ENVIRONMENTAL

RESEARCH BRIEF

The Impact of Ground-Water/Surface-Water Interactions on 

Contaminant Transport with Application to an Arsenic 

Contaminated Site 

Robert Ford* 

Introduction 

It is recognized that physical and chemical interactions between adjacent ground water 

and surface water bodies are an important factor impacting water budget and nutrient/ 

contaminant transport within a watershed (Winter et al., 1998). This observation is 

also of importance for hazardous waste site cleanup within the United States, since 

about 75% of sites regulated under the Resource Conservation and Recovery Act 

(RCRA) and the Comprehensive Environmental Response, Compensation, and Liability 

Act (CERCLA or Superfund) are located within a half mile of a surface-water body 

(Tomassoni, 2000; Biksey and Gross, 2001). The boundary between adjacent groundwater and surface-water bodies is referred to as the ground-water/surface-water (GW/ 

SW) transition zone. The transition zone plays a critical role in governing contaminant 

exchange and transformation during water exchange between the two water bodies. 

The intervening transition zone between a stream and its adjacent aquifer has 

historically been referred to as the hyporheic zone (see Triska et al., 1993 for specific 

definition). The more general terminology, GW/SW transition zone, is used throughout 

this document to stress the importance that water and solute exchange is not limited 

only to streams. 

The purpose of this document is to provide a brief overview of the dynamics of chemical 

processes that govern contaminant transport and speciation during water exchange across 

the GW/SW transition zone and to present results from a field study examining the fate 

of arsenic during ground-water discharge into a shallow lake at a contaminated site. A 

conceptual model of the GW/SW transition zone is defi ned to serve as a starting point for 

prioritizing tasks carried out during site characterization to define contaminant mass flux 

across the GW/SW transition zone. This information is a critical component towards establishing site-specific risks and alternatives for remedial intervention to reduce or eliminate 

these risks. Developing a knowledge base for delineating the biogeochemical processes 

controlling subsurface transport of the contaminant is one component of the investigative 

effort to define human or ecological risk. The discussion that follows necessarily ignores 

specific risk receptors. Since risk is dependent on the degree of current and future contaminant exposure to the receptor, it is important to define contaminant mass distribution (aqueous, solid, gas) and the dynamics of mass re-distribution within the regulatory 

boundaries established for the site. However, receptor response to contaminant exposure 

may not vary proportionally to contaminant mass or media-specific concentration, so the 

overall risk characterization effort must be guided by knowledge of both contaminant and 

receptor(s) distribution within the investigative boundary. 

* U.S. EPA, Office of Research and Development, National Risk Management 

Research Laboratory, Ground Water and Ecosystems Restoration Division, Ada, OK. 

contact: 580-436-8872, ford.robert@epa.gov 
Assessment of the factors controlling contaminant transport 

and distribution between interacting ground-water and surface-water bodies will be guided by the conceptual model 

that delineates the relevant hydrologic and biogeochemical processes. The conceptual model is developed based 

on site-specific data and will be improved or revised with 

continued accumulation of data that document relevant 

processes active on site. The complexity of the conceptual 

model and the extent of required refinement or revision will 

be dictated by site heterogeneity and process variability in 

time and space. While similarities may exist between different sites with regard to hydrology, contaminant identity, 

and system biogeochemistry, the initial conceptual model 

should always be viewed as a prototype for each given site. 

Therefore, site-specific monitoring plans should include 

iterative review and revision of the site conceptual model 

concurrent with improvements in the knowledge base. 

Evaluation of GW/SW Interactions at Waste Sites 

Regulatory Framework 

There is general recognition within the USEPA regulatory 

program of the significance of water transport across the 

GW/SW transition zone as a pathway for contaminant 

transport (Tomassoni, 2000). Programs administered in 

support of achieving requirements set forth in RCRA and 

CERCLA have common goals of returning usable ground 

water to beneficial uses where practical and preventing 

contaminant migration and exposure to surface-water 

receptors. The Environmental Indicators (EI) assessment 

programs administered through the RCRA Corrective Action (CA) Program and the Superfund Program provide an 

example where the Agency provides guidance for evaluating 

the importance of ground water/surface water (GW/SW) 

interactions for contaminant transport and exposure. These 

indicators are used to monitor intermediate progress in environmental terms at sites identified for restoration. Under 

Superfund, the Migration of Contaminated Groundwater 

Under Control (GM) Environmental Indicator is used for 

assessment of contaminant levels in ground water and 

the potential for contaminant migration and discharge to 

surface water (http://www.epa.gov/superfund/accomp/ei/ 

gwsurvey.pdf). Under the RCRA CA Program, the Migration 

of Contaminated Groundwater Under Control EI (RCRIS 

code CA750) determination is used to monitor whether 

any existing plumes of contaminated ground water are 

getting larger or adversely affecting surface-water bodies 

(http://www.epa.gov/epaoswer/hazwaste/ca/eis/ei_guida. 

pdf). Guidance documents developed to support these 

EI assessments point towards the potential need to collect cross-media data (e.g., ground water, surface water, 

sediment and ecological data) for evaluating site-specific 

progress towards meeting cleanup goals. 

Agency Documentation 

There are two documents published by the USEPA that 

directly address scientific and policy aspects of chemical 

transport between ground-water and surface-water bodies 

as well as approaches to site characterization. A review 

of methods for assessing contaminated ground-water 

discharge to surface water was published by the Office of 

Water in 1991 (EPA, 1991). While there have been new 

technological developments that improve the capability for 

delineating GW/SW interactions, the general approaches 

and many of the specific techniques for mapping the discharge of non-point contaminant sources from ground-water 

to surface-water bodies are still applicable. The Office of 

Water has supplemented the 1991 document by outlining an 

approach for using biological indicators to assist in assessing the impact of contaminated ground-water discharge on 

surface-water quality (EPA, 1998a). The U. S. Geological 

Survey has also published examples where contaminant 

discharge area(s) have been mapped out using hydrologic 

and physicochemical techniques to assess the fate of 

volatile organic compounds and inorganic contaminants 

during transport across the GW/SW transition zone (e.g., 

Savoie et al., 2000; McCobb et al., 2003). 

The USEPA Office of Solid Waste and Emergency Response (OSWER) sponsored a workshop in January 

1999 to bring together a group of experts to address the 

ecological importance of the GW/SW transition zone and 

review approaches to characterize relevant hydrogeological, chemical, and biological processes within this zone. 

This workshop was jointly organized by the Ecological Risk 

Assessment Forum (ERAF) and the Ground Water Forum 

(GWF), which comprise ecological risk assessment and 

ground-water specialists, respectively, from USEPA Regional Offices, Headquarters, and the Office of Research and 

Development. A proceedings document was published to 

synthesize workshop discussions and to propose directions 

for future research that will benefit the USEPA regulatory 

mission (EPA, 2000a). This document also provides an 

update to many of the methods for site characterization 

published by the Office of Water and a compendium of 

case studies where GW/SW investigations have been a 

critical component of site characterization. Together, these 

publications provide a concise synthesis of the technical 

and policy issues relevant to contaminant transport across 

the GW/SW transition zone, and they should be consulted 

as resources for designing a site characterization plan. In 

addition, a draft ERAF-GWF Issue Paper is currently in 

preparation entitled, “Evaluating Ground-Water/SurfaceWater Transition Zones in Ecological Risk Assessment.” 

The intent of this document is to provide Agency guidance 

on data requirements and techniques for performing an 

ecological risk assessment. 

