent on the hydrologic and physical 

conditions at the site, and are subject to change with 

alterations within the watershed or in the site conditions. 

MNR requires confirmation that a site continues to be a 

deposition-dominated site over time without significant 

mixing of the near-surface sediment. 

1 
Another important mechanism in MNR is often the 

chemical sequestration, or sorption, of the contaminant to 

the sediment matrix. With organic COCs, this sorption 

process is generally most favorable with hydrophobic, 

non-volatile contaminants binding to organic matter 

within the sediment. With inorganic COCs and less, 

commonly, organic COCs, the binding of the contaminant 

may occur with the inorganic clay and mineral structures 

of the sediment. In either instance, this sequestration 

or sorption of the contaminant has been shown, in some 

cases, to be effective in reducing the ability of a receptor 

(human or ecological) to take up the contaminant if 

exposed. This reduced exposure may reduce the risk of 

the COC. However, studies have shown the sequestration 

of contaminants, whether organic or inorganic, and any 

reduced availability are related to site-specific conditions 

(sediment type, organic matter content and structure, 

oxidation/reduction state, presence of sulfide, etc.). 

These conditions, therefore, should be evaluated on a 

site-specific basis during initial design for MNR and 

should be incorporated into the long-term monitoring 

plan to assure site conditions remain viable for continued 

reductions in contaminant availability (Magar, 2009). 

The biological degradation, or biodegradation, of the 

COC should also be evaluated where considering 

MNR. If these processes are determined to be important 

mechanisms for either changing the availability or 

toxicity of the COC or reducing the inventory of the 

COC, biodegradation evaluations should be included as 

part of the long-term monitoring plan. Biodegradation is 

dependent on the physical/chemical characteristics of the 

contaminant itself and the viability of microfauna capable 

of biodegrading the COC. Biodegradation can reduce or 

increase the toxicity of COCs depending on the biological 

process and/or the specific COC. Finally, site conditions 

can inhibit or control the environment needed for optimal 

biodegradation rates. 

Site Background 

In 1994, U.S. EPA Region 4 issued a Record of Decision 

(ROD) (USEPA, 1994) establishing MNR as the remedyof-choice to remediate Operable Unit #2 (OU2) of the 

Sangamo-Weston, Inc./Twelve-Mile Creek/Lake Hartwell 

PCB Contamination Superfund Site (Figure 1). OU2 

consisted of the contaminated sediments in Lake Hartwell 

and Twelve-Mile Creek (TMC). OU2 was used as a 

field location by U.S. EPA-ORD to develop, test, and 

validate methods and tools to support the evaluation of 

MNR of PCB-contaminated sediments. A comprehensive 

set of measurements and analyses were conducted over 

a number of years to develop chemical, biological, and 

physical lines of evidence to document the progress of the 

MNR remedy in reducing PCB concentrations in surface 

sediments and the associated ecosystem. 

Sangamo, Inc. owned and operated a capacitor 

manufacturing plant in Pickens, South Carolina, from 

1955 to 1987. During operations, the facility used 

several varieties of fluids that contained PCBs. Waste 

was disposed of on the plant site and at six satellite 

disposal areas within a 3-mile radius of the plant. PCBs 

were also discharged with industrial wastewater effluent 

into Town Creek, a tributary of TMC. TMC is a major 

tributary of Lake Hartwell. Between 1955 and 1977, 

the average quantity of PCBs used by Sangamo-Weston 

ranged from 700,000 to 2,000,000 lb/yr. An estimated 

3% of the quantities received and used by the plant were 

discharged into Town Creek, resulting in an estimated 

cumulative discharge of 400,000 lb of PCBs. PCB use 

was terminated in 1977, prior to a U.S. EPA ban on its 

use in January 1978 (USEPA 1994). 

Lake Hartwell is a U.S. Army Corps of Engineers 

(USACE) reservoir located in the northwest corner of 

South Carolina along the Georgia state line (Figure 1). It 

is bordered by Anderson, Pickens, and Oconee Counties 

in South Carolina and Stephens, Franklin, and Hart 

counties in Georgia. It was created between 1955 and 

1963 when the USACE constructed Hartwell Dam on the 

upper Savannah River, 7 miles from the confluence of 

the Seneca and Tugaloo Rivers. Lake Hartwell extends 

49 and 45 miles up the Tugaloo and Seneca Rivers, 

respectively. At full pool elevation (660 ft MSL), it 

covers nearly 56,000 acres of water with a shoreline of 

962 miles (USEPA, 1994). 

Research Study Phases 

U.S. EPA-ORD conducted research at the Lake Hartwell/ 

TMC site between 2000 and 2008. This research evolved 

in phases as the study progressed and more information 

was required. 

Phases I and II of the project were completed in 2000­

2002 and consisted of chemical and age dating analyses 

of sediment core samples collected longitudinally and 

laterally on 10 transects previously established by U.S. 

