John A. Cherry

Migration of contaminants in groundwater at a landfill: A case study: 1. Groundwater flow and plume delineation

Abstract A landfill-derived contaminant plume with a maximum width of ∼600 m, a length of ∼700 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 point-samplers 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, which 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 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.

A Field Exercise on Groundwater Flow Using Seepage Meters and Mini-piezometers*

Basic principles of physical hydrogeology and the nature of the hydrologic interactions between groundwater and surface water can be convincingly demonstrated in the field using two inexpensive and...

Using models to simulate the movement of contaminants through groundwater flow systems

Prediction of the movement of contaminants in groundwater systems through the use of models has been given increased emphasis in recent years because of the growing trend toward subsurface disposal of wastes. Prediction is especially critical when nuclear wastes are involved. Contaminant transport models which include the effects of dispersion have been applied to several field situations. However, factors that limit the routine use of these models include the difficulty of determining the field coefficient of dispersion and numerical difficulties encountered when solving the dispersion equation. Regional size models which neglect the effects of dispersion have had limited success because of the scarcity and poor quality of field data. Another difficulty in the development of contaminant transport models is the current lack of knowledge regarding the quantification of chemical reaction terms. This review examines the formulation of contaminant transport models, application to field problems, difficulties ...

Arsenic species as an indicator of redox conditions in groundwater

Abstract Although the thermodynamically based concept of oxidation-reduction potential has for many decades been an accepted tool for interpretation of the chemistry of hydrochemical systems, attempts at measurement of actual redox levels in natural waters have been fraught with difficulty. Existing methods of measurement involve use of potential-sensing inert metal electrodes or analytical determination of redox-indicator species such as dissolved O2 or Fe2+ or redox couples such as SO2−4-HS− and HCO−3-CH4. As a result of recent advances in analytical methods, it is now possible to determine the concentrations of both As(III) and As(V) at sufficiently low levels so that the apparent redox condition, as pE or Eh, can be computed from measured concentrations of As(III) and As(V) species. The arsenic pE or Eh domain obtained using published thermodynamic data for As species and the assumption of redox equilibrium, provides a basis for obtaining an indication of redox levels within the central portion of the redox field for natural waters. The redox domain for the As couple is largest at high total dissolved As concentrations, but even at concentrations as low as 1–10 μg/l the domain has significant extent. Oxidation and reduction of As(III) and As(V) in laboratory trials with redox agents common to natural waters, such as O2, H2S and Fe, suggests that oxidation or reduction of As species in natural waters occurs at rates sufficiently slow to enable water samples to be collected, transported and analysed before excessive change in species distribution takes place, but rapid enough for As species to adjust to the dominant redox condition of the water if periods of years or longer are available for equilibration. Because of the long equilibration time and the position of the pE-pH domain for the As couple, groundwater is best suited for use of As as a redox indicator.

Migration of contaminants in groundwater at a landfill: A case study: 4. A natural-gradient dispersion test