National Research Initiatives 

As a regulatory body, the USEPA does not maintain/operate field sites to support long-term research in the area of 

GW/SW interactions. However, several site-specific case 

studies are documented within the 1999 OSWER-sponsored 

workshop on GW/SW interactions (EPA, 2000a) that can be 

consulted as points of reference for designing a monitoring plan to support a site-specific conceptual model. The 

Agency also maintains web-based resources pertinent 

to this technical area, e.g., the Ground Water Task Force 

website located at http://gwtf.cluin.org/resources/. This 

website provides a compilation of information sources for 

characterization of ground-water plumes, several of which 

2
pertain directly to GW/SW interactions. In addition, the 

USEPA partners with other federal agencies, such as the 

U. S. Geological Survey (USGS), that are better positioned 

to support long-term field-based research initiatives. The 

USGS supports several research programs that address 

chemical transport at the watershed level, including hydrogeologicalinvestigations pertaining to the GW/SW transition 

zone. The National Science Foundation also sponsors the 

Long Term Ecological Research (LTER) Network, which 

has as one of its core research areas the assessment of 

patterns of inorganic inputs and movements of nutrients 

through soils, ground water, and surface waters. Internet 

links are provided below to several of these programs: 

1) Water, Energy, and Biogeochemical Budgets Program 

http://water.usgs.gov/webb/ 

2) Toxic Substances Hydrology Program 

http://toxics.usgs.gov/ 

3) Shingobee Headwaters Aquatic Ecosystems Project 

http://wwwbrr.cr.usgs.gov/projects/SHAEP/index.html 

4) Long-Term Ecological Research Network 

http://lternet.edu/ 

While the primary focus of these research initiatives is not 

exclusive to contaminated sites, they provide a unique 

resource to practitioners due to the knowledge base 

developed relative to evaluation of successful tools for site 

characterization and modeling of water transport and solute 

interactions across the GW/SW transition zone. 

Conceptual Model Development 

Interacting System Components

A simplified diagram of the GW/SW transition zone is 

shown in Figure 1A. The graphic simplicity of this diagram 

is purposeful to emphasize the importance of process identification (hydrologic or biotic-abiotic chemical processes) 

in guiding design of the monitoring program for assessing 

contaminant transport and distribution. Three components 

areidentified within a hypotheticalboundary encompassing 

the ground-water and surface-water bodies that exchange 

water and the chemicalcomponents mobile within this phase. 

The three system components include 1) the ground-water 

aquifer, 2) the GW/SW transition zone, and 3) the surfacewater body. Arrows located at the water surface in both 

the ground-water aquifer and the surface-water body are 

shown to emphasize that water levels may fluctuate in 

Figure 1. Conceptual model of the system components that govern contaminant transport across the GW/SW transition 

zone. (A) Chemical fluxes entering the GW/SW transition zone (Zone 2) emanate from surface-water (Zone 

3) and/or ground-water (Zone 1) sources. Arrows shown at the surface of the ground- and surface-water 

components indicate the fluctuations in hydraulic head that may occur resulting in a potential change in the 

direction of water flow. (B) The soil/sediment material within the GW/SW transition zone can further be delineated into surface sediments (or gyttja) and aquifer solids. (C) Solid matter within the GW/SW transition 

zone includes inorganic/organic substrates of which the soils/sediments are composed as well as plant and 

animal biomass. 

3
either component. Differences in the relative water level 

between these two components will dictate both the direction 

and magnitude of water flux across the GW/SW transition 

zone (Winter et al., 1998). The detail shown in Figure 1B 

is included to emphasize that the physical composition of 

the GW/SW transition zone may incorporate both aquifer 

and surface-water sediments. The importance of plants 

as well as the microbial and benthic communities that inhabit the GW/SW transition zone is emphasized in Figure 

1C. These organisms play an important role in governing 

nutrient transport and the development of chemical gradients within the GW/SW transition zone (e.g., Jaynes and 

Carpenter, 1986). Chemical gradients develop across the 

GW/SW transition zone due to the inherent difference in the 

ground-water and surface-water chemistry and the biotic 

communities inhabiting this unique ecological niche. The 

microbial community within the GW/SW transition zone 

may, in part, govern contaminant transport across this 

zone through regulation of the chemistry of redox-sensitive 

elements such as carbon, iron, nitrogen, and sulfur. This 

will impact the chemical composition of water and solids 

within this zone and, therefore, contaminant mobility (EPA, 

2000b). 

Surface-water sediments are delineated separately from 

aquifer sediments in the diagram shown in Figure 1 due 

to their intimate contact with the surface-water body. As 

a result of inputs from terrestrial sources, surface-water 

sediments (referred to as gyttja) may develop chemical 

properties distinct from adjacent aquifer sediments. Due to 

these unique characteristics, contaminants that otherwise 

are not attenuated by sorption onto aquifer sediments may 

accumulate within surface-water sediments. While contaminant partitioning to surface-water sediments may mitigate 

transport into the surface-water body, sediments that have 

accumulated a significant mass of a recalcitrant contaminant may present either a separate exposure medium or 

pose a significant long-term source of contamination to the 

surface- or ground-water body as a result of contaminant 

release. In addition, contaminant reactions with surfacewater sediments may be controlled by light- or temperature-dependent factors, e.g., photosynthesis-induced pH 

changes (e.g., Jones et al., 2004 and references therein). 

These reactions are distinct from those that may be active 

outside of the zone of light influence in buried surface-water 

sediments or aquifer sediments within the GW/SW transition zone and must be accounted for separately in order to 

properly assess the influence of ground-water discharge 

on contaminant mass in the surface-water body. From 

this perspective, it is important that the site investigation 

provides data to assess contaminant mass loading and 

reactivity in surface-water sediments, including sorption 

capacity for continued contaminant uptake. 

Contaminant flux may be multi-directional and transient, 

so developing an understanding of contaminant flux within 

or between the relevant compartments will depend on 

ascertaining the dynamics and distribution of temporal and 

spatial factors governing contaminant reactive transport. 

Contaminant transport and, therefore, exposure-based risk 

estimates can best be understood in the context of a mass 

balance of the contaminant(s) of concern. A contaminant 

may remain mobile across the GW/SW transition zone, 

become immobilized on sediments within or adjacent 

to the boundaries of the transition zone, or undergo 

transformation in water to products with lower/greater toxicity. 

Development of an understanding of the processes that 

control contaminant mass distribution within the boundaries 

of the site investigation underpins the risk evaluation and 

the design of intervention strategies to minimize risk. 

Unifying Concept: Contaminant Flux 

While contaminant concentration is a determining factor 

for human or ecological risk, this metric does not provide 

a measure of contaminant mass and distribution within the 

system of interest. Determination of a contaminant mass 

balance is critical for determining changes in contaminant 

mass or speciation during water transport within and across 

the transition zone. The transition zone not only encompasses the transition from one flow regime to another, it is 

also commonly the location of dramatic chemical gradients 

driven by biotic/abiotic reactions. For organic contaminants, a mass balance calculation aids determination of 

whether contaminant degradation or sorption is occurring. 

For inorganic contaminants, a mass balance is required 

to assess changes in chemical speciation and mobility. 

The mass balance is inclusive of liquid and solid matrices 

(e.g., sediments) within the boundaries of the conceptual 

model. The distribution of contaminant mass is important 

with respect to projecting current and future risk. This is 

critical for recalcitrant organic compounds and inorganic 

contaminants that may accumulate on solid matrices within 

the transition zone, thus posing future risk due to processes 

such as contaminant desorption. 

Contaminant flux (M) is defined as the product of contaminant concentration (C) in the mobile phase (water and mobile 

colloids) and the volumetric flow of the mobile phase (Q). 

M = C * Q Eq. 1 

Contaminant flux can be calculated for point locations or 

cross-sectional areas perpendicular to water fl ux depending 

on the level of heterogeneity in water flow or contaminant 

concentration distribution (Einarson and Mackay, 2001; 

Buscheck et al., 2004). For inorganic contaminants, this 

general equation represents the flux of all contaminant 

species at a given point in time. Since a mobile inorganic 

contaminant can change chemical form, it may be useful or 

necessary to further define contaminant flux for individual 

contaminant species relevant to site-specific conditions.

 M = C * Q Eq. 2 i i

Contaminant transport occurs along water flow paths in 

the subsurface. Therefore, the first step to determining 

contaminant flux is developing an understanding of system 

hydrology. With the establishment of a water budget for 

the site (i.e., water flux distribution), then a mass budget 

can be developed to establish contaminant flux across the 

system component boundaries. 

