Migration of contaminants in groundwater at a landfill: A case study

Journal of Hydrology, 63 (1983) 1--29

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Elsevier Science Publishers B.V., Amsterdam -- Printed in The Netherlands

MIGRATION OF CONTAMINANTS IN GROUNDWATER LANDFILL: A CASE STUDY 1. G r o u n d w a t e r F l o w a n d P l u m e D e l i n e a t i o n

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D.S. MacFARLANE, J.A. CHERRY, R.W. GILLHAM and E.A. SUDICKY

Department of 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 at 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 ~ 6 0 0 m, a length of ~ 7 0 0 m 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 temperature of groundwater in the zone of contamination beneath the landfill is 12 C, whmh is 4--5 C above background. The thermal 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 leachate 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 o f 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 diRpersion. The plume passes beneath a small shallow stream near the landfill without significant influence on the stream. .

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INTRODUCTION The most c o m m o n 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 b y 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 n o t 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 c o m m o n in the glaciated region of North America. The abandoned landfill is located within the confines of the Canadian Forces Base (C.F.B.), Borden, ~ 8 0 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 ai., 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

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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 2 3 2 m above sea level (a.s.1.) at the landfill to a low of 2 2 2 . 5 m 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 s o u t h , east and west sides of the m o u n d 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 northwestem 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 4 0 m 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 t o w n of Angus until 1970 (Gartner Lee Associates Ltd., 1977). The mean daily temperature ranges from 12.4 ° to 0.6°C, averaging 6.5°C, with extremes of + 3 6 . 7 ° and --41.1°C. Mean total precipitation averages 82.8 cm yr_ 1 , of which 58.67 cm was in the form of rain, and 240.5 cm was snow. A typical year has 171 days of frost, 88 days of rain and 56 days of snow.

HISTORY OF 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 t h a t 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 to 10 m, with the thickest refuse in the north and northwest sections. The most rapid input of refuse likely occurred during the period of World War II when the Borden c o m m u n i t y

grew rapidly to a peak population. In 1973, modern landfilling practices were initiated and were continued until the operation was terminated in 1976. The final landfill surface was covered with "~0.5m of sand. From boreholes and test pits, Gartner Lee Associates Ltd. (1977) established that ~ 8 0 % of the refuse deposited during the life of the landfill operation consisted of ash, w o o d and construction debris. The remaining refuse is composed mainly of domestic and commercial food wastes. Hydrogeological studies at the Borden landfill began in 1974 when Gartner Lee Associates Ltd. installed 51 piezometers and water-table standpipes at 26 locations. The purpose of this monitoring was to determine, at a reconnaissance scale, the influence of the landfill on the quality of groundwater in the unconfined sandy aquifer. In this phase of the study, it was established that although the quality of water in the unconfined aquifer was degraded, the contaminated zone was separated b y 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 n o t predictable at that time and it was decided to close the landfill in favour of a new site on much less permeable deposits elsewhere on the base. A long-term program of research at the abandoned landfill was initiated in 1976 by Gartner Lee Associates 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 b y measurements of electrical conductance on groundwater samples obtained using an auger-head sampler designed specifically for the task (Anderson, 1977; Cherry et al., 1983). An additional 69 piezometers and water-table standpipes were installed during this investigative 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-piezometers were installed at the Borden site (Fig. 2). Twenty-nine multilevel groundwater samplers of the type described by Pickens et al. (1978) were also installed in 1978/1979 for sampling of groundwater at numerous depths at each monitoring site. More bundle-piezometers were installed in 1980/ 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 installed using 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 enabled 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

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depth, it was raised a few decimetres and the piezometer or multilevel sampler was lowered through the auger stem and then used to knock o u t the rubber stopper. As the augers were removed, the cohesionless sand collapsed tightly around the pipe. Water was n o t 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 at a few of the drill sites. At depths exceeding 20 m, the rubber stopper did n o t keep formation sand o u t of the auger, and a column of water was maintained in the annular space to 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 to prevent excessive infilling of silt or clay. The piezometer screens are 0.3--0.6 m long and the standpipes are screened from the b o t t o m of the pipe to above the water table. In 1979/1980, fourteen multilevel bundle-piezometers, with a total of 124 sampling points labelled T-series, were installed to various depths. Each bundle is comprised of eight 1.0-cm (~-in.) I.D. polyethylene tubes bound to a central rigid 1.3-cm (½-in.) I.D. PVC pipe. A 15-cm (6-in.)longperforated plastic screen wrapped with nylon mesh was attached to the end of each