EPA Region 4 for ongoing monitoring at Lake Hartwell 

(USEPA, 1994; USEPA, 2004). These studies focused 

on the use of chemical analyses of surface sediments 

and sediments at depth to characterize the decline of 

surface sediment concentrations over time (Brenner, 

Magar et al., 2004). Additionally, evidence of anaerobic 

dechlorination of higher-chlorinated PCBs was shown to 

be occurring at depth in the sediment, albeit, at very slow 

rates (Magar, Johnson et al., 2005). 

Results from Phases I and II along with the required ROD 

annual monitoring data (USEPA, 2004) were evaluated 

2 
in developing the subsequent studies at Lake Hartwell 

and TMC. Surface sediment PCB concentrations were 

declining at predicted rates; however, reductions in 

fish tissue did not appear to be declining similarly as 

projected. The result was an expansion of the project 

within ORD that encompassed Phases III-V over the 

next 6 years. 

Phase III (2002-2003) served as a proof-of-concept effort 

in the development of innovative tools for monitoring 

the mechanisms responsible for the transport and fate 

of PCBs in this system. Measures during Phase III 

focused on the elements that comprise the Lake Hartwell 

and TMC ecosystem including environmental matrices 

(sediment, benthos, and water column) and food web 

components (fish, macroinvertebrates, plankton, and 

organic matter). Additional studies concentrated on the 

diffusion and advection of PCBs from contaminated 

sediment/pore water into the water column, as well as 

surrogate biological measures using passive samplers. 

Phase IV (2004-2005) and V (2005-2008) of these 

studies focused on extending selected data sets for 

characterizing long-term trends and further validating 

selected methods and tools from prior studies at Lake 

Hartwell/TMC and other sites. Phase IV focused on 

characterizing seasonal variability of the mechanisms 

monitored. Phase V consisted of further food web and 

biotic studies along with studies of surface sediment 

stability and mixing. 

Methods to Assess MNR Processes 

MNR relies on several primary mechanisms to exist 

and progress at a predictable rate. These mechanisms 

generally involve naturally-occurring chemical, 

biological, and physical processes. Often, these 

processes are unrelated mechanistically with one another 

and may occur at different temporal and spatial scales. 

As such, multiple lines of evidence may be required 

to demonstrate success or progress of MNR toward 

the remedial objectives or goals. These site-specific 

lines of evidence are broadly categorized as chemical, 

biological, and physical lines of evidence. Through a 

Figure 1. Location of Lake Hartwell and Twelvemile Creek Site. 

3 
Figure 2. Conceptual Site Model for Monitored Natural Recovery as applied 

to Lake Hartwell (adapted from(Magar 2009). 

weight-of-evidence approach, the progress of MNR can 

be evaluated. 

Conceptual Site Model 

Prior to evaluating a remedy for a contaminated 

sediment site, a conceptual site model (CSM) should 

be developed to identify the key processes affecting the 

transport and fate of the contaminant and receptors at 

risk. Once the designed remedy has been completed, 

verification of remediation effectiveness is required. 

MNR may require long-term monitoring to verify that 

the naturally-occurring processes necessary to manage 

the contaminant risk are proceeding as predicted. The 

CSM again should be used to assist in the choice of 

critical sample matrices and the periodicity required 

to demonstrate progress. Figure 2 shows the CSM 

generated for Lake Hartwell to aid in the development of 

monitoring tools for documenting the lines of evidence 

needed to characterize the progress of the remedy. This 

CSM was utilized by U.S. EPA researchers to guide 

the development of sampling tools and methods briefly 

described in this Research Summary document. 

Chemical Lines of Evidence 

In most cases, direct measurement of the COC in surface 

sediments is necessary to document its reduction over 

time. At Lake Hartwell, surface sediment was defined 

as the top 10 cm of the sediment. This depth was 

established in the Remedial Investigation /Feasibility 

(RI/FS) study and was followed throughout this research. 

Surface sediment samples were collected for all phases 

of the research in Lake Hartwell and, where available, 

in TMC. Direct measures of the COC are a primary 

line of evidence in MNR, and several approaches were 

evaluated at Lake Hartwell. 

4 
Sediment Concentrations and Age Dating. 

Surface sediment concentrations were measured 

over the duration of the studies and combined 

with the data on sediment concentrations for 

samples collected under the ROD. In addition,

core sampling was completed at selected

transects to characterize the deeper sediments 

that reflect changes in historic surface sediment 

concentrations over time (Brenner, Magar et al. 

2004). Results indicated that age dating 

sediments can provide a historic record of the 

surface sediment concentrations and, in this

case, demonstrated a trend toward reduced 

concentrations of total PCBs (t-PCBs) over time 

(Figure 3). Researchers provided predictions of 

recovery time to reach human and ecologically

significant surface sediment concentrations 

based on historic sediment deposition rates 

Figure 3. Total PCB (t-PCB) Vertical Concentration Profile in Core L with Age Dated Sediment. 