Within the aquifer and the GW/SW transition zone, contaminant flux is dependent on water flow and contaminant 

4
Figure 2. Conceptual diagram illustrating potential spatial heterogeneity and variability in the magnitude of contaminant 

flux across the GW/SW transition zone: A) plan view and B) cross-sectional view through monitoring transect 

D-D’. Variation in magnitude of contaminant flux is indicated by thickness of arrows and is a function of the 

contaminant concentration and water flux within a given reaction volume. 

distribution as illustrated in Figure 2. The spatial and 

temporal distribution of the contaminant will depend on 

source location, the spatial distribution and velocity of 

water flow (or diffusion where advective flux is low), and 

the abundance and biotic/abiotic reactivity of aqueous 

and solid phase biogeochemical components along the 

paths of water flow. The extent to which biogeochemical 

processes will influence contaminant chemical speciation 

during transport will depend on the relative rates of fluid 

transport and chemical reactions (Morgan and Stone, 

1985). For example, contaminant transport will be relatively 

conservative for systems in which the timescale for water 

flow within a hypothetical reaction volume is much shorter 

than the timeframe for significant reaction to take place. 

Conversely, significant attenuation/transformation will be 

observed for contaminants in systems where reactions occur 

rapidly relative to the timescale for fluid transport. The latter 

situation is commonly assumed to apply to ground-water 

systems, but this needs to be confirmed during site characterization. This is particularly important for near-surface 

systems that may experience large variability in water fl ux 

and fluid velocities (e.g., Conant Jr., 2004). 

Sebestyen and Schneider (2004) present field results 

that illustrate the interdependence between fluid flow 

and chemical reactions. Trends in the sediment pore 

water concentrations of iron and zinc were related to the 

magnitude of ground-water seepage. Low seepage flows 

resulted in higher concentrations of these metals due to 

the development of anoxia from microbial reactions. The 

influence of these reactions was diminished with higher 

seepage flows, since solutes were rapidly flushed from 

the system. Harvey and Fuller (1998) have proposed a 

dimensionless index that may be used to assess the relative significance of chemical reactions within the hyporheic 

zone in mitigating solute transport within a drainage basin 

based on the assessment of chemical reaction rates and 

fluid residence times within the GW/SW transition zone 

(specifically, the hyporheic zone). These authors provide a 

detailed assessment of the influence of manganese uptake 

within the hyporheic zone in the Pinal Creek Basin, Arizona. 

Manganese was partially sequestered from contaminated 

ground water within the hyporheic zone due to the downward flux of oxygen from stream water, which stimulated 

microbial oxidation and precipitation of manganese within 

shallow sediments. As demonstrated in a subsequent 

publication (Fuller and Harvey, 2000), the manganese 

oxidation-precipitation process exerted direct influence on 

the transport of contaminant metals (cobalt, nickel, zinc) 

derived from ground-water discharge to the stream. These 

5
studies illustrate the importance of understanding the balance between fluid transport and chemical reaction rates 

for assessing contaminant migration across the GW/SW 

transition zone. 

A concise summary of relevant chemical processes that 

may influence contaminant flux is presented by Winter et 

al. (1998). Common types of biogeochemical reactions 

that impact contaminant transport across the GW/SW 

transition zone include acid-base reactions, precipitation 

and dissolution of minerals, sorption and ion exchange, 

oxidation-reduction reactions, biodegradation, and dissolution and exsolution of gases (Winter et al., 1998). 

Examples of these reaction types relevant to the GW/SW 

transition zone are presented in Table 1. These reactions 

will impact contaminant phase distribution (liquid, solid, gas) 

and chemical speciation within a given phase. Thus, it is 

critical to understand the chemical properties unique to the 

contaminant(s) under consideration in relation to the biogeochemical characteristics of the GW/SW transition zone. 

A detailed discussion of the physical and biogeochemical 

processes leading to organic contaminant (chlorinated 

solvents and fuel hydrocarbons) degradation/immobilization is provided by Wiedemeier and others (EPA, 1998b). 

Discussions of chemical processes that may influence 

inorganic contaminant transport in surface and subsurface 

systems are presented in texts by Chapelle (1993), Stumm 

and Morgan (1996), and Langmuir (1997). The reader is 

referred to these references for a more comprehensive 

overview of chemical processes relevant to lacustrine 

sediment and ground-water systems. 

Documented Impact of GW/SW Interaction on 

Chemical Transport 

Understanding of the factors controlling contaminant transport across the GW/SW transition zone is improving with 

increased publication of site-specific field studies in the 

literature. The literature review that follows highlights salient 

features that should be considered during assessment of 

contaminant transport across the GW/SW transition zone. 

This review targets only those studies where contaminant 

fate was examined in conjunction with assessment of 

system hydrology. There are numerous studies that focus 

solely on evaluating and modeling the hydrologic characteristics of GW/SW interactions for various flow regimes, 

and the reader should consider alternate sources for additional background on this aspect of site characterization. 

Literature reports included in Tables 2 and 3 below were 

predicated on thorough characterization of site-specific 

hydraulic characteristics prior to initiating or interpreting 

results from geochemical site characterization. 

Table 1. Classes of Biochemical Reactions with Examples Relevant to Contaminant Transport Across the GW/SW 

Transition Zone. 

Geochemical 

Reaction Relevant Process Example Reaction 

Acid neutralization 

Acid-Base by aqueous HCO3 

-

 + H+ = H2CO3 

o 

carbonate alkalinity 

PrecipitationDissolution of 

Minerals 

Precipitation of 

metal sulfide Zn2+ + HS-

 = ZnS(s) + H+ 

Sorption and Ion 

Exchange 

Ion exchange on 

feldspars KAlSi3O8(s) + NH4 

+

 = NH4AlSi3O8(s) + K+ 

Oxidation-Reduction 

Reductive 

dissolution of iron 

oxide coupled to 

organic carbon 

oxidation 

4Fe(OH)3(s) + 8H+ + CH2O = 4Fe2+ + CO2(g) + 

11H2O 

Biodegradation 

Benzene oxidation 

coupled to 

denitrification 

C6H6 + 6NO3 

-

 + 6H+ = 6CO2(g) + 6H2O + 3N2(g) 

Gas Dissolution and 

Exsolution 

Ammonia gaswater exchange NH3(g) + H2O = NH4 

+

 + OH6
Table 2. Examples from Peer-Reviewed Literature Documenting Field Case Studies Designed to Assess the Physicochemical Processes Infl uencing Inorganic Contaminant Transport Across the GW/SW Transition Zone. OC = organic carbon, DOC = dissolved organic carbon, EDTA = ethylene daminetetraacetic acid. 

CompoundIdentity Location Hydrologic Setting Chemical Processes Reference Nitrate Great South Bay, New York Submarine ground-water discharge to coastal marine ecosystem Denitrification (N 2 or N O production) and dissimilation 2 (NH4 -N production) in sediments Capone and Bautista, 1985 

Nitrate, Ammonia 

Seine River, France 

River-water infiltration into ground 

water

Nitrification-denitrification coupled to microbial 

mineralization of aqueous and solid OC; ammonia release from OC decay 

Doussan et al., 1997 

Nitrate Elbe River, eastern Germany water River-water infiltration into ground Denitrification using oxidizable aqueous and solid OC from river water and river bed/aquifer sediments Grischek et al., 1998 

Nitrate 

Western Sierra Nevada foothills, California 

Ground-water discha

temporal variations in mixing of bedrock and soil drainage sources 

rge into stream; Nitrification of ammonia; leaching of nitrogen 

compounds from soil solids 

Holloway and Dahlgren, 2001 

Nitrate Northwestern France discharge into second-order stream Shallow and deep ground-water Denitrification coupled to pyrite oxidation Grimaldi et al., 2004 

Nitrate Florida Keys Ground-water di canal system scharge to coastal Near-conservative transport of low-salinity plume; denitrification in off-axis portions of plume Griggs et al., 2003 Phosphate Tidal freshwater marsh, Virginia Marsh ground-water infiltration- exfiltration Iron oxidation-precipitation limits diffusive and advective phosphate transport into marsh Chambers and Odum, 1990 

Manganese, Nitrate 

Oder River, northeastern Germany River-water infiltration into ground water Denitrification and Mn reduction-oxidation coupled to microbial degradation of OC Massmann et al., 2004 

Manganese Lot River, Bordeaux, France River-water infiltration into ground water Weathering of Mn-bearing minerals coupled to bacterial degradation of OC; Mn(II) oxidation coupled to sorption-precipitation Bourg and Bertin, 1993 

