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Journal of Hydrology, 6 3 (1983) 1-29
Elsevier Science Publishers B.V., Amsterdam - Printed in The Netherlands
#
1
MIGRATION OF CONTAMINANTS IN GROUNDWATER AT A
- NDFILL: A CASE STUDY
houndwater Flow and Plume Delineation
'
D.S. MacFARLANE, J.A. CHERRY, R.W. GILLHAM and E.A. SUDICKY
' Dep~~rtmentof Earth Sciences,
..
University of Waterloo, Waterloo, Ont., N2L 3G1 (Canada)
(Accepted for publication July 13, 1982)
ABSTRACT
MacFarlane, D.S., Cherry, J.A., Gillham, R.W. and Sudicky, E.A., 1983. Migration of
contaminants in groundwater a t a landfill: a case study, 1. Groundwater flow and
plume delineation. In: J.A. Cherry (Guest-Editor), Migration of Contaminants in
Groundwater at a Landfill: A Case Study. J. Hydrol., 63: 1-29.
A landfill-derived contaminant plume with a maximum width of -600 m, a length of
-700m and a maximum depth of 20 m in an unconfined sand aquifer was delineated
by means of a monitoring network that includes standpipe piezometers, multilevel pointsamplers and bundle-piezometers. The extent of detectable contamination caused by the
landfill, which began operation in 1940 and which became inactive in 1976, was determined from the distributions of chloride, sulfate and electrical conductance in the sand
aquifer, all of which have levels in the leachate that are greatly above those in uncontaminated groundwater. The maximum gmperature of grpndwater in the zone of
contamination beneath the landfill is 12 C, which is 4-5 C above background. The
"
tnermal plume in the aquifer extends "150 m downgradient from the centre of the landfill. A slight transient water-table mound exists beneath the landfill in the late spring and
summer in response to snowmelt and heavy rainfall. Beneath the landfill, the zone of
leaclhate contamination extends to the bottom of the aquifer, apparently because of
transient downward components of hydraulic gradient caused by the water-table mound
and possibly because of the higher density and lower viscosity of the contaminated water.
Values of hydraulic conductivity, which show variations due to local heterogeneity, were
obtained from 'slug tests of piezometers, from pumping tests and from laboratory tests.
Because of the inherent uncertainty, in the aquifer parameter values, the 38-yr. frontal
position of the plume calculated using the Darcy equation with the assumption of plug
flow can differ from the observed frontal position by many hundreds of metres, although
the use of mean parameter values produces a close agreement.
The width of the plume is large relative to the width of the landfill and can be accounted
for primarily by variable periods of lateral east- and westward flow caused by changes in
water-table configuration due to the variable nature of recharge. Northward from the
landfill, the vertical thickness of the plume decreases and the top of the plume is farther
below the water table. The thickness of the zone of uncontaminated groundwater above
the plume increases northward as the area of recharge of uncontaminated water downflow from the landfill increases. Because dispersion in the vertical direction is weak, there
is very little mixing between the overlying zone of recharge water and the contaminant
plume. Concentration profiles are irregular beneath and near the landfill and become
smooth downgradient where the maximum concentrations are much less than those
beneath landfill. These features are attributed to a strong influence of longitudinal
dispersion. The plume passes beneath a small shallow stream near the landfill without
significant influence on the stream.
0022-1694/83/$03.00
@
1983 Elsevier Science Publishers B.V.
INTRODUCTION
The most common means of "managing" municipal refuse is by deposition
in sanitary landfills. In humid and semi-humid regions, infiltration through
landfills normally results in the migration of leachate from the refuse into
underlying groundwater zones. Studies of landfills on unconsolidated sand
and gravel aquifers by Golwer et al. (1975), Palmquist and Sendlein (1975),
and Kimmel and Braids (1975, 1980) have established that zones of leachatecontaminated groundwater can extend many hundreds of metres. In some
circumstances, leachate contamination can cause serious deterioration of
aquifers used for groundwater supply, as in the example described by Apgar
and Satherthwaite (1975). In other cases the contaminated zone may not
pose a significant hazard to useable water resources or to the ecological
system in the area. This can be the case because of favourable paths of
migration or because of attenuation of the contaminants by physical, chemical and biochemical processes.
In the selection process for new landfill sites and in the evaluation of the
hazard posed by existing landfills, hydrogeologists have the task of developing interpretations of the site conditions and of predictions of the future
behaviour of zones of contaminated groundwater. Consequently, knowledge
of the behaviour of landfill-derived contaminants in groundwater at representative sites that have been investigated in exceptional detail is desirable.
This paper is the first in a series of seven papers published in this Special
Issue, that pertain to an intensive investigation of an abandoned sanitary
landfill on an unconfined aquifer comprised of glacio-fluvial sand. Aquifers
of this type are common in the glaciated region of North America. The
abandoned landfill is located within the confines of the Canadian Forces
Base (C.F.B.), Borden, -80 km northwest of Toronto, Ontario (Fig. 1).
The objective of this paper is to describe the physical hydrogeological
conditions at the site and to relate these conditions to the existing shape
and extent of the zone of contaminated groundwater in the aquifer. This
paper provides a hydrogeological framework for the investigations described
in the other papers in this series, which pertain specifically to the dispersive
characteristics of the aquifer (Egboka et al., 1983; Sudicky et al., 1983),
to the hydrogeochemical behaviour of inorganic contaminants (Dance and
Reardon, 1983; Nicholson et al., 1983), to the use of geophysical methods
for detection and delineation of contaminated zones (Greenhouse and
Harris, 1983). Sykes and Farquhar (1980) have described numerical simulations of contaminant migration in the aquifer.
Several types of monitoring devices were used to obtain the data upon
which this paper is based. These devices are described briefly in this paper
and in detail by Cherry et al. (1983) in the second paper in this Special
Issue. These devices, which consist of normal standpipes, piezometers,
piezometer nests, piezometer bundles and multilevel point samplers were
installed in the aquifer during the period of 1974-1980. Most of the
Fig. 1. Site location and topographic setting.
hydrologic and contaminant concentration data upon which this paper is
based were obtained during 1979/1980. The interpretations also take into
account information obtained by other investigators during previous field
seasons. The data used in this paper include stratigraphic logs, measurements
of hydraulic head, temperature and electrical conductance of the groundwater, chloride and sulfate concentrations in the groundwater and measurements of hydraulic conductivity by several methods.