Table 1. Required sediment deposition (in cm) to achieve the 1.0 mg/kg t-PCB surface sediment 

 

0-5 cm 

5-10 cm
	

10-15 cm
	

15-20 cm
	

20-25 cm
	

25-30 cm
	

30-35 cm
	

35-40 cm
	

40-45 cm
	

45-50 cm
	

50-55 cm
	

55-60 cm
	

60-65 cm
	

65-70 cm
	

70-75 cm
	

75-80 cm
	

80-85 cm
	

Depth (cm) 

, 

0 10 20 30 40 50 60 

Concentration (mg/kg dry weight) 

2000 

1999 

1998 

1996 

1995 

1994 

1992 

1991 

1989 

1987 

1986 

1984 

1982 

Year 

When conducting a site investigation, it is important to 

ask: “What defines surface sediment?” The definition 

of surface sediment can be based on the depth of 

expected mixing due to hydrological events (high flow 

condition, propeller wash, etc.) and the biological 

active zone (BAZ) (MacDonald, Ingersoll et al., 2000; 

Eggleton and Thomas, 2004). The BAZ is inhabited 

by infaunal organisms including microbes, meiofauna, 

macroinvertebrates, and other organisms (i.e., early life 

stages of fish or amphibians) that spend all or part of 

their lives associated with sediments. The community 

of organisms present generally depends on the physical 

and chemical characteristics of the water body as 

determined by the watershed. The depth of the BAZ 

varies depending on sediment substrate characteristics 

(including particle size fractions, organic matter 

content, consolidation, and pore water geochemistry). 

These characteristics control the organisms present. In 

freshwater systems, the BAZ typically spans the top 20 

to 40 cm of the surface sediment (Clarke, 2001; Nogaro, 

Mermillod-Blondin et al., 2009). The majority of benthic 

organisms will usually be associated with the upper 15 

cm. However, certain invertebrate and/or amphibian 

species can extend the BAZ during a portion of their life 

history (e.g., up to 100 cm below the sediment surface) 

(Fleeger, Tita et al., 2006). Understanding the depth of 

the BAZ is critical in designing sediment and biological 

sampling plans for MNR sites. 

Sediment Concentrations and Age Dating. 

Surface sediment concentrations were measured over 

the duration of the studies and combined with the data 

on sediment concentrations for samples collected under 

the ROD. In addition, core sampling was completed at 

selected transects to characterize the deeper sediments 

that reflect changes in historic surface sediment 

concentrations over time (Brenner, Magar et al., 2004). 

Results indicated that age dating sediments can provide 

a historic record of the surface sediment concentrations 

and, in this case, demonstrated a trend toward reduced 

concentrations of total PCBs (t-PCBs) over time (Figure 

3). Researchers provided predictions of recovery time 

to reach human and ecologically significant surface 

sediment concentrations based on historic sediment 

deposition rates (Table 1). In addition, researchers also 

Figure 3. Total PCB (t-PCB) Vertical Concentration Profile in Core L with Age Dated Sediment. 

5 
Table 1. Required Sediment Deposition (in cm) to Achieve Selected t-PCB Surface Sediment Concentration Goals. 

evaluated reductive dechlorination of higher-molecularweight PCB congeners occurring at Lake Hartwell 

(Magar, Johnson et al., 2005). 

Passive Samplers. Researchers also developed and 

evaluated novel passive sampler designs utilizing 

semipermeable membrane devices (SPMDs) at three 

sites in Lake Hartwell, a background transect (BKG) 

with a low level of contamination and two contaminated 

transects (T-M/N and T-O) on Lake Hartwell. PCB 

availability was quantified using racks (RSPMDs) 

designed to hold SPMDs in contact with surface 

sediments, benthic domes (DSPMDs) designed to enclose 

and suspend SPMDs at the sediment-water interface, 

and commercially obtained SPMD cages suspended in 

the water column (WCSPMDs). The sampler designs 

and results from preliminary studies are described by 

Schubauer-Berigan et al. (Schubauer-Berigan, 2010 

-submitted). For a detailed review of the underlying 

principals of SPMDs, see (Gale, 1998; Booij, Smedes et 

al., 2006; Huckins, 2006). 

The above devices developed and tested at Lake 

Hartwell proved to be effective in providing reproducible 

quantitative estimates of the dissolved PCB fraction at 

the study areas and in the different spatial arrangements 

deployed (Table 2 and Figure 4). Measurable 

accumulation of PCBs was detected in all SPMD 

deployments above contaminated sediment levels in 

as little as 7 to 14 days. Time-weighted average total 

PCB concentrations were highly reproducible for field 

deployments, and variability was low within sampler 

type and study area (Table 2). A two-way analysis of 

variance showed significant effects by sampler type (P 

< 0.0001) and site (P < 0.0001) for mean time-weighted 

average (TWA) total PCBs (t-PCBs). Regardless of the 

SPMD sampler used, mean TWA t-PCBs were directly 

proportional to the level of contamination present 

(BKG < T-M/N < T-O). WCSPMDs and RSPMDs 

t-PCB uptake patterns were described by a significant 

exponential fit of the data (r = 0.9768, p = 0.000006 

and r = 0.9114, p = 0.0006, respectively [Figure 4]). 

DSPMD and RSPMD samples had a higher percentage of 

higher-chlorinated PCBs compared to WSPMD samples. 