Manganese Lot River, Bordeaux, France River-water infiltration into ground water

Weathering of Mn-bearing minerals coupled to microbial degradation of OC – spatial variability due to different OC or mineralogical Mn sources; Mn(II) oxidation coupled to sorption-precipitation 

Bourg and Bertin, 1994 

Cadmium, Copper, Iron, Manganese, Nitrate, Zinc 

Glatt River, northeastern Switzerland 

River-water infiltration into ground 

water

Mineral weathering/precipitation coupled to microbial 

degradation of OC; Cu-EDTA complexation; metal sorption-precipitation due to fluctuations in abiotic chemical equilibria; denitrification 

Jacobs et al., 

1988 

Cadmium, Copper, Manganese, Nitrate, Zinc 

Glatt River, northeastern Switzerland 

River-water infiltration into ground 

water

Mineral weathering/precipitation coupled to microbial degradation of OC; Cu-DOC complexation; metal sorption-precipitation due to fluctuations in abiotic chemical equilibria; denitrification 

von Gunten et al., 1991 

Cadmium, Copper, Iron, Manganese, Nitrate, Zinc 

Silver Bow Creek, Montana exfiltration 

River-water/ground-water infiltration- Oxidative precipitation of Fe and Mn; precipitationsorption of Cd, Cu, Zn; denitrification Benner et al., 1995 

Arsenic, Iron 

Silver Bow Creek, 

Montana 

River-water/ground-water infiltrationexfiltration Dissolution of As-bearing iron oxides from sediment deposited in reducing hyporheic zone Nagorski and Moore, 1999 

Iron, Zinc 

Several lakes, Adirondack Mountains, New York 

Lake/ground-water infiltrationexfiltrations Anaerobic microbial OC degradation, mineral weathering and ion exchange reactions Sebestyen and Schneider, 2004 

7 
Table 3. Examples from Peer-Reviewed Literature Documenting Field Case Studies Designed to Assess the PhysicoChemical Processes Infl uencing 

Organic Contaminant Transport Across the GW/SW Transition Zone. EDTA = ethylene diaminetetraacetic acid, NTA = nitrilotriacetic acid, NDC = naphthalene dicarboxylate, APEC = alkylphenol polyethoxy carboxylates, HAA = haloacetic acid, CFC = chlorofl uorocarbon. 

Compound Identity Location Hydrologic Setting Chemical Processes Reference Trichloroethylene, Tetrachloroethylene, 1,4-Dichlorobenzene, 1,3-Dimethylbenzene, Organochlorine 

Glatt and Aare Rivers, northeastern Switzerland 

River-water infiltration into 

ground water 

Conservative transport for some volatile organic compounds; aerobic and anaerobic microbial degradation; sorption of lipophilic compounds limited by low organic carbon content of aquifer solids 

Schwarzenbach et al., 1983 

Volatile, Semivolatile, Non-volatile compounds Croatia, Yugoslavia 

Sava River, northern 

ground water 

River-water infiltration into Inferred biological degradation and/or sorption onto solid OC in surface water and aquifer 

sediments 

Ahel, 1991 

Chloroform, Benzene, Tetrachloroethylene, Trichlorethylene 

Watson Creek, Aberdeen Proving Ground, Maryland 

Ground-water discharge 

into stream 

Not studied; passive samplers used to delineate contaminant plume discharge zone 

Vroblesky et al., 1991 

Volatile Organic Compounds 

Inner Piedmont, South Carolina into stream 

Ground-water discharge Not studied; passive samplers used to locate rock fractures that were conduits for 

contaminant discharge 

Vroblesky et al., 1996 

Alkylphenol polyethoxylate surfactants 

Glatt and Sitter Rivers, northern Switzerland 

River-water infiltration into 

ground water 

Biological degradation of organic compounds primarily within GW/SW transition zone (shallow aquifer) 

Ahel et al., 1996 

Chloroethene St. Joseph, Michigan 

Groun

recharge into lake with 

wave infiltration

d-water discharge- Chloroethene co-oxidation by methaneoxidizing bacteria Lendvay et al., 1998 

EDTA, NTA, NDC, APEC, HAA 

Santa Ana River, Orange County, California 

River-water infiltration into 

ground water Microbial degradation of NTA, APEC and HAA Ding et al., 1999 

Trichloroethylene, 1,1,2,2-tetrachloroethane 

West Branch Canal Creek, Aberdeen Proving Ground, Maryland 

recharge into wetland and 

Ground-water dischargestream Anaerobic biodegradation (hydrogenolysis and reductive dechlorination) under reducing conditions (iron-reducing, sulfate-reducing, methanogenesis) Lorah and Olsen, 1999 

Tetrachlorethylene Pine River, Angus Ontario, Canada 

Ground-water discharge 

into river (short circuiting and low-high dispersed flow)

Reductive dehalogenation producing cis-1,2-dichloroethylene and vinyl chloride (lowmoderate discharge zone) Conant Jr., 2000 

Tetrachloroethene Pine River, Angus Ontario, Canada into river (short circuiting and low-high dispersed 

Ground-water discharge 

flow)

Reductive dehalogenation producing cis-1,2-dichloroethylene and vinyl chloride (lowmoderate discharge zone) Conant Jr. et al., 2004 

CFC-11, CFC-12, CFC113 Everglades National Park, south Florida Surface-water infiltration into ground water CFC degradation coupled to microbial methanogenesis Happell et al., 2003 

αhexachlorocyclohexane Coastal northeastern Florida Ground-water discharge into a stream Biological degradation in non-acidic, anaerobic portions of plume Law et al., 2004 

8 
Summaries of published studies from the peer-reviewed 

literature that document the influence of GW/SW interactions 

on inorganic and organic contaminant fate are provided in 

Tables 2 and 3, respectively. Several general themes are 

echoed throughout the summarized studies: 

1) Hydrologic transients imposed by long-term (e.g., 

seasonal) and short-term (e.g., storm events) flow 

variations impact the types and intensity of chemical 

reactions that influence contaminant chemical speciation and transport; 

2) Chemical gradients of major and trace elements/contaminants are often greatest in the vicinity of the GW/SW 

transition zone; 

3) Contaminant transport is dependent on the diversity 

and interactions between the biotic and abiotic solid 

components within the GW/SW transition zone; and 

4) Microbial degradation of natural and anthropogenic 

sources of organic matter and the availability of terminal 

electron acceptors influence the distribution of redox 

zones within the GW/SW transition zone. 

Hydrologic Transients. With regard to hydrology, it must be 

understood that water flow across the transition zone is a 

two-way street. The direction of water flow (and contaminant 

flux) will respond to changes in water level in surface water 

or ground water. Changes in water level may originate from 

natural events such as rainfall, snow melt, evaporation, or 

tidal/wave influences or from man-made events such as 

changes in water storage in a managed reservoir or from 

ground-water withdrawal. In wetland or peatland settings 

with dense plant growth, water transport due to evapotranspiration can also exert a signifi cant influence on ground-water 

flow patterns (Doss, 1993; Fraser et al., 2001). Water flow 

dynamics influence contaminant transport directly due to 

changes in the volume and direction of water flux. However, 

changes in water flow also exert an indirect effect by altering 

major element chemistry that can stimulate or quench biogeochemical reactions controlling the chemical speciation 

of contaminants. As an example, changes in the activity 

of organic matter degradation that accompany variations 

in the flux of dissolved organic matter due to ground-water 

recharge/discharge may impact iron and/or sulfur oxidationreduction reactions. Oxidation-reduction transformations of 

iron or sulfur drive the precipitation/dissolution of iron- and 

sulfur-bearing minerals involved in inorganic contaminant 

sorption (e.g., Moore et al., 1988; Moore, 1994) or abiotic 

degradation of organic contaminants (e.g., Ferrey et al., 

2004). Thus, the accuracy of a site-specific conceptual 

model depends on defining the interaction between fl uid fl ow 

and the biogeochemical processes governing contaminant 

transformations and mobility. 