PHYSIOGRAPHY, CLIMATE AND GEOLOGY
The topography in the vicinity of the landfill is flat to undulating, with
local elevations at the site ranging from a high of 232m above sea level
(a.s.1.) at the landfill to a low of 222.5m in the abandoned sand pit to the
north of the landfill (Fig. 1). The landfill has an area of 5.4 ha and forms a
step-shaped mound. The south,.east and west sides of the mound have steep
slopes, and the northern side slopes much more gently.
The area to the south, east and north of the landfill is covered by oak,
pine, birch and scrub brush. A grassed field covers the western and northwestern part of the area. An abandoned flat-bottomed sand pit, which provided cover material for the landfill, forms a broad depression in the northern
part of the site. A small northward-flowing intermittent creek passes within
40m of the eastern edge of the landfill (Fig. 1). Seasonal marshes occupy
the low areas to the south and east of the landfill and part of the sand pit.
A climatic station, from which the following data were obtained, was
maintained at the town of Angus until 1970 (Gartner Lee Associates Ltd.,
1977). The mean daily temperature ranges from 12.4" to 0 . 6 " ~ ,averaging
6.5"C, with extremes of +36.7" and - 4 1 . 1 ~ ~ . Mean total precipitation
averages 82.8cm yr.-', of which 58.67cm was in the form of rain, and
240.5 cm was snow. A typical year has 1 7 1 days of frost, 8 8 days of rain and
56 days of snow.
HISTORY O F LANDFILLING AND SITE STUDIES
Landfilling operations at the Borden site spanned a period of 36 years,
from 1940 to 1976. Gartner Lee Associates Ltd. (1977) indicate that the
landfill formed as refuse was dumped over the edge of a minor sand ridge
and that the filled area expanded to the south. The site had little or no
onsite management from 1940 to 1973. During this period, refuse was
periodically burned and then covered with sand. Layers of ash were encountered during exploratory drilling in the deepest parts of the landfill.
The landfill ranges in thickness from 5 t o 10m, with the thickest refuse in
the north and northwest sections. The most rapid input of refuse likely
occurred during the period of World War I1 when the Borden community
1
1
I
I
,
['
1
1
'
i ~ c ~rapidly
w
to a peak population. In 1973, modern landfilling practices
ware initiated and were continued until the operation was terminated in
1976. The final landfill surface was covered with -0.5 m of sand. From
I
,
1
1
I
horeholes and test pits, Gartner Lee Associates Ltd. (1977) established
!hat -80% of the refuse deposited during the life of the landfill operation
ronsisted of ash, wood and construction debris. The remaining refuse is
~~Onlposed
mainly of domestic and commercial food wastes.
Hydrogeological studies at the Borden landfill began in 1974 when Gartner
Loe Associates Ltd. installed 51 piezometers and water-table standpipes
; ~ t26 locations. The purpose of this monitoring was to determine, at a
rt~connaissancescale, the influence of the landfill on the quality of ground:r.ntcr in the unconfined sandy aquifer. In this phase of the study, it was
t~stahlishedthat although the quality of water in the unconfined aquifer was
~ l ~ , ~ ~ athe
d e dcontaminated
,
zone was separated by an extensive zone of clay
and silt from a deeper aquifer used for water supply. The long-term influence
of the landfill on the shallow aquifer and the local stream was not predictable
;it that time and it was decided to close the landfill in favour of a new site on
niuch less permeable deposits elsewhere on the base. A long-term program of
research at the abandoned landfill was initiated in 1976 by Gartner Lee
.\ssociates Ltd. and the University of Waterloo with support provided by
government agencies. In this year, the spatial distribution of the contaminated zone was delineated in detail by measurements of electrical conductance
groundwater samples obtained using an auger-head sampler designed
q1ecifically for the task (Anderson, 1977; Cherry et al., 1983). An additional
69 piezometers and water-table standpipes were installed during this investicative phase.
GROUNDWATER MONITORING DEVICES AND SOIL SAMPLING
Between the initial studies in 1974 and the end of the 1979 field season,
78 water-table standpipes, 56 standpipe piezometers and 14 bundle-piezot~leters were installed at the Borden site (Fig. 2). Twenty-nine multilevel
uoundwater samplers of the type described by Pickens et al. (1978) were
also mstalled in 1978/1979 for sampling of groundwater at numerous depths
nt each monitoring site. More bundle-piezometers were installed in 19801
1981. The standpipes and standpipe piezometers were used primarily for
monitoring of water levels. The piezometer bundles were used for water-level
monitoring and for water sampling at many depths in the aquifer.
The standpipes, standpipe piezometers and multilevel samplers were
~nstalledusing truck-mounted drill rigs with hollow-stem flight augers. For
most installations, a rubber stopper was inserted in the opening in the auger
bit to prevent the entry of sand into the auger stem during drilling. This
vnabled drilling to proceed without the use of the center rods and without
use of drilling water. When the auger flight was advanced to the desired
NEST OF TWO OR THREE
PIEZOMETERS OR STANDPIPES
PUMPING T E S T SITE
Fig. 2. Location of multilevel-point samplers, bundle-piezometers, pumping test sites and
piezometer nests.
depth, it was raised a few decirnetres and the piezometer or multilevel
sampler was lowered through the auger stem and then used t o knock out
the rubber stopper. As the augers were removed, the cohesionless sand
collapsed tightly around the pipe.
Water was not used during installation of the multilevel samplers. During
drilling of deeper holes for installation of the bundle-type multilevel piezometers in 1979, foreign water was used in the augers a t a few of the drill
sites. At depths exceeding 20 m, the rubber stopper did not keep formation
sand out of the auger, and a column of water was maintained in the annular
space t o equalize the hydraulic pressure and prevent sand inflow.
For the standpipe piezometers installed in the clayey-silt zone below the
aquifer, a sand pack was placed around the screened interval. A bentonite
seal was installed above the sand pack. The sand and bentonite were introduced through the hollow stem of the augers.
The piezometers and standpipes designated as the M-, 77- and P-series
piezometers (see Fig. 5), are constructed of 1.9-, 2.2- and 3.2-cm I.D.
PVC pipe, respectively. Screens consist of sawed slots covered with fiberglass or nylon cloth t o prevent excessive infilling of silt or clay. The piezometer screens are 0.3-0.6 m long and the standpipes are screened from the
bottom of the pipe t o above the water table.