tube. These piezometers were installed through the hollow stem of the flight augers in the manner described above. Where water was used during installation, representative chemical samples could n o t 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 b y 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, ~ 6 0 0 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 in April 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 cm s were p u m p e d 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 to 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 m a n y 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 b y 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 glaciofiuvial origin, the unconfined aquifer is locally very heterogeneous d u e 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 b o t t o m 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 ~'9.5 m in the northern end of the area (Fig. 3). Beneath the landfill, the surface of the clayey silt increases in elevation northward. In the b o t t o m 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 . 5 m in thickness and in some cases were less than 0.1 m thick. Because they are so thin, they may n o t have been detected in some of the deep boreholes. It is therefore n o t known if these beds are continuous in this area. The sand in the b o t t o m part of the aquifer appears to be more silty than elsewhere in the aquifer but this change is n o t very 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 t o p 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 m e t h o d utilizes the effective grain size (dl0)

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whereas the latter t w o methods include a measure of the actual distribution in grain size. Values of hydraulic conductivity of two samples from a borehole between $8 and S9 were subjected to permeameter tests. One sample, which was composed of clean medium-grained sand yielded a value of 10 -2 cm s -1 . A sample of silty fine-grained sand gave a value of 10 -4 cm s -1 . 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" 10 -a - - 5 " 10 -3 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 c o m p u t e d 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 from the initial water table to the b o t t o m of the well screen. At the site located at the southeastern edge of the landfill where the aquifer is thick, the large degree of partial penetration of the pumping well m a y 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 c m -a 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 to ground surface in the abandoned sand pit and in the

12 area south of the landfill, where it is generally less than 1 m 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 b o t t o m of the refuse indicate that during this period the water table in part of the landfill area exists within the b o t t o m metre of refuse. The water table gradually declines b y 0.75--1 m during the summer and early fall and then rises b y 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 dates in 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 east-west direction during the fall and winter months, as illustrated by the watertable map for October 17 (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 to fall reflect the formation of a water-table m o u n d 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 p o n d resulted mainly from surface runoff from the landfill mound. The isopotential diagrams for flow along the norih--south 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 m o u n d is situated beneath the landfill rather than beneath the periphery of

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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 exhibits closed isopotential lines in the mid-depth of the aquifer beneath the landfill. This reflects a c o m p o n e n t of groundwater flow that is not in the plane of the cross-section. This component of flow appears to be generated b y higher water-table levels beneath the western part of the landfill, which, for a month or two, causes eastward flow

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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, ELECTRICAL CONDUCTANCE

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longitudinal cross-section; the sampling points for chloride and electrical conductance are

17

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 m a x i m u m 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 10 mg 1-1 , whereas groundwater that is strongly contaminated by leachate has more than 500 mg1-1 . The results o f the chloride analyses are shown in plan view in Fig. 6b and along cross-section A--A' in Fig. 7c. The maps of chloride (Fig. 6b) and electrical conductance (Fig. 6a) were constructed using the m a x i m u m 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 m a x i m u m consistently occurs at or near an elevation of 2 1 4 m a.s.1. Fig. 8 shows depth profiles of CHLORIDE 200

O ~

~

4

~')

6

400

1

600

i

200

0

CONC E N T R A T I O N

400 [

600

o

IOO

(rag/L) 200

300

i

I

0,

2

2t

4

4,~

I00

200 r

500

6I

/ S~PLE POINT

8~

8

I

12

12

'°1

Q

14 ¢n

16

a.

150rn

,

®

I

400m FROM LANDFILL 300rn FROM LANDFILL

FROM LANDFILL

20 L

75rn

i

i

i

FROM LANDFLL

Fig. 8. Vertical profiles o f chloride ( m g l - l ) at sites along the longitudinal cross-section, from near the landfill on the left to greater distance from the landfill on the right.

18 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 SOl- to H2 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 n o t appear to be a significant mechanism of SO4:- 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 b o t t o m 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 2060mg1-1 under the landfill to background levels, which are generally less than 25 mg 1-1 . Oxygen-isotope analyses of sulfate suggest that the sulfate is likely derived from construction materials in the landfill (S. Feenstra pets. 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 ~ 7 0 0 m north along the principal direction of flow, and as much as 200 m laterally to 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 b o t t o m 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 Cl-, SO~- and electrical conductance obtained during 1977/1978 were 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

19

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 the aquifer, where the plume occurs, they are high.