Chemical Gradients. The dynamics of hydrologic events in 

concert with the diversity and activity of microscopic and 

macroscopic organisms that occupy the GW/SW transition zone can result in steep gradients in chemical mass 

in transitioning from the ground-water to the surface-water 

body (and vice versa). The chemical gradients that develop 

may pertain both to the major element composition as 

well as the contaminant mass in the water or solids in the 

GW/SW transition zone. Two examples are discussed to 

illustrate the magnitude of chemical mass changes relative 

to transport distances. The first example is taken from work 

conducted by Conant et al. (2004) to assess the extent of 

PCE discharge from a contaminated ground-water plume 

into a stream. Determination of the potential for natural attenuation/degradation of PCE during transport was a critical 

component of this assessment. The authors observed the 

following: 1) transport of PCE with little attenuation to the 

streambed, 2) the presence of both PCE and degradation 

products within the streambed, and 3) low detections of 

PCE and degradation products in the overlying stream. The 

predominant fraction of anaerobic degradation of the PCE 

plume occurred within the top 2.5 meters of the streambed. 

Solids within the streambed also served to concentrate 

volatile organic compounds from the plume due to sorption or retardation. Failure to monitor water and sediment 

concentrations within the streambed could have resulted 

in a misinterpretation of low surface water PCE concentrations as resulting only from dilution of the discharging 

plume. The second example illustrating the development 

of sharp chemical gradients is taken from work published 

by Benner et al. (1995). These authors observed changes 

in water chemistry during GW/SW interactions in a stream 

receiving discharge from an acidic, metal rich ground water. 

Order-of-magnitude changes in cadmium, copper, and 

zinc concentrations were observed over a distance of less 

than one meter during ground-water transport through the 

hyporheic zone into the overlying stream. The reduction in 

contaminant metal concentrations was due to partitioning 

to solid phases such as iron oxides that precipitated as pH 

and dissolved oxygen levels increased during transport from 

ground water to surface water. Knowledge of the solidphase partitioning process was critical in order to assess 

the potential long-term stability of the contaminant metals 

accumulating within streambed sediments. 

Transition Zone Composition. Solid components within a 

ground-water aquifer are primarily derived from weathered 

rock or soil materials and subsurface microbial communities. In contrast, solid components within the GW/SW 

transition zone may include both aquifer materials as well 

as microscopic/macroscopic biotic and abiotic components 

that are primarily derived from the surface terrestrial system. 

For example, both submerged and emergent plants can 

contribute a significant portion of the total mass of solid 

components within the GW/SW transition zone. These 

components can play two critical roles with regard to the 

chemistry within this zone 1) by influencing the transport 

of important chemical constituents directly through normal physiological functions (e.g., Jaynes and Carpenter, 

1986) or indirectly by hosting microbial communities that 

infl uence localized oxidation-reduction (redox) conditions, 

and 2) by contributing a source of organic material to 

surface sediments within the GW/SW transition zone at 

the end of their life cycle. The abiotic organic and mineral 

components within the GW/SW transition zone, in part, 

control the system chemistry by buffering changes in redox 

or pH and serving as temporary/permanent reservoirs for 

9
contaminants that are absorbed/adsorbed during transport 

across the transition zone. 

Redox Chemistry. The importance of redox processes for 

organic andinorganic contaminant transport and fatein subsurface systems at waste sites has been addressedin detail 

in a recent Agency publication (EPA, 2000b). Developing 

an understanding of the relevant redox processes controlling contaminant degradation or solid-liquid partitioning is a 

critical task for projecting contaminant flux. Due to spatial 

and temporal heterogeneity in the physical and chemical 

properties across the GW/SW transition zone, multiple 

reaction zones with differential impacts on contaminant 

transport may develop as illustrated in Figure 3. Many 

redox reactions that govern contaminant fate are directly 

Figure 3. Conceptualdiagram showing snapshot of chemical component profiles across the GW/SW 

transition zone. Redox reaction zones develop 

within fixed spatialregions due to biogeochemical reactions maintained by chemical fluxes. 

Conceptual adaptation of illustration from 

Figure 7 in Bourg and Bertin (1993). NOM = 

natural organic matter, TEAP = terminal electron acceptor process. 

or indirectly influenced by microbial degradation of natural 

or anthropogenic organic compounds coupled to electron 

transfer reactions with terminal electron acceptors such 

as oxygen, nitrate, ferric iron, sulfate, and carbon dioxide 

(Chapelle, 1993; Vroblesky and Chapelle, 1994). 

Variations in the mass flux of these chemical components in 

addition to temperature fluctuations will influence both the 

type and intensity of microbially-mediated redox reactions 

that affect contaminant chemical speciation. For example, 

Groffman and Crossey (1999) monitored the response 

of oxygen, iron, manganese, and sulfur distributions in a 

shallow alluvial aquifer in contact with a first order stream 

influenced by rain and snow melt infiltration events. The 

distribution of terminal electron acceptors or redox zones 

in the shallow aquifer shifted spatially and temporally in 

response to seasonal infiltration events. These variations 

were significant across spatial and temporal scales of 

centimeters and weeks, respectively. The impact of flow 

fluctuations on nitrate concentrations in a stream partially 

fed by ground-water discharge was illustrated by Grimaldi 

et al. (2004). Reduction of nitrate coupled to the oxidation 

of sulfur in the pyrite-bearing aquifer played a key role in determining stream nitrate concentration throughout the year. 

The magnitude of nitrate reduction was mediated by daily, 

storm-event, and seasonal variations in system hydrology. 

Both studies point to the importance of delineating the role 

of both soluble and solid-phase redox reactants/products 

toward developing a sound conceptual understanding of 

contaminant fate across the GW/SW transition zone. 

These studiesillustrate theimpact of redox-driven processes 

on the transport of inorganic and organic contaminants 

across the GW/SW transition zone. The evaluations documentedin these reports provide a usefulcontext for evaluating the spatial and temporal detail needed in a monitoring 

program designed to support development and validation 

of site-specific conceptual and analytic models. 

Application to an Arsenic Contaminated Site 

Contaminant source identification is a critical task during 

site characterization to assess contaminant transport across 

the GW/SW transition zone. Identification of the major 

contaminant sources provides the basis for assessing potential long-term contaminant flux to the risk receptor and 

for targeting intervention efforts to minimize contaminant 

transport and exposure. As previously noted, a common 

direct source of contaminant flux to a surface-water body 

is from discharge of a contaminated ground-water plume 

from an upgradient waste site. Removal or isolation of the 

contaminant source zone within the upgradient ground-water 

aquifer is a logical step to minimize continued contaminant 

flux. However, historical contaminant attenuation and accumulation within the GW/SW transition zone may pose 

a secondary long-term source of contaminant flux to the 

surface-water body. Determination of the fraction of the 

total contaminant flux derived from sediment desorption/ 

dissolution thus becomes an important aspect for apportioning risk to the various contaminant source terms and 

for evaluating whether intervention is required to manage 

estimated risk from sediment contamination. 

10
The following field study provides an example where surface-water sediment contamination due to arsenic transport 

from an upgradient ground-water plume poses a potential 

long-term source for arsenic flux into the surface-water 

body. Methods employed for site assessment are illustrated 

along with field data that provide evidence that the flux of 

arsenic into surface water is due to 1) direct discharge from 

contaminated ground water and 2) concurrent release of 

arsenic from contaminated sediments. 

Initial Site Conceptual Model 

Site Hydrology. The study location is down gradient from a 

Superfund site in Massachusetts in which historical on-site 

disposal of arsenic-bearing wastes has resulted in leaching 

of arsenic into site ground water (Durant et al., 1990; 

Davis et al., 1994). A portion of the ground-water plume 

discharges into a constructed shallow lake that also receives 

surface-water inputs from a perennial stream (Halls Brook) 

on its western edge and an intermittent runoff channel at 

its northwestern end (Figure 4). The constructed shallow 

lake was built for flood control and is referred to as the Halls 

Brook Holding Area (HBHA) Pond throughout this report. 