The SPMD samplers tested were particularly useful for 

quantifying spatial differences in the partitioning of PCB 

exposure within the water column and sediments. 

The SPMD deployment approaches designed and tested 

in this study demonstrated promise as a rapid and robust 

method to aid in developing contaminated sediment risk 

management strategies and assessing the mechanisms 

influencing the effectiveness of these strategies. Some 

of the benefits of using passive samplers in addition 

to biological monitoring are that they can be deployed 

in sufficient numbers to help overcome issues of both 

spatial and temporal sampling variability and the data 

from one geographic area to another can be compared 

directly. Also, passive samplers can be used in harsh 

environmental conditions that would be lethal to caged 

organisms. 

Biological Lines of Evidence 

The biological metrics of exposure can consist of 

biota impacted by the COC or other indicators used to 

6 
demonstrate exposure or effects. These metrics can use 

organisms that range from fish that have both ecological 

and human health risks to benthic infauna that can 

demonstrate both a direct impact of the COC or a link to 

higher trophic levels that are impacted. 

In reviewing ROD-mandated monitoring data for Lake 

Hartwell, two issues became evident: 1) sampling longerlived fish resulted in a bias based on gender, and 2) data 

appeared to indicate that persistent organic pollutants, 

like PCBs, may require as long as two generations of the 

targeted fish species (10-16 years) to show statistically 

significant changes in tissue concentrations. As a result 

of these observations, biological indicators of exposure 

were evaluated at Lake Hartwell that have shorter life 

spans and, therefore, were hypothesized to reflect shorter 

temporal responses to PCB exposure. 

Macrobenthos Body Burden. One indicator 

that was adapted to evaluate the long-term recovery 

of contaminated sediments and the associated biota 

Table 2. Mean Time-Weighted Average Concentrations of Total Dissolved PCBs as Measured Using Different SPMD 

Samplers (water column: WCSPMD; sediment surface rack: RSPMD; and submerged benthic dome: DSPMD) at 

Three Lake Hartwell Sites (Schubauer-Berigan 2010 - submitted). 

7 
Figure 4. Regression Analyses (using an exponential fit) of the Time Course 

Experiments Comparing WCSPMDs and RSPMDs at Site T-O (top panel) and a 

Comparison of the RSPMDs Deployed at T-O, T-M/N, and BKG (bottom panel) 

(Schubauer-Berigan, 2010 - submitted). 

8 
following remediation was the measurement of COC 

body burden in aquatic macrobenthos. Researchers 

utilized artificial substrates (Hester-Dendy substrates) to 

collect macrobenthos for harvesting and quantifying the 

COC body burden. 

Aquatic macrobenthic organisms were examined as an 

indicator of recent exposure to PCBs. A majority of the 

benthic invertebrates, such as midge larvae, annelids 

(aquatic worms), and other larvae, have life cycles 

that last 30-90 days (Pennak, 1978; Merritt, 2008). 

Contaminant tissue levels, referred to as body burden 

concentrations, in macroinvertebrates represent very 

recent contaminant exposure levels. At Lake Hartwell, 

the approach was first to determine if macroinvertebrates 

collected on artificial substrates, such as Hester Dendy’s 

(HDs) (Klemm 1990) could discriminate between 

various contaminated sites and compare the body burden 

concentrations to sediment concentrations. In 2002, 

20 nine-plate HDs were attached in four tiers of five 

to inverted wire cages. After 4 weeks of deployment 

in the water column, HDs were retrieved, dissembled, 

sorted, preserved (4oC), and analyzed for total PCBs by 

summing PCB congeners as total PCBs (t-PCBs). Figure 

5 compares t-PCB concentration in water, sediment, and 

macroinvertebrate tissue. Similar trends were noted 

in sediment t-PCB concentrations for macrobenthos 

body burden between sites, but some differences in the 

magnitude of PCBs were evident at site T-M/N. 

These results indicate that macroinvertebrates can be used 

to assess sediment contamination among sites that have 

different contamination levels of PCBs, but additional 

validation over time is needed to demonstrate a more 

Figure 5. t-PCB Concentrations in Water (ug/L or ppb) and Sediment and Macroinvertebrate 

Tissue (ng/g wet weight or ppb). 

9 
Figure 6. t-PCB Wet Weight Tissue (ng/g or ppb) for Two Indicators Used at Lake Hartwell: 

Collections of Indigenous SSLAF, Whitefin Shiners (Cyprinella nivea) and 14-day Deployed 

Adult (8-12 months old) Fathead Minnows (Pimephales promelas). 

rapid response to changes in sediment concentrations than 

measured in higher trophic level fish. 

Small Short-Lived Adult Fish (SSLAF). SSLAF 

were evaluated as an indicator of shorter-term responses 

within the indigenous fish populations as a short- and 

a long-term monitoring tool. As opposed to long-lived 

larger fish, SSLAF are expected to exhibit more rapid 

response to changes in the COC exposure in the system. 