In 1979/1980, fourteen multilevel bundle-piezometers, with a total of
124 sampling points labelled T-series, were installed t o various depths. Each
bundle is comprised of eight 1.0-cm ($-in.) I.D. polyethylene tubes bound
t o a central rigid 1.3-cm (&-in.)I.D. PVC pipe. A 15-cm (6-in.) long perforated
plastic screen wrapped with nylon mesh was attached t o the end of each
tube. These piezometers were installed through the hollow stem of t!e flight
augers in the manner described above. Where water was used during installation, representative chemical samples could not be taken for several weeks
due to contamination of the area immediately adjacent to the screens. Details
of the construction, installation and advantages and disadvantages of the
instrumentation used at Borden are discussed by Cherry et al. (1983) in Part
2 of this series of papers.
At many of the sites at which piezometers were installed, occasional
split-spoon samples were taken to supplement the borehole description
obtained from auger samples. At six locations, cores from numerous depths
were obtained by Gartner Lee Associates Ltd. (1978) using a universal
split-tube sampler and a Dames & Moore-type cohesionless soil sampler.
Additional stratigraphic information was obtained from geophysical logs of
rotary-drilled mud.-filled holes (J.P. Greenhouse, pers. commun., 1980).
GROUNDWATER MONITORING AND HYDRAULIC CONDUCTIVITY TESTS
During the early stages of development of the monitoring network, water
levels were measured in standpipes and piezometers by Gartner Lee
Associates Ltd. (December 1974 and March 1977) and by the University of
Waterloo (April/May 1978). Water levels in the M-, 77- and P-series wells
were measured eleven times over the course of a 12-month period in 1979
using an electric tape. Water levels in the bundle-piezometers were measured
to the nearest 2 or 3 mm using a narrow-diameter electric tape consisting of
a 4-mm diameter coaxial cable.
During the period of August 8-11,1979, -600 water samples for analysis
of electrical conductance and chloride were collected from the piezometers,
standpipes and multilevel samplers. The sediment in all samples was removed
by centrifuging prior to analysis. Fifty-two samples were collected in August
1979 and 53 inApril 1980 for analysis of sulfate concentration.
Using a group of sample bottles connected to a manifold attached to a
vacuum bottle, samples were obtained 6 or 8 at a time from multilevel
samplers (S-series) and piezometer bundles (T-series). Before each sample
was taken, three volumes of -150 cm3 were pumped from each point to
remove any stagnant water. Samples were obtained from the PVC standpipes
and piezometers by means of a flexible rubber tube connected t o a handoperated vacuum pump.
Temperature profiles were measured on August 10,1979 and June 5,1980,
to determine the effect of the landfill on the thermal regime of the groundwater zone. The thermal resistance of the groundwater was determined using
a YSI@ model 44031 precision thermistor and a digital ohm meter.
Falling-head and rising-head response tests for hydraulic conductivity were
performed on many of the piezometers. Falling-head tests were conducted
by monitoring the rate of decline of the water level after the water level was
rapidly displaced upward by the lowering of an aluminum rod into the
water. After equilibrium was achieved, a rising-head test was performed by
monitoring the recovery of the water level to its original level after the rod
was quickly removed.
Detailed pumping tests were conducted at three locations where a pumping
well and numerous observation piezometers were located. One of these sites
is located near the southwestern side of the landfill and the other two are
located in or near the old sand quarry north of the landfill (Fig. 2). At the
first site the pumping well penetrates only the upper third of the aquifer. At
the other sites the pumping wells penetrate the full aquifer thickness.
AQUIFER PROPERTIES
Because of its glaciofluvial origin, the unconfined aquifer is locally very
heterogeneous due to complex distributions of beds and ,lenses of fine-,
medium- and coarse-grained sand. Some of the beds have a considerable silt
content. The unconfined aquifer lies on an extensive deposit of clayey and
sandy silt. This deposit is a regional stratigraphic unit that varies in thickness
from 15 to 3 0 m in the Borden area (Burwasser and Cairns, 1974). The top
of the clayey silt layer at the bottom of the aquifer was identified at numerous
locations because the augers yielded distinct samples of the material even
when they penetrated only a short distance into this zone. The aquifer is
thickest in the southwestern part of the area and thins to -9.5m in the
northern end of the area (Fig. 3). Beneath the landfill, the surface of the
clayey silt increases in elevation northward. In the bottom part of the aquifer
in the vicinity of the landfill a few beds of clayey silt or silty clay were
observed in some of the boreholes. These beds .were less than 0.5m in
thickness and in some cases were less than 0.1 m thick. Because they are so
thin, they may not have been detected in some of the deep boreholes. It is
therefore not known if these beds are continuous in this area. The sand in
the bottom part of the aquifer appears to be more silty than elsewhere in
the aquifer but this change is not vely distinct. Where the aquifer is much
thinner northward of the landfill, there is no trend toward much higher
silt or clay contents deeper in the aquifer, except within a metre or so of the
top of the clayey-silt unit. At this depth the contact between the sand
aquifer and the clayey silt is gradational and contains some thin beds of
silty sand, silt or clayey silt.
Values of hydraulic conductivity for the aquifer sand were obtained
from grain-size analyses, from water-level response tests in piezometers, from
pumping tests and from a permeameter test on two disturbed core samples
(Table I). From grain-size analyses of samples from four boreholes, hydraulic
conductivity values were obtained using the methods of Allen Hazen (in
Lambe and Whitman, 1969), and of Fair and Hatch (1933) and Masch and
Denny (1966). The Hazen method utilizes the effective grain size (dl,,)
100
-
50
0
SAND. MEDlUH AND FINE GRAINED
loom
VERT. E X l G ' 10
SAND. WITH SILTY ZONES AND
THIN SILTY CLAY BEDS
CLAY, SILTY, PEBBLY
Fig. 3. Longitudinal geological cross-section along the main direction of groundwater flow.
TABLE I
Hydraulic conductivity (K) obtained from grain-size methods, permeameter tests, pumping tests and piezometer response tests
Apparent lithologic unit
( I ) Grain-size correlation (197711978)
fine- and medium-grained sand
silty fine-grained sand
(2) Permeameter tests (1978)
fine- and medium-grained sand
silty fine-grained sand
(3) Pumping tests (1979) (at three test
fine- and medium-grained sand
sites)
(4) Piezometer response
P-series wells (1978)
W-series wells (1979)
P-series wells (1979)
Piezometer nest near 77-34(1979)
W-series wells (1980)
P-series wells (1979)
fine- and medium-grained sand
silty fine grained
G = geometric mean, S = standard deviation, N = number of tests.
Ran e of K
(lo-'cms-l)
N
whereas the latter two methods include a measure of the actual distribution
in grain size.
