United States
Environmental Protection


The Impact of Ground-Water/Surface-Water Interactions on
Contaminant Transport  with  Application to an Arsenic
Contaminated Site

Robert Ford*

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 ground-
water 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 defined 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 estab-
lishing 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 con-
taminant exposure to the receptor, it is important to define contaminant mass distribu-
tion (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 thefactors controlling contaminant transport
and distribution between interacting ground-water and sur-
face-water bodies will be guided by the conceptual model
that delineates the relevant hydrologic and biogeochemi-
cal 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 dif-
ferent 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 (El) assessment
programs administered through the RCRA Corrective Ac-
tion (CA) Program and the Superfund  Program provide an
example wherethe Agency provides guidance for evaluating
the importance  of ground water/surface water (GW/SW)
interactions for contaminanttransport and exposure. These
indicators are used to monitor intermediate progress in en-
vironmental 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). Underthe RCRA CA Program, the Migration
of Contaminated Groundwater Under Control El (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
pdf).  Guidance documents developed to support these
El assessments point towards the  potential need to col-
lect 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 dis-
chargeof non-point contaminant sourcesfrom 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 assess-
ing 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 Re-
sponse  (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 hydrogeologi-
cal, 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 Re-
gional 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/Surface-
Water 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/oper-
ate field sites to support long-term research in the area of
GW/SW interactions. However,  several site-specific case
studies are docu mented withinthe1999 OS WE R-sponsored
workshop on GW/SW interactions (EPA, 2000a) that can be
consulted as points of reference for designing a monitor-
ing 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

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 hydro-
geological investigations 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
2)  Toxic Substances Hydrology  Program
3)  Shingobee Headwaters Aquatic Ecosystems Project
4)  Long-Term Ecological Research Network
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 iden-
tification (hydrologic or biotic-abiotic chemical processes)
in guiding design of the monitoring program for assessing
contaminant transport and distribution. Three components
are identified within a hypothetical boundary encompassing
the ground-water and surface-water bodies that exchange
waterand the chemical components mobile within this phase.
The three system components include 1) the ground-water
aquifer, 2) the GW/SW transition zone, and 3) the surface-
water 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
                                                          GWSW Titntftion ZfiM
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 G W/SW transition zone can further be de-
           lineated 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.

either component.  Differences in the relative water level
between these two components will dictate both thedirection
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 in-
habit 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 gradi-
ents 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,
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-watersediments. While contami-
nant partitioning to surface-water sediments may mitigate
transport into the surface-water body, sediments that have
accumulated a significant mass of a recalcitrant contami-
nant 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 surface-
water sediments may be controlled by  light- or tempera-
ture-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 zoneof light influence in buried surface-water
sediments or aquifer sediments within the GW/SW transi-
tion 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.
Contaminanttransportand, 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 waterto products with lower/greatertoxicity.
Development of an understanding of the processes that
control contaminant mass distribution withintheboundaries
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 encom-
passes 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  contami-
nants, 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 accumulateonsolid 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 contami-
nant concentration (C) in the mobile phase (water and mobile
colloids) and the volumetric flow of the mobile phase (Q).
                  M = C*Q
Contaminant flux can be calculated for point locations or
cross-sectional areas perpendiculartowaterfluxdepending
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
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, con-
taminant flux is dependent on water flow and contaminant

                               M = C*Q

Figure 2.   Conceptual diagram illustrating potential spatial heterogeneity and variability in the magnitude of contaminant
           flux across the G W/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 react ions occur
rapidly relative to the timescaleforfluid transport. The latter
situation is commonly assumed to apply to ground-water
systems, but this needs to be confirmed during site char-
acterization. This is particularly important for near-surface
systems that may experience large variability in water flux
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 rela-
tive 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 Final Creek Basin, Arizona.
Manganese  was partially sequestered from contaminated
ground water within the hyporheic zone due to the down-
ward 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

studies illustrate the importance of understanding the bal-
ance 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
ai. (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 dis-
solution 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 bio-
geochemical 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/immobiliza-
tion 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 trans-
port across the GW/SW transition zone is improving with
increased publication of site-specific field studies in the
literature. The literature reviewthat 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 charac-
teristics of GW/SW interactions for various flow regimes,
and the reader should consider alternate sources for ad-
ditional 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.
Dissolution of
Sorption and Ion
Gas Dissolution and
Relevant Process
Acid neutralization
by aqueous
carbonate alkalinity
Precipitation of
metal sulfide
Ion exchange on
dissolution of iron
oxide coupled to
organic carbon
Benzene oxidation
coupled to
Ammonia gas-
water exchange
Example Reaction
HC03 + H+ = H2C03°
Zn2+ + HS = ZnS(s) + H+
KAISi308(s) + MM/ = NH4AISi308(s) + rC
4Fe(OH)3(s) + 8H+ + CH2O = 4Fe2+ + CO2(g) +
C6H6 + 6N03- + 6H+ = 6C02(g) + 6H2O + 3N2(g)
NH3(g) + H2o = NH; + OH-