TEMPERATURE DISTRIBUTION

Biochemical processes within the landfill cause the landfill to be much warmer than the surrounding soft. In longitudinal cross-section (Fig. 9), a zone enclosed by the l l ° C contour, 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

l,,,

,

~

o.

,,..

-

,...

,

~

.

210

200

..~-~--'---~--~-~-~"

I00 "

------~-50 0

TEMPERATURE

°C

lOOm

VERT. EXAG. = I0

Fig. 9. Temperature distribution along the longitudinal cross-section (contour intervals 0.5°C).

20 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 land filling 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 silty~layey 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.,

21 1977) does n o t differ appreciably from the general form of the maps presented in this paper based on data obtained in 1 9 7 9 / 1 9 8 0 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 b y 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 . 3 5 , expressed as a fraction. From data on hydraulic gradient (i), effective porosity (n) and hydraulic conductivity (K), values of the average linear groundwater velocity (~) in the northward direction were obtained from the relation: =

-

Ki/n

(1)

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 - 1 0 -2 cm s -1 . For this t y p e of sand the geometric mean values of the hydraulic conductivity obtained from piezometer response tests are in the range of 3" 1 0 - 3 - - 5 . 1 0 - 3 c m s - 1 . 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 17-28 m yr_ 1 . Based on the hydraulic conductivity values from the permeameter test and the pumping tests, the range of ~ is 8--14 m yr.-1 . From 1940 to 1979 (approximately 38 yr. of travel time), the minimum estimated northward movement of the leachate front (~ = 8 m) would be 300 m and the maximum (~7 = 28 m) would be 1060 m. The observed position of the northward front of the plume is ~ 7 0 0 m 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 c o m p u t e d 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 b y Sudicky et al. (1983) in Part 4 of this series and by boreholedilution tests of the t y p e 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

22 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 y r . - ' . The borehole~lilution tests provided velocity values in the range of ~ 3 0 - - 9 0 m yr_ ~ . The velocity values from the tracer test and the borehole
23 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-migrating plumes 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 north--south 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 mixing 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 to 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 ~ 2 0 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

24 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 to cause a vertical driving force in addition to that caused b y 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 t w o 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 to rationalize the downward movement of warm leachate into colder ambient groundwater in an aquifer in which the flow is predominantly horizontal. With TDS of ~ 2000mg1-1 and a temperature of l l ° C , contaminated water beneath the landfill has a mass density of 1 . 0 0 1 2 g c m -3. Uncontaminated groundwater with 100mg1-1 TDS and a temperature of 7°C has a mass density close to 1.000g cm- 3. The net density driving force can be obtained from the relation AP/Po, where Ap is the density difference and P0 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 to 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 c o m p o n e n t of the hydraulic gradient and is n o t much smaller than the average horizontal c o m p o n e n t 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 to be correspondingly higher. In the Borden aquifer, this effect is small because a temperature difference of 4°C 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.

25 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 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 b y Kimmel and Braids (1980) is the presence at the Borden site of the transient water-table m o u n d and transient components of flow transverse to the general northward direction of flow. At the sites on Long Island studied b y 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 m o u n d 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

26 of rain and snowmelt into the landfill to occur. Estimates of the mean annual rate of recharge to the aquifer through the landfill were obtained using a mass balance for b o m b tritium in the contaminant plume. These results are described b y 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 ~ 3 9 h a and extends more than ~ 7 0 0 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 b y road-salt near the front of the plume. Beneath the landfill, the zone of contamination extends to the b o t t o m of the aquifer, 20 m below the water table. This deep penetration of the plume is attributed to small b u t significant transient downward components of the hydraulic gradient caused by a slight water-table m o u n d 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 a b o u t 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