Values of hydraulic conductivity of two samples from a borehole between
S8 and S9 were subjected to permeameter tests. One sample, which was
composed of clean medium-grained sand yielded a value of
cm s-' . A
cm s-' .
sample of silty fine-grained sand gave a value of
The results of water-level response tests are listed in Table I. The values of
the geometric mean and the arithmetic mean for the non-silty fine- and
medium-grained sand are all within the range of 3
-5.
cm s-' .
The fact that this range is small suggests that the non-silty segment of the
aquifer is relatively homogeneous when viewed on a large scale even though
numerous small-scale variations due to bedding are exhibited in samples
from each borehole.
The values of hydraulic conductivity obtained from the pumping tests at
the three sites are in the same range as the mean values obtained from the
piezometer response tests. At each pumping test site the drawdowns in
several observation piezometers were monitored as one well was pumped.
Values of hydraulic conductivity were computed from the drawdown data
for each observation piezometer. The results provided in Table I are representative results based on the values computed from the early time data from
the observation piezometers that provided the most detailed early-time
drawdown vs. time record. The analysis of drawdown data from the pumping
tests provided values of transmissivity, which were converted to hydraulic
conductivity by dividing the distance, fronthe-initisll-water table to the
bottom of the wellscreen. At the site,located at the southeastem edge of the
landfill wherethe aquifer is thick, the large degree of partial penetration of
the pumping well may cause considerable error in the computed values of
hydraulic conductivity. At the other sites located in the northern part of the
plume where the aquifer is thin, the aquifer conditions are nearly ideal for
the evaluation of pumping test data using standard methods of analysis
based on type curves.
A porosity value equal to 0.38 was calculated on the basis of the difference
between saturated and air-dried weight of relatively undisturbed cores
obtained from the site of the natural gradient dispersion test (Sudicky et al.,
1983). A dry bulk density equal to 1.76 g ~ m was
- ~ also obtained from the
core samples. In comparison to the range in hydraulic conductivities of the
sand, the expected range for porosity is small. The spatial variation is probably less than 10% of the value reported above.
GROUNDWATER FLOW SYSTEM
The water table is farthest below ground-surface in the landfill area and
in the area northeast of the landfill. The maximum water-table depth is
7--9 m. It is closest t o ground surface in the abandoned sand pit and in the
-
area south of the landfill, where it is generally less than 1m below surface.
The water-table fluctuation in 1979 was generally between 1.2 and 0.5 m,
with the smallest fluctuations occurring in the abandoned sand pit, near the
stream, and in the northernmost part of the study area.
The refuse in the landfill is entirely above the water table except from late
March to June, when the water table is at its highest because of recharge
from snowmelt and spring rain. Piezometers with tips at the bottom of the
refuse indicate that during this period the water table in part of the landfill
area exists within the bottom metre of refuse. The water table gradually
declines by 0.75-1 m during the summer and early fall and then rises by a
few decimetres in response to fall rainfall.
Contour maps of water-table elevation and longitudinal cross-sections with
isopotential lines are displayed in Figs. 4 and 5, respectively. The maps are
based on water-level measurements made on five datesin 1979. The locations
of the data points upon which the contours are based are shown in Figs. 4a
and 5a, which also indicate the position of the zone of contaminated groundwater in the unconfined aquifer. Isopotential diagrams for other cross-sections
and water-table maps for other dates of monitoring are presented by
MacFarlane (1980).
The contour lines on the water-table maps are relatively linear in the eastwest direction during the fall and winter months, as illustrated by the watertable map for October 1 7 (Fig. 4). During this time, groundwater flow in the
study area is northward. During the spring and summer, however, the watertable contours have an arcuate shape, which indicates that during this period,
groundwater flow in the aquifer is directed radially from the landfill. There
is northward flow from the central part of the landfill, with flow veering to
the northeast and northwest relative to the position of the cross-section
A-A'. The changes in water-table configuration from spring t o fall reflect the
formation of a water-table mound near and beneath the landfill in the
spring and its subsequent decay during the summer and complete disappearance in the fall and winter.
On May 9, when the highest water levels observed in 1979 occurred, a
small pond existed near the eastern edge of the landfill. This pond resulted
mainly from surface runoff from the landfill mound.
The isopotential diagrams for flow along the northsouth cross-section
(Fig. 5) are consistent with the trends indicated by the water-table contour
maps. During the fall and winter months, the isopotential lines are vertical
or near-vertical, which indicates horizontal flow northward in the aquifer.
During the spring and early summer, however, semi-circular isopotential
lines represent the hydraulic-head conditions beneath or very near the landfill. For example, the surface water ponding alongside the landfill in late
April and May is reflected in the semi-circular arrangement of isopotential
lines beneath the southern segment of the landfill on the April 27 and May
9 monitoring dates (Fig. 5). In the late spring and summer, the water-table
mound is situated beneath the landfill rather than beneath the periphery of
Fig. 4. Water-table maps, 1979 (metres above sea level, contour interval is 0.1 m).
the landfill and the semi-circular isopotential lines are also entirely beneath
the landfill, as indicated on the July 3 monitoring dates.
The diagram for this date also exhibit8 closed isopotential lines in the
mid-depth of the aquifer beneath the-landfill. This reflects a component of
groundwater flow that is not in the plane of the cross-section. This component of flow appears t o be generated by higher water-table levels beneath the
western part of the landfill, which, for a month or two, causes eastward flow
METRES (A.S.L.1
METRES (A.S.L.)
METRES ( A . S . L . )
METRES (A.S.L.)
METRES ( A . S . L . 1
METRES (A.S.L.)
n
r
I?.
C
3.
0
2
p
beneath much of the landfill as indicated by cross-sections prepared by
MacFarlane (1980) for other dates in the spring and early summer. During
the late summer and fall transverse cross-sections indicate little tendency for
flow eastward or westward.
DISTRIBUTION OF CHLORIDE, SULFATE AND ELECTRICAL CONDUCTANCE
Rainwater or snowmelt that infiltrates through the refuse at the Borden
landfill leaches constituents from the refuse. The leachate enters the groundwater zone below the landfill. During migration in the groundwater system,
Fig. 6. Maps of electrical conductance ( P S cm-' ) and chloride (mg 1-' ); contours based
on maximum values in the plume at each multilevel sampling site.