Table 2.    Examples from Peer-Reviewed Literature Documenting Field Case Studies Designed to Assess the Physicochemical Processes In-
           fluencing Inorganic Contaminant Transport Across the GW/SW Transition Zone.  OC = organic carbon, DOC = dissolved organic car-
           bon, EDTA = ethylene daminetetraacetic acid.
Nitrate, Zinc
Nitrate, Zinc
Nitrate, Zinc
Iron, Zinc
Great South Bay,
New York
Seine River,
Elbe River, eastern
Western Sierra
Nevada foothills,
Florida Keys
Tidal freshwater
marsh, Virginia
Oder River, north-
eastern Germany
Lot River,
Bordeaux, France
Lot River,
Bordeaux, France
Glatt River,
Glatt River,
Silver Bow Creek,
Silver Bow Creek,
Several lakes,
Mountains, New
Hydrologic Setting
Submarine ground-water discharge to
coastal marine ecosystem
River-water infiltration into ground
River-water infiltration into ground
Ground-water discharge into stream;
temporal variations in mixing of
bedrock and soil drainage sources
Shallow and deep ground-water
discharge into second-order stream
Ground-water discharge to coastal
canal system
Marsh ground-water infiltration-
River-water infiltration into ground
River-water infiltration into ground
River-water infiltration into ground
River-water infiltration into ground
River-water infiltration into ground
River-water/ground-water infiltration-
River-water/ground-water infiltration-
Lake/ground-water infiltration-
Chemical Processes
Denitrification (N2 or N2O production) and dissimilation
(NH^-N production) in sediments
Nitrification-denitrification coupled to microbial
mineralization of aqueous and solid OC; ammonia
release from OC decay
Denitrification using oxidizable aqueous and solid OC
from river water and river bed/aquifer sediments
Nitrification of ammonia; leaching of nitrogen
compounds from soil solids
Denitrification coupled to pyrite oxidation
Near-conservative transport of low-salinity plume;
denitrification in off-axis portions of plume
Iron oxidation-precipitation limits diffusive and
advective phosphate transport into marsh
Denitrification and Mn reduction-oxidation coupled to
microbial degradation of OC
Weathering of Mn-bearing minerals coupled to
bacterial degradation of OC; Mn(ll) oxidation coupled
to sorption-precipitation
Weathering of Mn-bearing minerals coupled to
microbial degradation of OC - spatial variability due to
different OC or mineralogical Mn sources; Mn(ll)
oxidation coupled to sorption-precipitation
Mineral weathering/precipitation coupled to microbial
degradation of OC; Cu-EDTA complexation; metal
sorption-precipitation due to fluctuations in abiotic
chemical equilibria; denitrification
Mineral weathering/precipitation coupled to microbial
degradation of OC; Cu-DOC complexation; metal
sorption-precipitation due to fluctuations in abiotic
chemical equilibria; denitrification
Oxidative precipitation of Fe and Mn; precipitation-
sorption of Cd, Cu, Zn; denitrification
Dissolution of As-bearing iron oxides from sediment
deposited in reducing hyporheic zone
Anaerobic microbial OC degradation, mineral
weathering and ion exchange reactions
Capone and
Bautista, 1985
Doussan et al.,
Grischek et al.,
Holloway and
Dahlgren, 2001
Grimaldi et al.,
Griggs et al.,
Chambers and
Odum, 1990
Massmann et al.,
Bourg and Berlin,
Bourg and Berlin,
Jacobs et al.,
von Gunten et al.,
Benner et al.,
Nagorski and
Moore, 1999
Sebestyen and
Schneider, 2004