The Cyprinidae (minnow family) is the largest of all 

fish families. Minnows are very important components 

within the food web because they are food to larger sport 

fish (e.g., largemouth bass, striped bass, etc.). There are 

several species of minnows and shiners across North 

America that are omnivores preying on insects, algae, 

detritus, and microcrustaceans. These fish reach maturity 

in 4-6 months and live 1-3 years. Many have good 

site fidelity (do not forage or migrate great distances) 

and may be excellent indicators of recent (1-3 years) 

contaminant exposures in sediment and water. Therefore, 

SSLAF should be a good biological indicator to track 

exposure changes in shorter periods of time and space 

than do longer-lived fish. 

Figure 6 illustrates the wet weight tissue concentrations 

measured in two SSLAF indicators at Lake Hartwell: 

indigenous whitefin shiners, Cyprinella nivea, and 14­

day deployed adult (8-12 months old) fathead minnows, 

Pimephales promelas. Whitefin shiners and 14-day 

deployed adult fathead minnows showed similar trends 

in tissue concentrations to the macroinvertebrates and 

related well to sediment t-PCB concentrations (i.e., 

higher tissue concentrations were measured at T-O than at 

T-M/N as reflected in the sediment concentrations). 

Additional studies are being conducted to compare results 

of SSLAF to traditionally top carnivore, longer-lived 

adult fish, as are typically used for long-term monitoring. 

This shorter-term indicator may provide more rapid 

indications of improvements in sediment contaminant 

concentrations and ecological recovery and with less 

variance. 

10 
Food Web Studies to Identify Biological 

Indicators. In an effort to begin to integrate a number 

of biological indicators and measures and explore 

alternative indicators of ecological recovery, food web 

studies were also conducted at TMC and Lake Hartwell. 

Several studies were conducted to characterize the 

aquatic and riparian food webs associated with the 

TMC-impacted area. These studies combined a number 

of novel techniques used in contaminated small stream 

systems, including: stable isotope analyses to predict 

trophic position and contaminant levels (Walters, Fritz et 

al., 2008), congener and chiral PCB chemistry to indicate 

uptake and biological processing (Dang, Walters et al., 

2010), and riparian export of PCBs to terrestrial predators 

(Walters, Fritz et al., 2008; Walters, Mills et al., 2009). 

Food web studies in TMC and Lake Hartwell revealed 

that t-PCB body burden was largely explained by an 

organism’s trophic position (i.e., the relative position 

in the food web as determined by stable isotopes of 

nitrogen, δ15N) and lipid content (Figure 7A, (Walters, 

Fritz et al., 2008; Walters-in review)). Ongoing high 

levels of PCBs were found in food web organisms 

throughout TMC suggesting that TMC could be 

an ongoing source of PCBs to the Lake Hartwell 

food web. This hypothesis was supported through 

ecosystem modeling for Lake Hartwell demonstrating 

that contaminated detritus loaded to the system was 

incorporated in the food web, thereby maintaining high 

levels of PCBs in fish even though concentrations in 

lake sediments were declining (Rashleigh, Barber et al., 

2009). 

Congener and chiral analyses were used in concert to 

identify patterns in transformation within TMC and Lake 

Hartwell. Biological and physical factors both strongly 

influenced congener composition. Higher-chlorinated 

congeners (HCCs) tend to be more hydrophobic and 

recalcitrant than lower-chlorinated congeners, and their 

proportional contribution to t-PCBs increased with 

trophic position in the food web (Walters, Fritz et al., 

2008; Walters - in review). Likewise, the proportion 

of HCCs increased with downstream distance from the 

Sangamo - Weston site, indicating that lighter, lesschlorinated compounds were lost to the system via 

volatilization (Walters, Fritz et al., 2008). In contrast, 

patterns in chirality in TMC were unrelated to distance 

from the Sangamo-Weston plant. Rather, biological 

factors were important regulators in these types of 

transformations. Chirality of PCB signatures varied 

among organic matter types such as fine benthic organic 

matter (FBOM), coarse particulate organic matter 

(CPOM), and benthic algae, suggesting that distinctive 

microbial communities among these organic matter types 

have unique biotransformation processes (Dang, Walters 

et al., 2010). 

Figure 7. A) δ15N Predicts ΣPCBs in the Stream Food 

Web; B) δ13C Predicts ΣPCBs in the Riparian Food 

Web (open symbols are outliers excluded from model); 

C) Sediment PCB Concentrations Strongly Predict 

Spider PCB Concentrations Along a PCB Gradient in 

Lake Hartwell. 

11 
 

Riparian studies of aquatic-to-terrestrial flux of PCBs 

focused on three key questions: 1) Which species in 

diverse riparian predator communities are at highest risk 

of exposure?, 2) Are concentrations in predators linked 

to concentrations in sediments?, and 3) How far inland 

do these fluxes extend? PCB concentrations in riparian 

predators along TMC were largely driven by their 

dependence on adult aquatic insects (Fig. 7B) (Walters, 

Fritz et al., 2008). That is, predators that consume a 

high proportion of aquatic insects had much higher PCB 

concentrations than predators relying on uncontaminated, 

terrestrial prey. Ambient PCB sediment concentrations 

were highly correlated with riparian spider concentrations 

(Fig. 7C), suggesting that riparian spiders are effective 

indicators of ecosystem recovery as well as indicators 

of potential risks to other terrestrial species such as 

arachnivorous birds (Walters, Blocksom et al., 2010). 