The HBHA Pond has a maximum depth of approximately 5 

meters, and it discharges into a heavily vegetated wetland 

area. The HBHA Pond discharge reconstitutes Halls Brook, 

which meanders through the wetland until discharge into 

the Aberjona River. Based on eight measurements during 

the period of 7/1/1997 to 1/27/1998, Wick et al. (2000) 

determined a range in surface-water inputs and HBHA 

Pond outflow of approximately 48,380-459,090 ft3d-1, 

0-423,776 ft3d-1, and 118,304-953,496 ft3d-1 for Halls 

Brook, the intermittent runoff channel, and the HBHA 

Pond outflow, respectively. Assuming steady state for the 

lake water balance, this resulted in an estimated total flow 

input from ground-water discharge ranging from 27,898 to 

185,755 ft3

d-1 (ranging from approximately 7% to 60% of 

the water fl ux). This ground-water discharge corresponds 

with an independent estimate of approximately 91,818 

ft3

d-1 based on surface-water flow measurements at the 

Halls Brook discharge into the HBHA Pond and the point 

of discharge into the Aberjona River (Aurilio et al., 1994). 

These estimates are higher than ground-water discharge 

rates determined using seepage meters installed along 

the center-line of the lake (706 ft3

d-1; Davis et al., 1996). 

However, based on salt balance calculations, Wick and 

others (1998a, 2000) have estimated that the contaminated 

ground-water plume contributes 1,412-16,951 ft3

d-1 to the 

HBHA Pond. Visual observations of ground water seeps 

along the northern edge of the HBHA Pond (Aurilio et al., 

1994) and salt balance calculations within the HBHA Pond 

(Wick et al., 2000) indicated that contaminated ground 

water primarily discharges into the northern end of the 

HBHA Pond. 

Figure 4. (A) Aerial photograph and (B) aerial schematic of study site including approximate locations of ground-water 

and surface-water monitoring points. The major sources of water input into the pond include discharge from 

1) Halls Brook, 2) site-derived ground water with overlapping plumes of arsenic and BTEX compounds (approximately 130 and 100 m of shoreline, respectively), and 3) an intermittent runoff channel. 

11
Monitoring Strategy. Based on this information, the site 

characterization effort for this study was focused towards 

defining the spatial extent of contaminated ground-water 

discharge into the northern portion of the HBHA Pond. 

The spatial location of ground-water and surface-water 

monitoring points is shown in plan-view relative to the 

configuration of the pond (Figure 4). Tubing wells were 

employed at multiple depths adjacent to and within the 

lake to capture the vertical extent of the GW/SW transition 

zone. A combination of tubing wells, diffusion samplers, 

and depth-resolved sampling within the pond was used to 

measure water chemistry parameters in-situ and collect 

water samples for laboratory analyses. Tubing wells consisted of six-inch stainless steel screens attached to the end 

of Teflon-lined tubing. The multi-level tube cluster (NML) 

consisted of a series of ten six-inch screens with attendant 

Teflon-lined tubing attached to a PVC pipe anchored in the 

shallow aquifer. Sampling from tubing wells was carried 

out under low-flow sampling conditions (Puls and Paul, 

1995). The locations of key monitoring points including the 

elevation at which water samples were collected within the 

HBHA Pond and adjacent aquifer are shown in Figure 5. 

Discrete monitoring points within the HBHA Pond boundary were used to assess the magnitude of time-dependent 

variations in arsenic concentration. Water chemistry data 

collected from ground-water monitoring points adjacent 

to the HBHA Pond perimeter indicated that the highest 

arsenic concentrations were being discharged proximate 

to well-cluster location TW07. 

Arsenic Plume Characteristics. Historical data collected 

from the site indicated that ground water contaminated with 

arsenic and hydrocarbons discharges into the lake along 

the north-northeastern portion of the HBHA Pond (Davis et 

al., 1994;Wick and Gschwend, 1998b). Site-derived ground 

water contains elevated concentrations of ferrous iron and 

sulfate (Davis et al., 1994), which leads to the production of 

iron oxides and iron sulfi des in oxic and anoxic portions of 

the HBHA Pond, respectively. Shallow ground-water seeps 

are distributed along the northern boundary of the HBHA 

Pond as evidenced by the formation of orangish-red iron 

oxide precipitates where reduced ground water comes in 

contact with air. These shallow oxidized sediments contain 

arsenic concentrations ranging from 500-850 mg/kg based 

on measurements in this study. The ground-water input 

represents a significant fraction of the volumetric flow into 

the lake (Wick and Gschwend, 1998a; Wick et al., 2000). 

Contaminants discharging into the lake via the ground-water 

input potentially interact with aqueous and solid-phase 

components in the sediments and water column. Signifi cant 

portions of the site under study consist of former wetland 

areas in-filled with native soil materials and waste products 

from historical industrial activities. The general chemistry 

of the shallow unconsolidated aquifer is thus reflective of 

the high productivity commonly observed in wetland areas 

with significant rates of organic carbon turnover. In general, 

site ground water is reducing due to this native productivity 

and from anthropogenic factors such as degradation of 

buried waste materials (Davis et al., 1994). Ground water 

Figure 5. (A) Cross-sectional view of surface elevation through the northern end of the HBHA Pond and the upgradient 

aquifer. (B) Relative depths of sampling intakes for ground water and surface water are shown for locations 

TW10, TW07, TW02, and NML. (C) Aerial photograph showing the cross-section used to construct the diagram in Panel A. Land surface elevation at locations TW10 and TW07 is 67.25 and 58.75 feet above see 

level, respectively. A scale used to track water level in the HBHA Pond is installed at the location marked 

‘culvert’. 

12
is characterized by high concentrations of iron, sulfate, and 

organic carbon (natural and anthropogenic). Published data 

for sediment samples collected from the lake indicate that 

a fraction of the arsenic transported across the GW/SW 

transition zone is partitioned to sediments (Aurilio et al., 1994; 

Davis et al., 1996; Wilkin and Ford, 2002; Ford, 2004). 

An initial conceptual model was developed around the assumption that transformations of redox-sensitive elements 

such as iron, sulfur, and carbon would in part govern the 

speciation and transport of arsenic across the GW/SW transition zone. Differences in specific conductance between 

surface-water and ground-water inputs into the HBHA Pond 

result in nearly continuous chemical and physical stratification of the water column throughout the year (Wick et al., 

2000). Seasonal profiles of specifi c conductance, dissolved 

oxygen, and oxidation-reduction potential measured in 

the HBHA Pond water column during this study support 

this observation (Figure 6). It was hypothesized that the 

chemocline that develops within the lake water column 

would maintain distinct oxic and anoxic zones at shallow 

and deep depths, respectively. It was assumed that iron 

oxides derived from oxidation-precipitation of ferrous iron 

supplied continuously from discharging ground water would 

control the concentration of soluble arsenic within the lake 

water column. However, the fate of arsenic partitioned to 

iron oxides formed in the presence of dissolved oxygen 

near the chemocline was not known. Data from sediment 

characterization by Wilkin and Ford (2002) indicated that a 

fraction of the arsenic carried down to the sediments with 

settling iron oxide particles would ultimately partition to 

authigenic sediment minerals such as iron sulfides. There 

was no direct method to identify whether dissolved arsenic 

measured in the water column was derived from groundwater discharge or desorption/dissolution of arsenic from 

iron oxides deposited onto the reduced sediments. Thus, 

delineation of the relative contributions of arsenic to the 

lake water column from direct ground-water discharge or 

release from previously contaminated sediments was critical towards developing an understanding of arsenic fluxes 

across the GW/SW transition zone. 

Identifying Arsenic Sources to Surface-Water 

Receptor 

Monitoring the recovery of the HBHA Pond water column to 

steady-state conditions following a surge in surface-water 

input was used to assist delineation of arsenic sources. 

During Spring 2001 of our sampling campaign, a significant 

rainfall event coupled with snow melt resulted in a major 

flow event within the Aberjona watershed. Surface-water 

flow data for a monitoring station at Winchester, MA (USGS 

01102500 approximately 2.5 miles down gradient of the 

HBHA Pond), are shown in Figure 7 to document the relative 

intensity of this flow event. Depth-resolved sampling of the 

HBHA Pond water column was carried out over a period of 

several months to observe the recovery to pre-storm stratification. The time trend in specific conductance measured 

for ground water collected at the TW10, TW07, and TW02 

sampling locations is illustrated in Figure 8. These data indicated little variation in ground-water specifi c conductance 

over a period of approximately 1.5 years. However, there 

is a slight reduction in specific conductance for sampling 

Figure 6. Conceptual diagram depicting important hydrologic and chemical processes controlling water column chemistry within the HBHA Pond. Predominant inputs of water into the surface-water body include site-derived 

ground water and discharge from Halls Brook. Iron and sulfate reduction processes are active at the sediment-water interface and within shallow sediments. Representative vertical profiles of specific conductance 

(Sp. Cnd.), dissolved oxygen (DO), and redox potential (Eh) are shown to illustrate the chemical stratification 

that prevents mixing of shallow oxic and deep anoxic water (April 4, 2000; north end). 