         Table 3.   Examples from Peer-Reviewed Literature Documenting Field Case Studies Designed to Assess the PhysicoChemical Processes Influencing
                   Organic Contaminant Transport Across the GW/SW Transition Zone. EDTA = ethylene diaminetetraacetic acid, NTA = nitrilotriacetic acid,
                   A/DC = naphthalene dicarboxylate, APEC = alkylphenol polyethoxy carboxylates, HAA = haloacetic acid, CFC = chlorofluorocarbon.
1 ,4-Dichlorobenzene,
1 ,3-Dimethylbenzene,
Volatile, Semivolatile,
Chloroform, Benzene,
Volatile Organic
CFC-11.CFC-12, CFC-
Glatt and Aare Rivers,
Sava River, northern
Croatia, Yugoslavia
Watson Creek,
Aberdeen Proving
Ground, Maryland
Inner Piedmont, South
Glatt and Sitter Rivers,
northern Switzerland
St. Joseph, Michigan
Santa Ana River,
Orange County,
West Branch Canal
Creek, Aberdeen
Proving Ground,
Pine River, Angus
Ontario, Canada
Pine River, Angus
Ontario, Canada
Everglades National
Park, south Florida
Coastal northeastern
Hydrologic Setting
River-water infiltration into
ground water
River-water infiltration into
ground water
Ground-water discharge
into stream
Ground-water discharge
into stream
River-water infiltration into
ground water
Ground-water discharge-
recharge into lake with
wave infiltration
River-water infiltration into
ground water
Ground-water discharge-
recharge into wetland and
Ground-water discharge
into river (short circuiting
and low-high dispersed
Ground-water discharge
into river (short circuiting
and low-high dispersed
Surface-water infiltration
into ground water
Ground-water discharge
into a stream
Chemical Processes
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
Inferred biological degradation and/or sorption
onto solid OC in surface water and aquifer
Not studied; passive samplers used to
delineate contaminant plume discharge zone
Not studied; passive samplers used to locate
rock fractures that were conduits for
contaminant discharge
Biological degradation of organic compounds
primarily within GW/SW transition zone
(shallow aquifer)
Chloroethene co-oxidation by methane-
oxidizing bacteria
Microbial degradation of NTA, APEC and HAA
Anaerobic biodegradation (hydrogenolysis
and reductive dechlorination) under reducing
conditions (iron-reducing, sulfate-reducing,
Reductive dehalogenation producing cis-1 ,2-
dichloroethylene and vinyl chloride (low-
moderate discharge zone)
Reductive dehalogenation producing cis-1 ,2-
dichloroethylene and vinyl chloride (low-
moderate discharge zone)
CFC degradation coupled to microbial
Biological degradation in non-acidic,
anaerobic portions of plume
Schwarzenbach et
al, 1983
Ahel, 1991
Vroblesky et al.,
Vroblesky et al.,
Aheletal., 1996
Lendvay et al.,
Ding et al., 1999
Lorah and Olsen,
Conant Jr., 2000
Conant Jr. et al.,
Happell et al.,
Lawet al., 2004

Summaries of published studies from the peer-reviewed
literaturethat documentthe influenceof 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 specia-
    tion and transport;
2)  Chemical gradients of major and trace elements/con-
    taminants 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 waterf low (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, watertransport due to evapotranspi-
ration can also exert a significant influence on ground-water
flow patterns (Doss, 1993; Fraseretal., 2001). Water flow
dynamics  influence contaminant transport directly due to
changes in the volume and direction of waterflux. However,
changes in waterflowalso exert an indirect  effect by altering
major element chemistry that can stimulate or quench bio-
geochemical 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/orsulf uroxidation-
reduction 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 fluid flow
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 transi-
tion 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 at-
tenuation/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 sorp-
tion or retardation. Failure to monitor water and sediment
concentrations within the streambed could have resulted
in a  misinterpretation of low surface water PCE concen-
trations 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 dischargef rom 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 solid-
phase 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 communi-
ties.  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 nor-
mal physiological functions (e.g., Jaynes and Carpenter,
1986) or indirectly by hosting microbial communities that
influence 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

contaminants that are absorbed/adsorbed during transport
across the transition zone.
Redox Chemistry. The importance of redox processes for
organicand inorganic contaminant transport and fate in sub-
surface systems at waste sites has been addressed in detail
in a recent Agency publication (EPA, 2000b).  Developing
an understanding of the relevant redox processes control-
ling 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
                               J PTOCMBM
Figure 3.   Conceptual diagram showing snapshotof chem-
           ical component profiles across the GW/SW
           transition zone. Redox reaction zones develop
           within fixed spatial regions due to biogeochemi-
           cal reactions maintained by chemical fluxes.
           Conceptual adaptation of illustration  from
           Figure  7 in Bourg and Berlin (1993). NOM =
           natural organic matter, TEAP = terminal elec-
           tron 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 de-
termining 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 studies illustratethe impact of redox-driven processes
on the transport of  inorganic and organic  contaminants
across the GW/SW transition zone. The evaluations docu-
mented in these reports provide a useful context for evaluat-
ing 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 po-
tential 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 ac-
cumulation 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 appor-
tioning risk to the various contaminant source terms and
for evaluating whether intervention is required to  manage
estimated risk from sediment contamination.