Concentrations in spiders along the shoreline were high, 

in some cases >10-fold higher than calculated wildlife 

risk values for arachnivorous birds. However, the spatial 

extent of this contamination and risk was small and 

limited to less than 5 m of the lake edge. Aquatic insect 

prey were limited to this narrow band of terrestrial habitat 

close to the lake margin, and spiders beyond this zone 

demonstrated consistent and low levels of PCB exposure 

(Walters, Mills et al., 2009). 

Physical Lines of Evidence 

The accumulation of progressively cleaner sediment 

was documented early in the research conducted at Lake 

Hartwell (Brenner, Magar et al., 2004). Using age dating 

coring techniques in conjunction with PCB chemistry, 

researchers demonstrated that accumulation of cleaner 

sediment over time reduces surface sediment contaminant 

concentrations. The age dating approach was used to

section through which intact sediment cores are inserted. 

The core has a rectangular cross-section (10 cm x 15 cm) 

and is less than 1 m in length. 

The measures of erosion rates for sediments as a function 

of shear stress and depth below the sediment surface are 

briefly described here. The undisturbed sediment core 

was inserted into the flume until the sediment surface 

was even with the bottom of the SedFlume channel. The 

flume was run at a specific flow rate corresponding to a 

particular shear stress (McNeil, 1996). Erosion rates are 

estimated by measuring the core length at different time 

intervals as shown in Table 3. 

The critical shear stress of a sediment bed, τcr, is defined 

as the shear stress at which a very small, but accurately 

measurable, rate of erosion occurs. For SedFlume 

studies, this rate of erosion has been practically defined 

as 10-4 cm/s. This rate represents 1 mm of erosion in 

approximately 15 minutes. Since it is difficult to measure 

τcr exactly at 10-4 cm/s, erosion rates were determined 

above and below 10-4 cm/s. The τcr was then determined 

by linear interpolation. The technique produces a τcr 

measurement with at least 20% accuracy (McNeil, 1996; 

Roberts, Jepsen et al., 1998). 

In addition to erosion rate measurements, samples were 

collected to determine the water content, bulk density, 

and particle size distribution of the sediments. Bulk 

density was determined by water content analysis using 

methods outlined in (Hakanson, 2002). The water 

content, W, was determined (Table 3) and then used to 

calculate bulk density, ρb, (Table 3). Lastly, particle 

size distributions were determined using laser diffraction 

analysis. The method was valid for particle sizes between 

0.04 and 2000 mm. Fractions over 2000 mm were 

weighed to determine the weight percentage greater than 

2000 mm. 

predict the time necessary for surface sediment 

concentrations to recover to a remediation target 

concentration. This approach and how this was 

further developed at Lake Hartwell (as well as other 

contaminated sediment sites) has been previously 

described (USEPA, 2008). 

Sediment Stability. The stability of the deposited 

sediment is a key component for successful 

MNR remedies. It is critical that clean sediment 

accumulation over the contaminated sediment is 

stable and not subject to scour, thereby preventing 

exposure of the contaminated sediment in the future. 

One technique for measuring sediment stability 

consists of simulating erosion events using any 

number of commercially available sediment flumes 

devices. A diagram of a SedFlume used to evaluate 

Lake Hartwell sediment stability is shown in Figure Figure 8. Schematic of the SedFlume Devise Used to 

8. SedFlume is a straight flume with an open bottom test Measure Sediment Stability. 

12 
Table 3. Sediment Stability Parameters Measured and Computed for Lake Hartwell. 

Figure 9. Picture of Core TN-SSE-01 Aligned with SedFlume Erosion Rate Data. 

13 
 

Figure 10. Erosion Rates (cm/s) as a Function of Depth (cm) for Three Cores on the O Transect at a Shear Stress 

of 1.6 N/m2. 

As an example, Figure 9 depicts the results for the 

SedFlume erosion analysis of a core, TN-SSE-01. The 

sediment found in this core consisted of fine silt over 

a sand/silt mixture over a stiff, silty sand layer. Small 

worms were visible near the surface of the core, and no 

other organisms were observed in the sediments. Core 

TN-SSE-01 contained an easily eroded sandy layer 

from 6 to 18 cm depth. The bulk density measurements 

increased markedly here with a corresponding increase in 

particle size. Below this sand layer was a silty sand layer 

that became progressively more resistant to erosion (stiff) 

with depth. This trend was evident by the decrease in 

bulk density, particle size, and erosion rates at the deepest 

depths of the core. 