27

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-piezometers an 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 to 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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28 Bear, J., 1972. Dynamics of Fluids in Porous Media. American Elsevier,N e w York, N.Y., 764 pp. Burwasser, G.J. and Cairns, B.D., 1974. Quaternary geology of the Barrie area (western half), southern Ontario. Ont. Div. Mines, Prelim. M a p p. 978, Geol. Ser.,scale 1:50,000. Chapman, L.J. and Putnam, D.F., 1951. The Physiography of Southern Ontario. University of Toronto Press, Toronto, Ont., 284 pp. Cherry, J.A., Gillham, R.W., Anderson, E.G. and Johnson, P.E., 1983. Migration of contaminants in groundwater at a landfill: a case study, 2. Groundwater monitoring devices. In: J.A. Cherry (Guest-Editor), Migration of Contaminants in Groundwater at a Landfill: A Case Study. J. Hydrol., 63:31--49 (this Special Issue; in this paper referred to as Part 2). Dance, J.T. and Reardon, E.J., 1983. Migration of contaminants in groundwater at a landfill: a case study, 5. Cation migration in the dispersion test. In: J.A. Cherry (Guest-Editor), Migration of Contaminants in Groundwater at a Landfill: A Case Study. J. Hydrol., 63:109--130 (thisSpecial Issue; in this paper referred to as Part 5). Dickin, R.C., 1979. Landfill leachate contaminant patterns and soil water geochemistry at C F B Borden, Ontario. M.Sc. Project, Department of Earth Sciences, University of Waterloo, Waterloo, Ont. Egboka, B.C.E., Cherry, J.A., Farvolden, R.N. and Frind, E.O., 1983. Migration of contaminants in groundwater at a landfill: a case study, 3. Tritium as an indicator of dispersion and recharge. In: J.A. Cherry (Guest-Editor), Migration of Contaminants in Groundwater at a Landfill. A Case Study. J. Hydrol., 63:51--80 (this Special Issue; in this paper referred to as Part 3). Fair, G.M. and Hatch, L.P., 1933. Fundamental factors governing the streamline flow of water through sand. J. A m . Water Works Assoc., 25: 1551--1565. Freeze, R.A. and Cherry, J.A., 1979. Groundwater. Prentice-Hall Inc., Englewood Cliffs, N.J., 604 pp. Gartner Lee Associates Ltd., 1975. Groundwater monitor installationsfor existing and proposed sanitary landfill sites,C F B Borden. Prepared for Environ. Prot. Serv., Ont. Reg., Environ. Can., Toronto, Ont., Gartner Lee Associates, Toronto, Ont., March 18, 1975. Proj. No. 74--80, 29 pp., 1 plate. Gartner Lee Associates Ltd., 1977. Leachate migration study, C F B Borden. Fin. Rep. prepared for Environ. Can., Toronto, Ont., Gartner Lee Associates, Toronto, Ont., M a y 1977, 24 pp. and appendices. Gartner Lee Associates Ltd., 1978. Stratigraphic sampling at C F B Borden for the University of Waterloo, Waterloo, Ont., Gartner Lee Associates, Toronto, Ont., July 28, 1978. Germain, M.S.D., 1981. Quasi-stable concentration distributions in saturated porous media with a constant solute source. M.Sc. Thesis, Department of Earth Sciences, University of Waterloo, Waterloo, Ont. Gillham, R.W., Sudicky, E.A., Cherry, J.A. and Frind, E.O., 1983. An advection--diffusion concept for solute transport in heterogeneous unconsolidated geological deposits. Water Resour. Res. (in press). Golwer, A., Matthess, G. and Schneider, W., 1975. Effects of waste deposits on groundwater quality. Proc. of the M o s c o w Symposium, Groundwater Pollution, I.A.H.S.-A.I.S.H. Publ. No. 103, pp. 159--166. Greenhouse, J.P. and Harris, R.D., 1983. Migration of contaminants in groundwater at a landfill: a case study, 7. DC, V L F and inductive resistivitysurveys. In: J.A. Cherry (Guest-Editor), Migration of Contaminants in Groundwater at a Landfill: A Case Study. J. Hydrol., 63:177--197 (thisSpecial Issue;in this paper referred to as Part 7). Kimmel, G.E. and Braids, O.C., 1975. Leachate plumes in a highly permeable aquifer. Ground Water, 12: 388--392. Kimmel, G.E. and Braids, O.C., 1980. Leachate plumes in groundwater from Babylon and IslipLandfills,Long Island, N e w York. U.S. Geol. Surv., Prof. Pap. 1085, 38 pp. Lambe, T.W. and Whitman, R.V., 1969. Soil Mechanics. Wiley, N e w York, N.Y., 553 pp.

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