The following field study provides an example where sur-
face-watersediment 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 forsite 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 ft3d'1 (ranging from approximately  7% to 60% of
        the water flux). This ground-water discharge corresponds
        with an independent estimate of approximately 91,818
        ft3d~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 ft3d1; 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 ft3d~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.
                     Image derived from May 1995 aerial
                     photograph obtained from MassGIS.
                                                    I Arsenic Plume
                                                i	• Hydrocarbon Plume
• Tube Wells (TO)
O Multi-level Tube Sampler (NML)
                                                                      (Halls Brook)  (,\\
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 (ap-
           proximately 130 and 100 m of shoreline, respectively), and 3) an intermittent runoff channel.

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 con-
sisted 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 bound-
ary 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; Wickand 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 sulfides 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 intothe lake via the ground-water
input potentially interact with aqueous  and solid-phase
components in the sediments and water column.  Significant
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 dia-
           gram 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

is characterized by high concentrations of iron, sulfate, and
organiccarbon (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 (Aurilioetal., 1994;
Davis et al., 1996; Wilkin and Ford, 2002; Ford, 2004).

An initial conceptual model was developed around the as-
sumption that transformations of redox-sensitive elements
such as iron,  sulfur, and carbon would  in part govern the
speciation and transport of arsenicacross the GW/SWtran-
sition zone. Differences in specific conductance between
surface-water and ground-water inputs into the HBHA Pond
result in nearly continuous chemical and physical stratifica-
tion of the water column throughout the year (Wick et al.,
2000). Seasonal profiles of specific 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 ground-
water 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 criti-
cal towards developing an understanding of arsenic fluxes
across the GW/SW transition zone.
Identifying Arsenic Sources to Surface-Water
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 todocument 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 strati-
fication. The time trend in specific conductance measured
for ground water collected at the  TW10, TW07, and TW02
sampling locations is illustrated in Figures. These data in-
dicated little variation in ground-water specific conductance
over a period of approximately 1.5 years.  However, there
is a slight reduction in specific conductance for sampling
                                     Sp. Cnd. (nS/cm)  DO(mg/L)    Eh (mV)
                                       0 4000 8000 0   4  8 0  200 400
                                                                       Ground Water
                                                                       High conductivity
                                                                       As, Fe(ll), SOf organic carbon
                                                                       Suboxlc or anoxic
                                                                       \ Water Column

                                                                       | Sediment

                                                                       I Aquifer
Figure 6.   Conceptual diagram depicting important hydrologic and chemical processes controlling water column chem-
           istry 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 sedi-
           ment-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).

Figure 7.   (A) Daily streamflow 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 filled triangles show sampling dates for the TW07 and TW10 ground-water wells. Inset (B)
           shows NML and TW sampling dates relative to surface-water flow 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 fixed scale mounted to the culvert on the eastern side of the
           HBHA Pond (see Figure 4 for culvert location).
                                                             -•-TW10-1 (10.58fbgs)
                                                             -ffl-TW10-2 (15.58 fbgs)
                                                             -D- TW10-3 (20.58 fbgs)
                                                              -•-TW07-1 (12.25 fbgs)
                                                              -0-TW07-2 (15.00 fbgs)
                                                              -O- TW07-3 (20.00 fbgs)
                                                               o NML Sampling Dates
                                                               • Peak Surface Flow
                                                               v HBHA Pond South
                                                                  Sampling Dates
                                   3/1/2000  9/1/2000  3/1/2001
Figure 8.
Trends in specific 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 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 specific conductance 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.

locations TW10-1,TW10-2,TW07-2, and TW07-3during 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 inter-
ruption of the chemical stratification within the HBHA Pond
immediately afterthe March 2001 peak surface flow. Water
column monitoring was subsequently used totrack recovery
of the specific 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 rela-
tive 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
conductanceoccurredoveradepth interval of approximately
8-11 feet below the water surface, which brackets the ap-
proximate depth of ground-water monitoring point TW07-2.
The water column remained stratified 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
1 -
1 -
1 -
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n 	 n ^ ^\
t-e-- 	 _ _e
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— • ^ •
O-O- -O ^ O ^
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T -£-Sep2001
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-•- TW07-1
^ TW07-2
-O- TW07-3

o NML Sampling Dates
• Peak Surface Flow
v HBHA Pond South
Sampling Dates

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.