13
14

Figure 7. (A) Daily streamfl ow for USGS 01102500 Aberjona River at Winchester, MA, located approximately 2.5 miles 

downgradient of the HBHA Pond. Open circles show the NML sampling dates following the March 2001 

storm event, and fi lled triangles show sampling dates for the TW07 and TW10 ground-water wells. Inset (B) 

shows NML and TW sampling dates relative to surface-water fl ow in greater detail. Two apparent phases of 

the post-storm water column recovery in the HBHA Pond are noted with labels ‘1’ and ‘2’ in inset (B). Inset 

(C) shows relative water levels measured at a fi xed scale mounted to the culvert on the eastern side of the 

HBHA Pond (see Figure 4 for culvert location).

Figure 8. Trends in specifi c conductance with depth for ground-water upgradient to NML: (A) TW10, (B) TW07, and (C) 

at TW02 below the HBHA Pond sediments adjacent to NML (fbgs = feet below ground surface). The dates 

for maximum fl ow observed at the USGS 01102500 station and the subsequent NML sampling dates are 

shown in Panel C with open and closed triangles, respectively. The inset in Panel C documents changes 

in the specifi c conductance depth profi le in the south end of the HBHA Pond as a result of the March 2001 

peak surface fl ow. Error of duplicate measurements is within the size of the data symbol.
locations TW10-1, TW10-2, TW07-2, and TW07-3 during the 

period of March 29 to April 3, 2001, immediately following 

the major flow event (Figure 8). Also documented in this 

figure is the change in HBHA Pond stratification for three 

sampling dates that bracket the major flow event (inset to 

Figure 8C). The nearly continuous chemical stratification of 

the HBHA Pond water column was interrupted as a result 

of the March 2001 peak surface flow. 

Data collected from upgradient ground-water wells during 

the period of observation indicated a relatively constant 

concentration of arsenic discharging into the HBHA Pond 

(Figure 9). Monitoring well TW10 is adjacent to a channel 

used to route storm water runoff from a large parking facility 

(see Figure 4). The channel is lined with rip-rap to minimize 

erosion, but it is hydraulically connected to the adjacent 

aquifer. Changes in arsenic concentration at downgradient 

monitoring locations TW07-2, TW07-3, and TW02 showed 

little or no response to upgradient surface-water recharge 

(Figure 9B and 9C). The inset in Figure 9C displays the 

vertical distribution of dissolved arsenic within the HBHA 

Pond water column (south) for the same sampling dates 

shown in the inset of Figure 8C. These data confirm interruption of the chemical stratification within the HBHA Pond 

immediately after the March 2001 peak surface flow. Water 

column monitoring was subsequently used to track recovery 

of the specifi c conductance and dissolved arsenic vertical 

profiles within the north end of the HBHA Pond adjacent to 

the primary discharge of arsenic-laden ground water. 

Time-series monitoring data collected from a tubing well 

cluster (NML) installed within the northern portion of the 

HBHA Pond immediately downgradient from ground-water 

monitoring cluster TW07 are shown in Figure 10. The relative vertical locations of surface-water and ground-water 

monitoring points are shown at the right-hand side of the 

figure. Vertical profiles of specific conductance are shown 

in the left panel for five sampling dates from April 5, 2001, 

to September 13, 2001. The greatest variation in specific 

conductance occurred over a depth interval of approximately 

8-11 feet below the water surface, which brackets the approximate depth of ground-water monitoring point TW07-2. 

The water column remained stratifi ed in this portion of the 

HBHA Pond since it is north of the primary contributor of low 

conductivity surface water (Halls Brook; see Figure 4), and 

it is close to the main discharge point for high conductivity 

ground water. However, specific conductance and other 

water parameters were sufficiently depleted due to storm 

Figure 9. Trends in total dissolved arsenic concentration with depth for ground-water upgradient to NML: (A) TW10, 

(B) TW07, and (C) at TW02 below the HBHA Pond sediments adjacent to NML. The dates for maximum 

flow observed at the USGS 01102500 station and the subsequent NML sampling dates are shown in Panel 

C with open and closed triangles, respectively. The inset in Panel C documents changes in the dissolved 

arsenic depth profile in the south end of the HBHA Pond as a result of the March 2001 peak surface flow. 

Error of duplicate measurements is within the size of the data symbol. 

15
16

September 13, 2001, and this increase was most signifi cant 

near the sediment-water interface. Insignifi cant changes 

in dissolved arsenic were observed in sediment pore water 

(NML-9) and the underlying shallow aquifer (NML-10; TW02 

in Figure 9C) at this location within the HBHA Pond. The 

concentration of dissolved arsenic at the sediment-water 

interface (NML-8) exceeded the concentrations observed at 

all proximate ground-water monitoring locations (compare 

Figures 9 and 10). For this reason, it was proposed that two 

separate processes were contributing arsenic to the HBHA 

Pond water column: 1) direct ground-water discharge and 

2) dissolution of fresh arsenic-bearing suspended solids 

deposited within the reducing sediments subsequent to 

the March 2001 peak surface fl ow. Based on the arsenic 

vertical profi les, it appeared that ground-water discharge 

dominated arsenic fl ux into the northern end of the HBHA 

Pond during the period April 5 to at least May 14, 2001 

(Phase 1 in Figure 7). After May 14, 2001, the arsenic fl ux 

was derived from a combination of ground-water discharge 

and the apparent dissolution of arsenic-bearing iron oxides 

that had settled into the anoxic portion of the water column 

(Phase 2 in Figure 7). This data set would have been 

strengthened by additional sampling at the NML location 

during the period of May-September 2001. However, the 

water dilution to allow observation of changes in water 

chemistry following the peak fl ow event. According to specifi c 

conductance measurements, the pre-storm stratifi cation in 

the water column was re-established by May 14, 2001, or 

earlier. Variations in specifi c conductance observed beyond 

this date within the water column and the underlying shallow 

aquifer (see Figure 10 caption) were smaller in magnitude 

and representative of non-storm fl ow conditions. The stable 

specifi c conductance profi le indicates re-establishment of 

a new steady-state mixing profi le between ground-water 

and surface-water inputs into the HBHA Pond.

In contrast, evolution of the vertical distribution of dissolved 

arsenic did not parallel that observed for specifi c conductance (Figure 10, right panel). Assuming that specifi c 

conductance can be used as a conservative indicator for 

water chemistry in this system, the observed changes in 

the vertical distribution of arsenic indicated the infl uence 

of processes other than mixing of different water sources. 

Up to May 14, 2001, the vertical distribution of arsenic 

appeared to evolve in a fashion similar to that observed 

for specifi c conductance. The greatest change was again 

observed over the depth interval of approximately 8-11 feet. 

However, arsenic concentrations continued to increase up to 

Figure 10. Water column recovery following the March 2001 peak surface fl ow event based on specifi c conductance 

and dissolved arsenic measurements. The vertical location of the NML sampling ports is shown relative 

to upgradient (TW07) and adjacent (TW02) ground-water sampling locations (see aerial view in Figure 4). 

Error of duplicate measurements is within the size of the data symbol. The mean and standard deviation 

for specifi c conductance at NML-2, NML-6, NML-8, and NML-10 for the period of April – September 2001 

are 804 ± 136 (n=5), 8006 ± 2964 (n=5), 13665 ± 515 (n=4), 15445 ± 876 (n=4) μS/cm, respectively. For 

comparison, the mean and standard deviation for specifi c conductance at TW07-2 and TW02 for the period 

of April 2000 to September 2001 are 960 ± 312 (n=5) and 8852 ± 342 (n=5) μS/cm, respectively.
17

for the system. However, the estimated values for arsenic 

fl ux indicate that sediment dissolution can account for a 

signifi cant fraction of the total mass of dissolved arsenic 

within the HBHA Pond water column. 