Abstract A natural-gradient tracer test using a chloride solution with an initial injection volume of 0.7 m 3 was performed in the sandy aquifer at the Borden site. The solution was injected into five well points set ∼1 m below the water table in an uncontaminated zone situated above the contaminant plume at a location ∼450 m downflow from the landfill. Under conditions of natural groundwater flow, the tracer slug was monitored for a period of 4 months by withdrawing small-volume samples from points in a three-dimensional array of bundle-type multilevel samplers. Measurements of hydraulic head were obtained from a network of miniature piezometers. Soon after injection, the tracer slug gradually split into two halves, one half moving horizontally at an average velocity of 2.9 · 10 −6 ms −1 and the other horizontally at 8.2 · 10 −7 ms −1 . Although the split has been attributed to local lateral heterogeneity, the nature of the heterogeneity and its influence on the hydraulic-head distribution were not clearly distinguishable in the field data obtained before, during or after the test. The chloride patterns for each of the two halves of the tracer slug evolved into Gaussian forms although the patterns demonstrated some irregularity at early time. The relatively smooth Gaussian forms were unexpected because the aquifer has numerous small-scale heterogeneities observed in vertical cores obtained from the tracer zone and because the glaciofluvial origin of the aquifer suggests that heterogeneities are not laterally continuous. Simulated chloride distributions from a three-dimensional analytical solution to the advection-dispersion equation were fitted to the field data to obtain best-fit estimates of the values of longitudinal, transverse-horizontal and transverse-vertical dispersivity at various travel distances for each of the two halves of the tracer zone. This is the first known field test that has permitted the estimation of three principal dispersion coefficients in layered media. The longitudinal dispersivity was found to increase from 0.01 m at a distance of 0.75 m from the source to 0.08 m at 11.0 m. The transverse-horizontal dispersivity increased also to a value of 0.03 m at 11.0 m. Transverse-vertical dispersion was very weak and was accounted for by molecular diffusion. The relative lack of vertical dispersion is consistent with the shape of the plume of leachate contamination from the landfill. It was concluded that the observed increase in dispersivity along the path of migration is likely caused by heterogeneities. Information on the dispersion-controlling heterogeneities is not yet available as practical field methodologies for their identification and description have not yet been developed. Until such information is incorporated into mass-transport models, a realistic solution of the dispersion problem in heterogeneous media will remain elusive.

Transport of organic contaminants in groundwater

Distribution et devenir des polluants dans des aquiferes de sable et gravier. Processus chimiques, physiques et biologiques dans la zone saturee

The formation and potential importance of cemented layers in inactive sulfide mine tailings

Abstract Investigations of inactive sulfide-rich tailings impoundments at the Heath Steele (New Brunswick) and Waite Amulet (Quebec) minesites have revealed two distinct types of cemented layers or “hardpans.” That at Heath Steele is 10–15 cm thick, occurs 20–30 cm below the depth of active oxidation, is continuous throughout the tailings impoundment, and is characterized by cementation of tailings by gypsum and Fe(II) solid phases, principally melanterite. Hardpan at the Waite Amulet site is only 1–5 cm thick, is laterally discontinuous (10–100 cm), occurs at the depth of active oxidation, and is characterized by cementation of tailings by Fe(III) minerals, principally goethite, lepidocrocite, ferrihydrite, and jarosite. At Heath Steele, an accumulation of gas-phase CO2, of up to 60% of the pore gas, occurs below the hardpan. The calculated diffusivity of the hardpan layer is only about 1 100 that of the overlying, uncemented tailings. The pore-water chemistry at Heath Steele has changed little over a 10-year period, suggesting that the cemented layer restricts the movement of dissolved metals through the tailings and also acts as a zone of metal accumulation. Generation of a cemented layer therefore has significant environmental and economic implications. It is likely that, in sulfide-rich tailings impoundments, the addition of carbonate-rich buffering material during the late stages of tailings deposition would enhance the formation of hardpan layers.

AN IN SITU STUDY OF THE OCCURRENCE AND RATE OF DENITRIFICATION IN A SHALLOW UNCONFINED SAND AQUIFER

Denitrification in shallow groundwater flow systems has been inferred from the observation of declining nitrate concentrations below the water table, and corresponding decline in dissolved oxygen concentrations. To provide direct evidence of denitrification within the saturated zone, and to determine the rate of denitrification, an in-situ injection experiment was conducted using a specially designed injection-withdrawal-sampling drive point. Nitrate and a conservative tracer (bromide) were added to natural groundwater and injected at 3 m depth into a shallow, unconfined sand aquifer. The relative changes in concentration were then observed with time. After 356 h the concentration of nitrate-N in the injected water declined from the initial 13 g m−3 to less than 0.1 g m−3. The decrease in nitrate concentration was much greater than the corresponding decrease in the concentration of bromide, confirming a preferential loss of nitrate. The loss of nitrate was preceded by a decline in dissolved oxygen concentration to less than 0.1 g m−3, and coincided with an increase in bicarbonate concentration of 142 g m−3. The production of bicarbonate observed in the injection experiment, 2.59 mmole HCO−3 per mmole nitrate denitrified, agreed with that calculated using an equilibrium geochemical model of the denitrification process. An increase in the population of denitrifying organisms from 1 to 23 organisms per gram of soil was detected in core samples collected at the depth of injection 169 h after the start of the experiment. The measured rate of denitrification ranged from 0.0078 to 0.13 g m−3 NO−3/1bN h−1, and is in reasonable agreement with published rates for saturated soils. The organic carbon source required for denitrification is either dissolved organic carbon or soil organic carbon. Soil organic carbon, at 0.08–0.16% by weight, is adequate to denitrify large amounts of nitrate.