                                                    i GW Discharge
                Water Surface
                                                  k2) Sediment Dissolution
2 -
4 -
6 -

£ 12-

1 Q
lo I

• April 5, 2001
* A April 10, 2001
4 * May 14, 2001
K * May 31, 2001
% 	 • 	 September 12-13, 2001
"•'•• '"''•'"Ğ.:.
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0 4000 8000 12000 16000

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D TW07-1
D D TW07-2

D TW02
             Specific Conductance, u,S/cm
Dissolved As, ppm
Figure 10.  Water column recovery following the March 2001 peak surface flow event based on specific 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 specific conductance at NML-2, NML-6, NML-8, and NML-10 for the period of April - September 2001
           are 804 ħ 136 (n=5), 8006 ħ 2964 (n=5), 13665 ħ 515 (n=4), 15445 ħ 876 (n=4) pS/cm, respectively.  For
           comparison,  the mean and standard deviation for specific conductance at TW07-2 and TW02 for the period
           of April 2000 to September 2001 are 960 ħ312 (n=5) and 8852 ħ 342 (n=5) pS/cm, respectively.
water dilution to allow observation of changes in water
chemistry followingthepeakflow event. According to specific
conductance measurements, the pre-storm stratification in
the water column was re-established by May 14, 2001, or
earlier. Variations in specific conductance observed beyond
this date within the watercolumn and the underlying shallow
aquifer (see Figure 10 caption) were smaller in magnitude
and representative of non-storm flow conditions. The stable
specific conductance profile indicates re-establishment of
a new steady-state mixing profile 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 specific conduc-
tance (Figure 10, right  panel).   Assuming that specific
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 influence
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 specific conductance. The greatest change was again
observed overthe depth  interval of approximately 8-11 feet.
However, arsenic concentrations continued to increase upto
September 13,2001, and this increase was most significant
near the sediment-water interface.  Insignificant changes
in dissolved arsenic were observed in sediment pore water
(NML-9) andthe 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). Forthis 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 flow.  Based on the arsenic
vertical profiles, it appeared that ground-water discharge
dominated arsenic flux 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 flux
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

relative stability of measured specific conductance above
the sediment-water interface and the underlying aquifer
indicated that the monitoring data sufficiently captured the
endpoints of the transition following post-storm recovery.

Estimates of the magnitude of the arsenic flux from ground-
water 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 cross-
sectional 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 flux of 0.0018 mg As L~1 ft"2
was  derived from  ground-water discharge at the north
end of the HBHA Pond. The flux 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
flux of arsenic derived from sediment dissolution/desorption
on September 13, 2001, near NML-8 (approximately 5 mg
As L1) was 0.0013 mg As L1 ft'2. The reader is cautioned
that these are estimates that represent a snapshot in time
                                     for the system. However, the estimated values for arsenic
                                     flux  indicate that sediment dissolution can account for a
                                     significant 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 ar-
                                     senic 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 dis-
                                     solution of As-bearing iron oxides deposited at the bottom
                                     of the HBHA Pond. A general schematic of arsenic fluxes
                                     into  and out of the HBHA Pond water column is shown in
                                     Figure 11. A key component that contributes to the arse-
                                     nic mass balance  within the  water column is the internal
                                     recycling of arsenic originally derived from ground-water
                                     discharge (Ford, 2004).
                      Surface Water
                      (Dissolved &
                    Suspended Solids)
                   Internal Recycling Process
As-HFO(s) +  H2S(aq) (or other reductants)

         İ j

   (or other Fe(ll)-bearing minerals
       and organic matter)
                                                             Water Surface
   I       Iron      '
r-J Oxidation-Precipitation
I   |    and Settling
                                                            Solids Dissolution
                                                           and/or Re-suspension
                                                                           - Chemocline
                                                                              3*    Water
                                                                              Q   Discharge
Figure 11.  Revised conceptual model illustrating the arsenic flux balance within the HBHA Pond water column and em-
           phasizing 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 significant 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 (Har-
rington et al.,  1998; Nagorski and Moore, 1999; Senn and
Hemond, 2002).
Conclusions and Implications for Site

The importance of arsenic partitioning to Fe-bearing sol-
ids 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 specific 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 separatefrom direct ground-waterdischarge. 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 indi-
cates 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 selec-
tion 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 lo-
cal 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 use-
ful 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 pro-
cesses controlling contaminant transport into or out of the
surface-water body.
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
benefited significantly from critical and constructive reviews
from Bart Faulkner (USEPA), Douglas Kent  (USGS), Sonia
Nagorski (University of Alaska Southeast), Donald Rosen-
berry (USGS), Rebecca Schneider (Cornell University), and
representatives from the USEPA Ground Water Forum and
Ecosystem Risk Assessment Forum.
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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