A'
-
w
I
200
-
SAMPLE POINTS
FOR CP-AND ELECTRICAL CONDUCTANCE
-
0
230
-
A'
A-
-
-
w
I
200
-
ELECTRICAL CONDUCTANCE (/AS)
AUGUST
1979
SULFATE CONTOUR ( m g l L )
-
1
Fig. 7. Electrical conductance (pS cm-' ), chloride (mg 1-' ) and sulfate (mg 1-' ) along the
longitudinal cross-section; the sampling points for chloride and electrical conductance are
shown in the upper diagram.
attenuation of contaminants occurs by one or more of the following mechanisms: dispersion, adsorption, ion exchange, precipitation, coprecipitation
and biochemical degradation. Some of the leachate-derived constituents are
transported by groundwater with little or no influence by chemical or biochemical processes. These constituents can be used as convenient indicators
of the maximum extent of readily-identifiable contamination. At the Borden
site, chloride and sulfate are in this category. Chloride is relatively nonreactive and, in the leachate: occurs at concentration levels much above the
natural groundwater. The natural groundwater has less than 1 0 mg 1-' ,whereas
groundwater that is strongly contaminated by leachate has more than
500 mg 1-' .
The results of the chloride analyses are shown in plan view in Fig. 6 b and
along cross-section A-A' in Fig. 7c. The maps of chloride (Fig. 6b) and
electrical conductance (Fig. 6a) were constructed using the maximum values
of each parameter obtained from each multilevel sampling site. Data from a
few shallow piezometers were used where there were no multilevel samplers
or piezometer bundles. On the cross-sections (Fig. 7b and c), only values
obtained for point intakes and piezometers with short screens (0.3 m) were
used (Fig. 7a).
The greatest chloride concentrations occur in the middle of the aquifer
50-160 m downgradient of the refuse, and the plume maximum consistently
occurs at or near an elevation of 214m a.s.1. Fig. 8 shows depth profiles of
CHLORIDE
CONCENTRATION
(mg/L)
4 0 0 m FROM LANDFILL
3OOm FROM LAWFILL
lSOm FROM LANDFILL
75m FROM LANDFLL
Fig. 8. Vertical profiles of chloride (mgl-' ) at sites along the longitudinal cross-section,
from near the landfill on the left t o greater distance from the landfill on the right.
chloride at selected multilevel samplers located along cross-section A-A'.
The chloride profiles are irregular near the landfill and become more uniform
with distance downgradient from the landfill.
At a late stage in the investigation, some emphasis was placed on the use
of sulfate in addition to chloride for delineation of the front of the contaminant plume. This was done to overcome difficulties in identification of
the frontal position of the plume. These difficulties arose due to road-salt
contamination of groundwater near the front of the plume. In contrast to
the use of chloride in landfill studies, sulfate is rarely used in this manner
because it commonly occurs only in very low concentrations in leachatecontaminated groundwater. In some situations, sulfate can undergo considerable concentration decline as a result of biochemical reduction that
converts SO:- to H, S or HS-. At the Borden site, sulfate concentrations in
the leachate-contaminated water are exceptionally high in most parts of the
plume. Biochemical reduction does not appear to be a significant mechanism
of SO:- attenuation in the plume downgradient of the landfill, although
some SO:- may be lost locally as a result of reduction beneath the southern
part of the landfill and near the clay at the bottom of the aquifer as described
by Nicholson et al. (1983).
The results of the sulfate analyses for selected bundle-piezometer and
multilevel samplers are presented on cross-section A-A' in Fig. 7d. Sulfate
concentrations range from 2060mgl-' under the landfill to background
levels, which are generally less than 25 mg 1-' . Oxygen-isotope analyses of
sulfate suggest that the sulfate is likely derived from construction materials
in the landfill (S. Feenstra pers. commun., 1980).
In plan view, both chloride (Fig. 6b) and electrical conductance (Fig. 6a)
show that contamination from landfill leachate encompasses a large area
downgradient from the landfill. The leachate-contaminated groundwater
extends -700 m north along the principal direction of flow, and as much as
200 m laterally t o the east and west of the landfill mound. The highest values
occur below the areas of thickest refuse in the landfill. The distributions of
chloride, sulfate and electrical conductance along the longitudinal crosssection (Fig. 7) are quite similar except beneath the southern part of the
landfill where sulfate is generally absent. The chloride plume extends downward from the refuse all the way to the bottom of the aquifer where the
clayey silt unit prevents further downward migration. Sulfate contamination
does not extend as deep as the chloride. Samples taken from within the
clayey silt were not above background chloride levels, which suggests that
the unit acts as an effective barrier to the downward movement of leachate.
contaminated groundwater. Distributions of C1-, SO:- and electrical conductance obtained during 1977/1978were generally similar to those presented
here (Dickin, 1979; Nicholson et al., 1983). The data sets from these earlier
years, however, have less detail.
Road-salt contaminates the shallow part of the aquifer near the extreme
downgradient end of cross-section A-A' where high values of electrical
I
conductance and chloride near the water table were observed. In the middle
" of the aquifer in this area, the values are much lower and at the bottom of
Sf
the aquifer, where the plume occurs, they are high.
1.
t
I
'
TEMPERATURE DISTRIBUTION
Biochemical processes within the landfill cause the landfill t o be much
warmer than the surrounding soil. In longitudinal cross-section (Fig. 9), a
zone enclosed by the llOccontour, exists within the contaminant plume
immediately downgradient from the landfill. The temperature drops to background values at a distance of -150-200 m downgradient from the landfill.
The zone of elevated temperature extends to the bottom of the aquifer
beneath the northern part of the landfill and corresponds to the zone of
deep penetration of the chloride and sulfate (Fig. 7). The existence of the
zone of elevated groundwater temperature extending a considerable distance
from the landfill indicates that the aquifer has sufficient capacity for heat
retention to allow the higher-temperature zone to persist even after many
years of travel away from the heat source at the landfill.
The results of the two temperature surveys (August 1979 and June 1980)
are nearly identical. The shallow groundwater upgradient from the landfill
exhibits a sharp temperature decline with depth. The shallowest groundwater is warmest in this area because the depth to the water table is sufficiently small for the spring and summer air temperatures to cause rapid
warming of the water. Similar temperature profiles exist downgradient of the
landfill beyond the main zone of landfill influence (Fig. 9). It is expected
that in the winter these areas have much lower temperatures at shallow
depth.The temperature regime within and beneath the landfill is probably
W
I
TEMPERATURE
VERT. EXAG.= I
0
OC
-
---Fig. 9. Temperature distribution along the longitudinal cross-section (contour intervals
0.5'~).
not influenced significantly by changes in seasonal air temperature. The
existence of a plume of warm water in a sandy aquifer downflow of a
landfill has also been documented by Kimmel and Braids (1980).