Origin, age and movement of pore water in argillaceous Quaternary deposits at four sites in southwestern Ontario

Hydrogeologic information in the form of textural properties, field- and laboratory-determined hydraulic conductivity values, vertical hydraulic gradients from piezometer nests, and major-ion and isotopic contents (18O, 2H, 3H, 13C, 14C), was obtained from four sites located in southwestern Ontario on thick deposits of clayey till and glaciolacustrine clay. At each site the piezometers were installed from the water-table zone, situated 1–4 m below the surface, to a maximum depth between 20 and 35 m below the surface. Values of hydraulic conductivity determined in the laboratory by consolidometer and triaxial methods are in the order of 10−8–10−7 cm/s and are similar to values calculated from rates of water-level rise in piezometers. Mean values of average vertical groundwater velocity, calculated with the measured values of hydraulic conductivity, porosity and hydraulic gradient, using the Darcy equation for a saturated non-fractured porous medium, range from 0.13 to 0.26 cm/yr. Tritiated pore waters only occur within 3–6 m of ground surface, thereby indicating that groundwaters below this depth recharged prior to at least 1952. 18O in the pore waters exhibit a distinctive regular shift with depth from water-table values of −9 to −10‰ (SMOW), characteristic of present-day precipitation, to values between −14 and −17‰ at depths of 20–40 m, which are characteristic of much cooler waters. Corrected 14C ages of the groundwater are greater than 8000 yr. B.P., thereby suggesting that the groundwaters at depth entered these deposits during or after the formation of these deposits approximately 11,000 to 14,000 yr. ago. 18O, 2H and Cl− concentration profiles were simulated with a one-dimensional model for transport by advection and diffusion in a saturated porous medium. Reasonable agreement between the model and the field profiles was obtained with values of effective diffusion coefficients of 3.0 · 10−6 cm2/s for 18O, 2H and Cl− and with groundwater velocities between 0.03 to 0.05 cm/yr. Based on these results, it is concluded that the pore water in these deposits is a mixture of late Pleistocene and modern waters and that the distribution of 18C, 2H and Cl− in these deposits is influenced predominantly by molecular diffusion, apparently more so than by hydraulic flow.

Field experiments in a fractured clay till: 2. Solute and colloid transport

A field tracer experiment was conducted in a lateral flow field in the weathered and highly fractured upper 6 m of a 40-m-thick clay-rich till plain in southwestern Ontario. In the upper 3 m where fractures are closely spaced ( 5 m/d. Simulations with a discrete fracture/porous matrix flow and transport model, which used the cubic law for flow in fractures, showed that diffusion of the solutes, but not the much larger colloids, into the matrix pore water between fractures is sufficient to cause the observed difference in solute and colloid transport rates. Transport-derived and hydraulic conductivity-derived fracture aperture values were similar, within a factor of 3 and falling mainly within a range of 5–40 μm. In the upper 3 m the solute tracers were evenly distributed between pore water in the fractures and the matrix, and as a result, solute transport can be closely approximated with an equivalent porous medium (EPM) approach. Below this depth, fractures are more widely spaced (0.13 to >1 m) with concentration peaks tending to occur near visible fractures, and solute transport cannot be adequately described with an EPM approach.