To show the variable nature of sediments with respect to 

depositional characteristics and related stability within a 

small area (< 3-m radius), the average erosion rates (at 

the 1.6 N/m2 shear stress condition) for three replicate 

cores collected at Transect O are discussed below (Figure 

10). The erosion rates demonstrate the same trend for 

all three replicate cores with a pronounced stiff (more 

erosion resistant) layer in the 10 to 15 cm interval. The 

magnitude of shear stress in the shallower and deeper 

regions of the cores show some variation, but this is 

likely due to heterogeneity in the sediments supported by 

corresponding particle size and bulk density differences 

and rapid sediment losses (“blowouts”) in the core caused 

by gas pockets. Measurement of sediment stability is 

important to understand the permanence of the natural 

isolation that occurs during the accumulation of cleaner 

sediment over historically contaminated sediment 

surfaces. 

Porewater Exchange through Groundwater/ 

Surface Water Interactions. The movement of water 

(sediment porewater) into and out of sediment can be 

controlled by the flux of water from the groundwater 

to the surface water. This groundwater/surface water 

interaction and various means of measuring porewater 

PCB concentrations were also investigated at Lake 

Hartwell. This displacement of the sediment porewater 

to the overlying water column and the associated 

receptors was considered a possible pathway to 

provide continued exposure to the aquatic system. To 

evaluate this mechanism, methods to characterize the 

groundwater/surface water interaction and porewater 

measurement methods were tested and analyzed. The 

14 
Figure 11: Piezometer and In Situ Sediment Deployments for Phase 4 and Phase 5 

approach presented here uses a weight-of-evidence 

approach to characterize the sediment porewater to 

quantify the degree of mixing between groundwater and 

surface water, and with differential depth sampling, to 

quantify the magnitude and direction of the porewater 

movement. The approach relies on the collection of 

porewater, groundwater, and surface water samples and 

stable isotope data (oxygen and hydrogen) to define the 

proportion of groundwater and surface water in sediment 

porewater. Where available, traditional data (hydraulic 

head differential, conductivity, and temperature) were 

used as additional lines of evidence to corroborate the 

findings. 

Three transects were monitored using verticallystratified, clustered piezometer wells for quantifying 

the groundwater-surface water interaction. Two 

contaminated transects (T-O and T-N) and a background 

transect (BKG) were monitored continuously for 18 

months. At each transect, a series of well clusters was 

installed to evaluate the vertical hydraulic gradient. 

Figure 11 shows the well clusters relative positions 

compared to the shoreline and transects. One well was 

screened in the water column to measure surface water 

elevations. A second well was screened just below the 

sediment/water interface (shallow). A third well was 

screened at mid-depth (mid-level), and the last well was 

screened at the deepest point (deep). Each well was 

instrumented with a pressure transducer/temperature/ 

conductivity datasonde (600LS, YSI, Yellowsprings, 

Ohio). Additionally, the wells were sampled to 

characterize the porewater for PCBs and stable isotopes. 

Three additional porewater methods were co-located with 

these well clusters to allow comparison between the four 

porewater measurement approaches. 

The wells were used to measure head potential (the 

potential for water to move from one point to another 

based on pressure or head differentials) at different 

sediment depths. The water level in the wells at each 

depth within the cluster was monitored to calculate the 

head difference between the three depth increments. 

The surface water level was an important parameter in 

determining the direction (gaining from the lake or losing 

from the aquifer) and magnitude of head potential. This 

reservoir along with the entire Savannah River basin 

suffered drought conditions during these studies, with 

surface water elevations varying as much as 25 ft. This 

monitoring method resulted in observed head potential 

trends that varied as much as ±2.0 ft. Typically, the 

head potential between the deepest well and the surface 

water was less than 12 in., indicating the local aquifer 

and surface water were connected and did not take 

long to equilibrate. Figure 12 illustrates an example of 

differential heads for Lake Hartwell sediment over time. 

The differential head measurements indicated that in the 

shallow sediment depths (10-25 cm) the sediment was 

hydraulically connected directly to the surface water. 

This observation indicates the sediment is unconsolidated 

and free to exchange with the water column. In the 

mid-depth wells (80-130 cm), the degree of connectivity 

depended on the sediment characteristics at the transect. 

15 
Figure 12. Head Potentials Shown for Vertically Stratified Wells in Lake Hartwell Sediment. 

At T-O, the mid-depth was still connected with a slight 

dampening effect. At T-N, less connectivity was noted 

due to a more consolidated sediment. The deepest 

screened wells (180- 250 cm) indicated some connection 

to the water column, but the deepest screened wells did 

not exhibit day-to-day variations measured in the water 

column, but did exhibit the general declining trend of the 

surface water. 

Sediment Porewater Exchange Using Stable 

Isotopes. Another approach to support the exchange 

of sediment porewater with the overlying surface waters 

was stable isotope measurements of porewater for 

oxygen and hydrogen. Variations in the stable isotope 

ratios in natural waters are widely studied to interpret 

hydrological processes. Isotopes of hydrogen and 

oxygen, being inherent in the water molecule, can be 

effective conservative tracers in understanding movement 

and mixing within the hydrologic compartments. An 

abundance of literature exists on the application of stable 

isotopes in hydrologic investigations (Clark I, 1997; 

Kendall, 1998; Machavaram, Whittemore et al., 2006) 

and is only briefly described here. 