Revised Site Conceptual Model

The patterns in ground-water and surface-water solute 

chemistry can be used to revise the initial site conceptual 

model. Upgradient ground water is clearly a source of arsenic into the HBHA Pond. A fraction of the arsenic derived 

from ground water is partitioned to sediments along the 

margin and the bottom of the pond (400-1500 mg As/kg for 

sediments at the sediment-water interface over the aerial 

extent of the HBHA Pond). However, patterns in dissolved 

As and Fe observed in the north end of the HBHA Pond 

indicated that an internal recycling process contributed 

a fraction of the observed concentration of these solutes 

within the intermediate and deep portions of the pond. The 

elevated concentration of As observed at the NML sampling 

location on September 2001 was likely derived from dissolution of As-bearing iron oxides deposited at the bottom 

of the HBHA Pond. A general schematic of arsenic fl uxes 

into and out of the HBHA Pond water column is shown in 

Figure 11. A key component that contributes to the arsenic mass balance within the water column is the internal 

recycling of arsenic originally derived from ground-water 

discharge (Ford, 2004). 

relative stability of measured specifi c conductance above 

the sediment-water interface and the underlying aquifer 

indicated that the monitoring data suffi ciently captured the 

endpoints of the transition following post-storm recovery.

Estimates of the magnitude of the arsenic fl ux from groundwater discharge and dissolution of arsenic-bearing iron 

oxides formed within the HBHA Pond water column can 

be used to assess the relative importance of these two 

sources of arsenic to the water column. Assuming a crosssectional area for the arsenic-bearing plume of 852 ft2

(approximately 426 ft of shoreline over a depth interval of 

2 ft between sampling locations NML-4 and NML-7) and 

an average arsenic concentration of 1.5 mg As L-1, an 

instantaneous arsenic aerial fl ux of 0.0018 mg As L-1 ft-2

was derived from ground-water discharge at the north 

end of the HBHA Pond. The fl ux of arsenic from sediment 

dissolution/desorption can be estimated assuming that 

all arsenic is derived from a depth within the HBHA Pond 

corresponding to the location of NML-8. A bathymetric 

survey of the HBHA Pond (data not shown) indicated an 

approximate sediment area of 3,726 ft2

 at this depth within 

the north end of the pond. Thus, the instantaneous aerial 

fl ux of arsenic derived from sediment dissolution/desorption 

on September 13, 2001, near NML-8 (approximately 5 mg 

As L-1) was 0.0013 mg As L-1 ft-2. The reader is cautioned 

that these are estimates that represent a snapshot in time 

Figure 11. Revised conceptual model illustrating the arsenic fl ux balance within the HBHA Pond water column and emphasizing the importance of internal recycling of arsenic between the water column and sediments. Inputs 

of arsenic into the water column (shown with a plus sign) include discharge of site-derived ground water and 

dissolution/re-suspension of contaminated sediments. Removal of arsenic from the water column (shown 

with a minus sign) is due to a combination of discharge at the pond outlet and arsenic removal during iron 

oxidation-precipitation (HFO = hydrous ferric oxide) and settling at the chemocline and partitioning of arsenic 

to reduced sediments during diagenesis and burial.
The internal recycling process is due to the coupling of 

iron oxidation-reduction processes that, in part, control the 

distribution of arsenic between water and solids within the 

water column (Figure 11). Observations within the HBHA 

Pond water column indicate that ferrous iron generated 

from dissolution of iron minerals diffuses upward from within 

the water column (Process 1). When diffusing ferrous iron 

encounters dissolved oxygen at the chemocline within the 

water column, it is oxidized and precipitated (Process 2). 

Poorly crystalline iron oxides collected within the HBHA 

Pond at the chemocline contain signifi cant concentrations 

of arsenic, since these precipitates efficiently sequester 

arsenic at the pH within the HBHA Pond (Ford, 2002). The 

iron oxides subsequently settle back to the sediment-water 

interface and are subject to iron reduction and dissolution, 

again releasing coprecipitated/adsorbed arsenic (Process 

3). A fraction of the arsenic derived from settled suspended 

solids is likely incorporated into the sediments (Process 

4). A similar process has been documented in lake and 

stream sediments where iron oxides were abundant (Harrington et al., 1998; Nagorski and Moore, 1999; Senn and 

Hemond, 2002). 

Conclusions and Implications for Site 

Characterization 

The importance of arsenic partitioning to Fe-bearing solids within the HBHA Pond impacts both the approach to 

characterizing the GW/SW transition zone and assessing 

the fate of arsenic within the surface-water body. Temporal 

and spatial measurements of arsenic within ground water 

and surface water coupled with knowledge of arsenic 

partitioning to solids within the HBHA Pond demonstrated 

that ground-water discharge was not the only contributor 

to dissolved arsenic concentrations within the pond water 

column. Measured specifi c conductance proximate to the 

sediment-water interface at the NML monitoring location and 

adjacent shallow aquifer locations served as a constraint for 

delineating an arsenic flux at the sediment-water interface 

that was separate from direct ground-water discharge. This 

observation supported the importance of recognizing the 

distinct properties of sediments in direct contact with the 

surface-water body as well as the aquifer solids that may 

exist within the GW/SW transition zone (as illustrated in 

Figure 1). The apparent instability of arsenic associated 

with sediments at the bottom of the HBHA Pond also indicates that elimination of the ground-water arsenic source 

will not necessarily eliminate dissolved arsenic within the 

lake water column. The potential for sediments to serve 

as a secondary source of arsenic to surface water is an 

important factor to be recognized with respect to selection of remedial alternatives and evaluation of subsequent 

performance monitoring data. Due to the importance of 

inorganic contaminant sorption to solids, the approach to 

characterization of the GW/SW transition zone employed 

at this site has general applicability to other sites impacted 

by inorganic contaminants in ground water. Specifically, 

the value of detailed temporal monitoring during hydraulic 

transients within the site can provide useful information for 

delineating coupled processes that are difficult to discern 

within these complex systems. 

Overall, the behavior of arsenic at this contaminant site 

illustrates the level of complexity one may anticipate for 

evaluating fate and transport of contaminants across the 

GW/SW transition zone. This and other studies indicate 

that site monitoring to support a reliable conceptual model 

must properly address spatial and temporal variations 

that are a common trait of this environmental setting. The 

density of data collection will be dictated by the complexity 

of the geomorphic setting as well as the variability in local climatic conditions that can influence both the system 

hydrology and biogeochemistry. The field study presented 

here and the literature studies highlighted earlier all point to 

the need to monitor during base-flow conditions and during 

times when system hydrology is perturbed from base-flow, 

e.g., during a storm cycle. Observation of the response 

and recovery of the GW/SW transition zone to significant 

perturbations to the hydrologic system could provide useful insight into the system functionality. First, this type of 

monitoring data will provide a realistic assessment of the 

range in the magnitude and directionality of contaminant 

flux that can be anticipated for the system. In addition, 

for complex systems with multiple potential contaminant 

sources this type of monitoring data may provide the most 

straightforward means to delineate the relative importance 

of the various contaminant sources and the dominant processes controlling contaminant transport into or out of the 

surface-water body. 

Acknowledgement 

Richard Wilkin, Frank Beck, Jr., Patrick Clark, Thomas 

Holdsworth, Cynthia Paul, and Joseph LeMay provided 

valuable assistance and guidance during field sampling. 

Ning Xu and Sandra Saye provided analytical support for 

determination of metals, and Victor Murray provided support 

for GIS applications under Contract #68-C-98-138. Martha 

Williams provided support for document formatting under 

Contract #68W01032, Task Order 2018. This manuscript 

benefi ted significantly from critical and constructive reviews 

from Bart Faulkner (USEPA), Douglas Kent (USGS), Sonia 

Nagorski (University of Alaska Southeast), Donald Rosenberry (USGS), Rebecca Schneider (Cornell University), and 

representatives from the USEPA Ground Water Forum and 

Ecosystem Risk Assessment Forum. 

Notice 

The U. S. Environmental Protection Agency through its 

Office of Research and Development funded the research 

described here. This research brief has been subjected 

to Agency's peer and administrative review and has been 

approved for publication as an EPA document. Mention of 

trade names or commercial products does not constitute 

endorsement or recommendation for use. 

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