DISCUSSION
The hydrogeological information that is normally collected in assessments
of the potential impact of proposed landfills at new sites includes stratigraphic, hydraulic conductivity, and hydraulic-head data. The stratigraphic
and hydraulic-head data are used to develop predictions of the travel paths
that contaminants will follow. By means of the Darcy equation the hydraulic
conductivity data in combination with the hydraulic-head data provide
estimates of groundwater velocities. These velocity values or the velocity
values estimated for the new flow conditions that are expected to occur
during landfilling operations are normally used to predict the extent to
which a zone of contaminated groundwater will develop.
At the Borden site, predictions of this nature can be developed and they
can be compared with the known position of the contaminant plume. The
approach taken in this paper includes no quantitative analysis of the effects
of hydrodynamic dispersion, although attention is drawn qualitatively to the
apparent effects of dispersion. Quantitative analyses of dispersion are presented in the other papers in this Special Issue (Egboka et al., 1983; Sudicky
et al., 1983) and by Sykes and Farquhar (1980).
The large-scale stratigraphy of the site, comprising the sand aquifer overlying the silty-clay aquitard was easily determined at the earliest stage of the
site investigation by Gartner Lee Associates Ltd. (1975) who conducted
routine auger drilling. Although much more drilling was done in later years,
the essence of this initial stratigraphic interpretation was not altered. It was
known from the early work that the aquifer was locally heterogeneous and
that, at the larger scale, it represented a major permeable hydrogeologic
unit with a silty-sandy zone near the bottom.
The siltyclayey unit beneath the aquifer was identified as a relatively
impervious unit that separated the unconfined aquifer from deeper aquifers
used for water supply. From these observations it was concluded that the
landfill was not a threat to the quality of groundwater in deeper aquifers.
Our subsequent detailed investigation of the contaminant plume and monitoring of groundwater conditions in the uppermost part of the silty-clay
unit have provided confirmation for this early geologically-based conclusion.
The simplest approach for the development of predictions of the general
directions of contaminant migration in an unconfined aquifer is based on
the water-table configuration. In the early phase of the site investigation
(Gartner Lee Associates Ltd., 1975), the water table was monitored by the
M-series wells. These wells indicated a general northward slope of the water
table. A water-table map prepared in 1976 (Gartner Lee Associates Ltd.,
1
I
1
1
1977) does not differ appreciably from the general form of the maps presented in this paper based on data obtained in 1979/1980 from many more
monitoring sites. Thus, given the stratigraphic setting and a water-table map
based on relatively sparse data, northward migration of landfill-derived
contaminants in the aquifer was an obvious expectation that was confirmed
by the detailed monitoring undertaken in later phases of the study.
The next step in the assessment of the potential for contaminant migration
is to obtain estimates of the horizontal hydraulic gradients in the aquifer
based on the water-table slope. A representative value for the northward
slope of the water table across the study area is -0.003. The effective
porosity of the aquifer sand is -0.35, expressed as a fraction. From data on
hydraulic B d i e n t (i), effective porosity (n) and hydraulic conductivity (K),
values of the average linear groundwater velocity (5) in the northward direction were obtained from the relation:
The values for hydraulic conductivity from the permeameter test and
from the grain-size analyses of the fine- and medium-grained, relatively nonsilty sand are in the range of 6 10-3-10-2 cm s-' . For this type of sand the
geometric mean values of the hydraulic conductivity obtained from piezometer response tests are in the range of 3 lod3-5 lob3cm s- . Representative values obtained from the pumping tests at three sites are also in this
range (Table I). When the representative hydraulic gradient is used, the range
of %values for northward flow obtained from the range of hydraulic conductivity values from the grain-size analyses and permeameter test is 1728 m yr.-' . Based on the hydraulic conductivity values from the permeameter
test and the pumping tests, the range of Z is 8-14myr;'.
From 1940 to
1979 (approximately 38 yr. of travel time), the minimum estimated northward movement of the leachate front (5 = 8 m) would be 300 m and the
maximum (5= 28 m) would be 1060 m. The observed position of the northward front of the plume is -700m from the landfill. Considering the fact
that the sand aquifer has a considerable degree of small-scale heterogeneity
and the fact that the methods for determining hydraulic conductivity of
the aquifer vary in scale and have different sources of error, the computed
and observed travel distances are remarkably similar. In contrast to this
closeness of agreement, Kimmel and Braids (1980), in a study of two plumes
at landfills on an unconfined glaciofluvial aquifers, obtained calculated
frontal advance distances that were 3-4 times as large as the observed
frontal positions of the plumes.
The average linear horizontal groundwater velocity in the aquifer has also
been evaluated by means of a natural-gradient slug-injection tracer test
described by Sudicky et al. (1983) in Part 4 of this series and by boreholedilution tests of the type described by Freeze and Cherry (1979). The
natural gradient tracer test and nearly all of the borehole dilution tests were
conducted at a depth of 0.5-3 m below the water table in the uncontaminated
e
'
part of the aquifer above the plume in the old sand quarry or between the
sand quarry and the road (Fig. 2). These sites are within 100 or 200 m of
the front of the plume.
Because of the effect of local heterogeneity, the tracer zone in the natural
gradient test split into two distinct segments, the fastest segment travelled
northward at a velocity of 90 m yr.-' and the slowest at 25 m yr.-' . The bore,holedilution tests provided velocity values in the range of 30--90 m yr.-' .
- The velocity values from the tracer test and the boreholedilution tests
are greater than the values obtained from the Darcy equation, using mean or
representative hydraulic conductivity data. When the actual hydraulic
gradient in the northern part of the plume is used in the velocity estimates
from the Darcy equation; the velocity values are larger by a factor of 1.8
than the values obtained using the representative across-site gradient of
0.003. These velocity values are, however, still generally lower than the
tracer test and borehole dilution values. The significance of this difference
with respect t o the development of the plume is currently being investigated.
- The effect of hydrodynamic dispersion was not taken into account in the
travel-distance comparison described above. Dispersion may cause the front
of the plume to advance more rapidly than the average linear groundwater
velocity, as assumed by Egboka et al. (1983) in Part 3 of this series of papers,
or after a period of time, it may cause the apparent front to advance less
rapidly as the plume gradually evolves toward a quasi-steady state condition. The conditions necessary for a quasi-steady state to develop in
idealized plumes have been evaluated on a theoretical basis by Germain
(1981). It is not yet known which of these two conditions is most relevant
to the Borden plume. We expect that detailed monitoring of the frontal
position of the plume during a period of several years will provide a basis
for evaluation of this issue.