The ratios of most abundant stable isotope species 

of water molecules (HDO/HHO, HH18O/HH16O) 

in the water cycle are governed primarily by global 

precipitation/evaporation processes and are secondarily 

affected by local thermodynamic processes such as 

evaporation and mixing. Due to the differences in the 

vapor pressure and kinetic diffusivity of lighter and 

heavier isotopes, a preferential transfer of one isotope 

over the other occurs during phase changes. During 

evaporation, the lighter isotopes of H and 16O are 

preferentially transferred, thus causing the vapor to be 

isotopically ‘lighter’ (depleted) and the residual liquid 

to be ‘heavier’ (enriched). The reverse is true during 

condensation where D and 18O are preferred. This 

‘isotope fractionation’, which causes the ratios to change 

between the liquid and vapor phases, is governed by 

temperature, humidity, and the extent of transfer between 

the phases. Fractionation, in turn, results in distinct 

isotopic ratios for various hydrologic compartments that 

can be used as conservative tracers of water to investigate 

the interactions between compartments. Figure 13 

illustrates the application of this approach to show the 

different isotope end-points for surface waters compared 

to aquifer-influenced porewater samples. Mixing models 

are applied to estimate the contribution of groundwater 

and surface waters to the sediment porewater (Mills 2010 

- in preparation). 

16 
Figure 13. Mean d-excess values for Each Zone Measured and Plotted Against the Oxygen Isotope Data (BG = 

Background and O = Trasect O). 

Porewater PCB Concentrations. Porewater in 

contaminated sediments is a concern from the standpoints 

of the direct exposure of sediment dwelling organisms 

and the release of contaminated porewater to the 

overlying water, resulting in subsequent exposure to 

both humans and wildlife. Research has shown that 

measuring porewater in contaminated sediments can 

be difficult and often is biased by the measurement 

technique (Ehlers, 2006; Tomaszewski and Luthy, 2008). 

One aspect of the research conducted at Lake Hartwell 

was to compare four methods of measuring porewater 

PCB concentrations. The first method was a traditional 

technique of elutriating the existing porewater from core 

samples collected at depth. The second, third, and fourth 

methods, respectively, consisted of direct sampling of 

the porewater using lysimeter-type samplers inserted 

at depth, lysimeter points with SPMDs deployed for 

28 days, and SPMDs deployed in the subsurface wells 

used for head measurements. Further description and 

comparison of these methods are being developed for 

publication. 

Advective Porewater Transport. Advective 

transport of the COC with porewater through 

contaminated sediments has the potential to be a 

significant long-term contaminant source to surface 

sediments and the overlying water column and, 

ultimately, the resulting human or ecological receptors. 

Once in the water column or surface sediments, the 

contaminants may enter the food chain or a direct 

pathway to human or ecosystem exposure. Continued 

research is needed to define the magnitude of advective 

transport mechanisms and determine under what 

conditions this mechanism may pose a threat to managing 

the risk of contaminated sediments. 

Conclusions 

Research conducted at Lake Hartwell and TMC was 

initiated to provide technical support to U.S. EPA Region 

4 and to develop mechanistic-based tools and approaches 

to evaluate MNR and other contaminated sediment 

remediation strategies. The direct technical support 

provided to the Region 4 Project Manager resulted in 

modifications to the Lake Hartwell and TMC long-term 

monitoring plan to demonstrate MNR progress. Those 

changes included modifications to established fish tissue 

17 
monitoring protocols (a balance of male and female fish), 

addition of congener specific analyses on a subset of 

samples, and incorporation of bivalve monitoring in Lake 

Hartwell (USEPA, 2004). 

In addition, research initiated at Lake Hartwell has been 

further expanded to support additional contaminated 

sediment sites undergoing remediation by other 

management strategies such as capping and dredging. 

By basing the tools and approaches on sound science, 

conceptual site models, and mechanistic principles, 

combined with technical support as needed, these 

approaches and measures can generate reliable and useful 

data to support a wide variety of site conditions, COCs, 

and remediation strategies. 

Point-of-Contact. Marc A. Mills, Ph.D., 

U.S. Environmental Protection Agency, Office of 

Research and Development, National Risk Management 

Research Laboratory, 26 W. Martin Luther King Drive, 

Cincinnati, Ohio 45268 

Tel. (513) 569-7322, mills.marc@epa.gov. 

Acknowledgements 

A project of this scope and scale requires a large team 

of researchers and support staff. The authors, Marc 

Mills, James Lazorchak, Joseph Schubauer-Berigan, 

and David Walters, thank the following people for their 

significant contributions to this research. From EPA: 

Richard Brenner, Paul dePercin, Michael Griffith, Ken 

Fritz, Brent Johnson, Eric Kleiner, Terry Lyons, Frank 

McCormick, David Raikow, Brenda Rashleigh, Paul 

Wernsing, and Craig Zeller. From Battelle Memorial 

Institute: James Abbot, Greg Durell, Jon Eastep, Eric 

Foote, Greg Headington, Jennifer Ickes, Carol Pevin, and 

Shane Walton. From ENVIRON: Victor Magar. From 

Clemson University: Ryan Otter and Anthony Sowers. 

18 
 

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