It was indicated on p. 21, that with even a small amount of hydrogeologic
data, it was evident that the contaminant plume from the Borden landfill
moves northward in the unconfined aquifer. However, prior to the actual
delineation of the plume by three-dimensional monitoring, there was little
basis t o expect in the northern area that the plume would become very thin in
the vertical dimension and so-distinctly occupy only the bottom half of the
aquifer. Northward from the landfill, an increasing percentage of the aquifer
contains water that has recharged the aquifer in the forested area downgradient from the landfill. This water occupies the upper part of the aquifer.
, The existence of a distinct zone of uncontaminated water above the plume
and the relatively abrupt transition between this zone and the plume indicates
that hydrodynamic dispersion in the vertical direction is a sufficiently
ineffective process to cause very little mixing of these two zones. Even a
moderate degree of vertical mixing would cause the plume to spread vertically through the aquifer. This would be expected because the aquifer is
thin and the distance of northward travel of the plume is long.
In the literature there is a paucity of information on the tendency for dispersion in the vertical direction when contaminant plumes travel horizontally
-
-
I
through sand or gravel aquifers. Using a finite-element model based on the
advection-dispersion equation, Pickens and Lennox (1976) simulated the
movement of laterally-migratingplumes along cross-sections in a hypothetical
thin horizontal homogeneous aquifer. Because they assigned large values for
the dispersivity in the vertical direction, these simulations showed extensive
vertical mixing of the plumes in a manner quite dissimilar to the Borden
plume. Sykes and Farquhar (1980) performed finite-element simulations of
the Borden plume along the main n o r t h s o u t h cross-section. They also used
large values for the dispersivity in the vertical direction and obtained simulated plumes with much more vertical mixing than we have identified from
the field data. Although these modeling efforts utilized large values for
vertical dispersivity, field evidence of very weak vertical dispersion, which
is consistent with the Borden observations, is presented by Kimmel and
Braids (1980), who monitored two large plumes at landfills in an unconfined
sand and gravel aquifer on Long Island, New York, U.S.A. The tracer test
for dispersion conducted at the Borden site by Sudicky et al. (1983) also
produced weak vertical dispersion. If weak dispersion in the vertical direction
is a characteristic feature of horizontally-migrating contaminant plumes in
glaciofluvial aquifers, it can be concluded that monitoring at numerous
depths in these aquifers is essential for delineation of the zones of highest
contaminant concentration. Evaluation of this generalization must await the
results of detailed field monitoring at other contaminant plumes in glaciofluvial aquifers.
Although the vertical dispersion of the plume is sufficiently weak to produce a relatively abrupt mixing zone between the plume and the overlying
recharge water, mixing between layers or lenses within the plume is sufficiently effective to cause distinct irregularities in vertical concentration
profiles to become gradually obliterated in the direction of the flow. This
trend is illustrated by the chloride profiles in Fig. 8, which shows very
irregular concentration profiles beneath and near the landfill and smooth
profiles at sites that are hundreds of metres north of the landfill. These
trends suggest that, on the scale of individual heterogeneities within the
aquifer, small-scale niixing in the vertical direction is an important process.
Relative to the thicknesses of the aquifer and the plume, however, this
process causes very little vertical spreading. A theory of small-scale interlayer
mixing in stratified aquifers is described by Gillham et al. (1983).
Although during much of the year, flow in the aquifer is essentially
horizontal from the area south of the landfill t o the area north of the
landfill (Fig. 4), the zone of leachate contamination beneath the landfill
has penetrated downward through the entire aquifer thickness to a maximum depth of -20 m below the water table. If horizontal flow was the only
influence on the distribution of the zone of contamination beneath the landfill, the plume would only exist at shallow depth in this area.
One major cause of the downward movement of the zone of contamination
beneath the landfill appears to be the transient downward-directed hydraulic
gradient beneath the landfill. Downward differentials in hydraulic head exist
during the spring and summer. Although these gradient components have a
significant magnitude for only a few months of each year, they appear to be
capable of causing many metres per year of downward flow.
A second possible cause of downward migration of the zone of contamination is the greater density of the contaminated water relative to the natural
groundwater in the aquifer. The effect of the density contrast is t o cause a
vertical driving force in addition to that caused by a downward hydraulichead differential. Because values of density are locally dependent on
both the concentration and temperature and because the vertical hydraulic
gradient varies with depth and time, a comprehensive analysis would require
modeling techniques that are not currently available. Nevertheless, by
separating the domain below the landfill in two regimes, an upper one consisting of higher average total dissolved'solids (TDS) and temperature and a
lower one comprised of low TDS and temperature, a preliminary assessment
of the possible effect of density can be instructive. Kimmel and Braids
(1980) used this approach t o rationalize the downward movement of warm
leachate into colder ambient groundwater in an aquifer in which the flow is
predominantly horizontal. With TDS of -- 2000 mg 1-' and a temperature
of l l ° C , contaminated water beneath the landfill has a mass density of
-- 1.0012 g ~ m - Uncontaminated
~ .
groundwater with 100 mgl-' TDS and a
temperature of 7' C has a mass density close to 1.000 g cm- 3 . The net density
drivingforce can be obtained from therelation A p / p o , where A p is the density
difference and po is a reference density (Bear, 1972). This expression yields
a value of 0.001. This driving force can be compared directly with the downward components of hydraulic gradient beneath the landfill, which are -- 0.01
in the spring and early summer and on the order of 0.001 or less during the
remaining months of the year. This indicates that in the spring and early
summer, the driving force due to density is small in comparison t o the downward hydraulic gradient caused by mounding of the water table beneath the
landfill. During the rest of the year, however, the density force is of the same
general magnitude or greater than the downward component of the hydraulic
gradient and is not much smaller than the average horizontal component of
the hydraulic gradient in the vicinity of the landfill.
A third effect that could contribute to downward movement of contaminated water is the lower fluid viscosity in the zone of elevated temperature
beneath the landfill. Temperature has a much larger effect on viscosity (or
resistance to flow) of water than it does on the density. Lower viscosity
causes the hydraulic conductivity t o be correspondingly higher. In the
Borden aquifer, this effect is small because a temperature difference of 4 ' ~
only causes a viscosity difference of 0.1 cP, which corresponds to a relative
increase in hydraulic conductivity of only 10%. Kimmel and Braids (1980)
also considered the viscosity effect to be small because the presence of an
electrolyte (e.g., NaC1) in water causes the viscosity of water to be relatively
invariable over a wide temperature range.
The movement of the plume deep into the aquifer beneath the landfill
and the development of a zone of recharged water above the plume have
resulted in little or no influence of the plume on the small stream that flows
northward past the eastern edge of the landfill (Fig. 2). Miniature bundlepiezometers were installed at several locations below the stream bed in the
manner described by Lee and Cherry (1979). The plume extends eastward 1
beneath the stream (Fig. 6) but it is deep enough to avoid interaction with
the stream. The stream acts as a groundwater divide but the divide only
affects the very shallow and local zone of groundwater flow near the stream. '
Another feature of the contaminant plume that warrants explanation is
its large width. Although the general direction of groundwater flow is northward, the plume has spread east- and westward to the extent that it is nearly
as wide as it is long (Fig. 6). Two processes have caused the plume to become
wide; the main cause is east- and westward groundwater flow and a secondary
cause is transverse dispersion. The maps of the water table (Fig. 4) and
transverse isopotential cross-sections (MacFarlane, 1980) indicate that for
many months of the year there are significant east- and westward components
of flow from the landfill and from the central part of the plume where the
water table is commonly slightly elevated above the areas to the east and
west. The water-table maps for spring and summer have arcuate contours
that reflect this condition. If hydraulic-head data had been obtained only
in the late fall and winter months, there would have been little or no indication that east- and westward flow components could cause the development of a wide plume. In contrast, data from the summer months are quite
indicative of this condition. Transverse dispersion in the horizontal plane
also contributes to the lateral spreading of the plume but to an extent that
cannot be determined.
A major difference between the hydrogeologic conditions at the Borden
site and those at the two landfills studied by Kimmel and Braids (1980) is
the presence at the Borden site of the transient water-table mound and
transient components of flow transverse to the general northward direction
of flow. At the sites on Long Island studied by Kimmel and Braids, no watertable mounding was observed. The water-table slope was uniform and had
little temporal variation. The two large contaminant plumes at these sites
exhibited no significant spreading transverse to the direction of the watertable slope, so that the plumes were long and thin in plan view, rather than
fan-shaped as at the Borden site.
It was shown above that nearly all of the refuse in the Borden landfill is
situated above the water table and that although a water-table mound forms
beneath the landfill, the water table only rises -1 m in response to annual
recharge. The fact that the landfill has produced a large plume of contaminated groundwater indicates that burial of refuse above the water table does
not prevent groundwater contamination. The nearly flat surface of much of
the landfill and the use of clean permeable sand as cover on the refuse layer
and as the final landfill cover has made it possible for considerable infiltration
of rain and snowrnelt into the landfill t o occur. Estimates of the mean annual
rate of recharge t o the aquifer through the landfill were obtained using a
mass balance for bomb tritium in the contaminant plume. These results are
described by Egboka et al. (1983) in Part 3 of this series.
The refuse in the northern part of the landfill was placed in the landfill in
the 1940's and 1950's. The distribution of contamination in the aquifer
immediately below the landfill indicates that this part of the landfill still
produces leachate with electrical conductance as high or higher than is produced in the areas of younger refuse. We expect that leachate with high
electrical conductance will continue to be emitted from the landfill for many
decades or possibly even for hundreds of years. Many years of monitoring of
the groundwater zone beneath the landfill will probably be necessary to
determine whether or not any significant trend toward a change in leachate
concentrations is developing.
SUMMARY OF CONCLUSIONS
The Borden landfill, which occupies -- 5.4 ha, received refuse during 1940
-1976. The landfill has caused the development of a plume of contaminated groundwater that occupies -39 ha and extends more than -700 m
north of the landfill. Electrical conductance and chloride served as convenient
parameters for delineation of the zone of contamination. Sulfate was used to
differentiate the plume in an area affected by road-salt near the front of the
plume. Beneath the landfill, the zone of contamination extends to the bottom
of the aquifer, 20 m below the water table. This deep penetration of the
plume is attributed to small but significant transient downward components
of the hydraulic gradient caused by a slight water-table mound that exists
beneath the landfill during the late spring and early summer. Differences in
density and viscosity caused by differences in TDS and temperature between
the plume and the natural groundwater may also contribute to the downward movement. The plume is much wider in the horizontal plane than
would be the case if it were transported directly northward in a uniform
flow regime. Much of the east- and westward extension of the plume can be
attributed to east- and westward components of flow from the landfill that
occur due to the water-table mounding in the late spring and summer.
Based on the geometric means of the values of hydraulic conductivity
obtained from water-level response tests in piezometers, and on representative
values of hydraulic conductivity from pumping tests at two sites, calculated
positions of the front of the plume based on average linear groundwater
velocities are within about a factor of 2 of the actual measured front. Considering the heterogeneous nature of the aquifer and the potential sources of
error in the hydraulic conductivity values, the factor-of-2 discrepancy is
quite small and is probably better than would be expected in aquifers with
more complex stratigraphy than the Borden aquifer. The predicted frontal
,
I
positions of the plume based on a tracer test and on several borehole dilution
tests are considerably greater than the actual position and greater than the
maximum distances predicted using the Darcy equation.
The thinness of the transition zone between the plume and the zone of
recharge water that overlies it suggests that hydrodynamic dispersion in the
vertical direction is not a major influence on the plume. The thinness of the
plume in the vertical dimension made the use of multilevel samplers and
bundle-piezometersan essential component of the investigation. The decrease
in the degree of irregularity of vertical concentration profiles that occur at
sites at increasing distance northward of the plume is attributed t o vertical
mixing between small-scale heterogeneities.
ACKNOWLEDGEMENTS
Many individuals and organizations provided assistance during the various
phases of this study. The authors thank Paul Johnson, Tim Cosgrave and
Don Miller for assistance with the drilling, installation of monitoring devices,
and water sampling, Boniface Egboka for assistance with sampling of the
multilevel samplers and piezometer bundles, Jack Michels for obtaining
elevations of the piezometers and multilevel samplers, Sam Vales, Gary
Bayes, and Bernd Rehm for drilling the boreholes for installation of most of
the groundwater monitoring devices, Shrikant Limaye for monitoring the
temperature of the aquifer in 1979 and for providing the design of the
temperature monitoring equipment, Heidi Flatt for analysis of the chloride
and sulfate concentrations of the large number of samples collected in 1979,
and John Greenhouse for providing geophysical borehole logs from the site.
The 1974 and 1976 field investigations and the 1978 program of
stratigraphic drilling were conducted by Gartner Lee Associates Ltd. of
Buttonville, Ontario. The investigators made available to us all of the information obtained during these early phases of the study. Financial support
for our investigations was provided by the Ontario Ministry of the Environment, Environment Canada, and the National Science and Engineering
Research Council. The Ontario Ministry of the Environment and Environment Canada also assisted by allowing us to have chemical analyses of
groundwater samples conducted in their laboratories.
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