Abstract

At landfill disposal sites the unsaturated zone forms an important buffer between hazardous wastes and the water table. Understanding the mechanisms and rates of movement of pollutants in sedimentary rocks is an important step in the process of groundwater protection. To this end contaminant migration has been studied for more than six years in experimental in situ lysimeters in the Lower Greensand, at Uffington, Oxfordshire, U.K. Migration of organic compounds such as chlorinated phenols and carboxylic acids, and of some of the inorganic anions was found to be dominated by microbiological reactions, but the migration of the major cations and heavy metals is controlled by a far more complex interplay of geochemical and microbiological processes. These processes have been studied by analysis of sediment samples and fluid breakthrough data. A simple conceptual model has been developed to explain the retention pattern observed in the Greensand, and this suggests that pH buffering reactions are of primary importance, for heavy metals, rather than cation exchange phenomena, and have a significant effect on the migration of other contaminants.

In 1974, the U.K. Department of the Environment initiated a research programme into the disposal of hazardous wastes in landfill sites, which is still continuing. The British Geological Survey, as a major contractor to this programme, has undertaken detailed studies of existing landfill .sites representing a range of waste types and hydrogeological regimes (DOE, 1978). Extensive experimental lysimeter studies have also been conducted to identify the processes which control contaminant attenuation in the unsaturated zone.

Three formations representing the main types of aquifer in the U.K., were selected for the lysimeter studies; Lower Chalk, Lower Greensand and Plateau Gravels. The Sherwood Sandstone was also the subject of limited investigations (Stuart, 1977). The experiments in Chalk (Black and Kipp, 1983) and the Plateau Gravels (Stuart, 1976, DOE, 1978) were designed principally to evaluate physical processes and were relatively small in scale. Amore detailed study of contaminant interactions in the unsaturated zone was undertaken in the Lower Greensand, using large in situ lysimeters.

Although the lysimeters have now been decommissioned, work is still continuing on the characterisation of sediment samples collected at the end of these experiments. Detailed results from these lysimeters are presented elsewhere, (Campbell et al, 1981a, b; Parker et al, 1981; Rees et al, 1981; Ross et al, 1981) and only summarised briefly here. The main aim of this paper is to describe a conceptual model for attenuation of heavy metals,, which has been developed as the lysimeter studies have progressed. 

Fig. 1 Lysimeter cross-section showing details of instrumentation.

Lysimeter construction and operation

At the experimental site, near Uffington in Oxfordshire, the Lower Greensand forms an outlier 3-3.5m thick, unconformably overlying Kimmeridge Clay. At this locality, the Greensand comprises a well bedded sequence of calcareous, poorly cemented fine sands and sandy clays. Four undisturbed blocks of the Lower Greensand, each 4.5m square and approximately 3m deep were isolated within reinforced concrete walls penetrating to the Kimmeridge Clay.

Two pairs of lysimeters were built with a central access trench between them, which allowed instrumentation to be introduced horizontally. The principal instrumentation is summarized in the schematic cross-section shown in Fig. 1.

The lysimeters were irrigated with synthetic leachate solutions, containing high concerlltrations of a wide range of contaminants, at a rate lof 730 mm annum giving average pore fluid velocities of 6mm day Leachate migration was monitored by analysis of pore fluid and gas samples collected at various depths in the lysimeters.

Leachate Composition.
Four major groups of contaminant were studied:

a) aromatic organic compounds including aniline, phenol and chlorophenols;
b) aliphatic carboxylic acids;
c) major anions including chloride, sulphate and nitrate, together with ammonia;
d) heavy metals.

Both the "aromatic organic" arid the "anionic" leachate supported intense fungal and bacterial growth. Attenuation was dominated by microbiological reactions and only chloride and sulphate proved persistent. Results from these lysimeters are discussed in more detail by Parker et al (1981), Campbell et al (1981a and b) and by Rees et al (1981). Leachate/rock interactions proved more important in the attenuation of the synthetic landfill leachate containing elevated concentrations of heavy metals and carboxylic acids.

The composition of the synthetic heavy metal solution (Table 1) was selected to simulate leachate resulting from the codisposal of domestic and industrial wastes, such as metal finishing sludges, but with the concentrations of individual heavy metals, except mercury, at 100mg1 (about 10 times the concentrations usually observed).The experiment was conducted in two parts and after 1140 days irrigation with the heavy metal leachate, irrigation was continued with a "blank" solution excluding heavy metals.

Solution adjusted to pH5 using ammonia
Table 1 Composition of synthetic heavy metal leachate

Results

Carboxylic acid attenuation

As with the aromatic organic and anionic leachates the attenuation of carboxylic acids in the mixed heavy metal solution proved dependent on microbial degradation. During the first year of the experiment virtually no attenuation of the carboxylic acids was observed (Fig .2) as a result of the inhibition of microbial colonization at the lysimeter surface by high concentrations (1000ug q^-1) of heavy metals. Subsequent development of an adapted population of microorganisms beneath the zone of heavy metal inhibition resulted in a marked reduction in the concentrations of carboxylic acids eluted and anaerobic conditions developed rapidly with the production of carbon dioxide and methane. Degradation rates were found to increase with substrate concentration and temperature and an optimum rate of 100mgl^-1 day^-1 was obtained for all three carboxylic acids (normalized to a 5000mgl input of each acid). The rate dependence on temperature led to an increase in the concentrations of carboxylic acids eluted during the colder winter periods, and a cyclic release pattern (Fig.2) was established following the winter/summer temperature variations.

Simple adsorption - cation exchange model.
The first model adopted for the attenuation of the heavy metals was a simple adsorption model. Studies such as those by Griffin et al (1976, 1977) and Fuller (1978) had shown that the attenuation of landfill leachate contaminants, particularly heavy metals, could be correlated directly with the cation exchange capacity and the surface area of the sediment. It was envisaged that the Greensand would have a finite "adsorption capacity" which once saturated would allow the contaminant to pass throught at input concentration.

This model is illustrated in Fig. 4; no allowance has been made for dispersion. In Fig. 4a. the concentration of contaminant has reached saturation in the upper layer of sediment and a fluid sample taken at the base of this layer would show the contaminant to have reached input concentration. After twice the initial period of irrigation (Fig. 4b) the sediment would be saturated to twice the depth of the first layer, and fluid samples taken at the lower depth would take twice as long to reach input concentration. This pattern of breakthrough would be repeated with increasing depth.

It was assumed that the 'adsorption capacity' of the Greensand would approximate to the measured cation exchange capacity of 50 meq kg On this basis input concentrations of the heavy metals would have been expected at the first fluid sampling point, 400mm below the lysimeter surface, after 150 days. As Fig. 3 shows, not even the most mobile metals, nickel and cadmium, were detected within this time. Therefore the simple cation exchange model had to be refined.

Refinement of cation-exchange model - analysis of sediment core samples.
The limited migration of the heavy metals made it difficult to deduce the processes controlling their migration from the fluid data alone. Therefore as the experiment proceeded a number of core samples were taken from the lysimeter. Preliminary analysis of the sediment (Campbell et al, 1981a, Ross, 1980) showed that the total amounts of heavy metals retained, up to 75 meq kg were much higher than the can on exchange capacity of the Greenland. However the order of retention of the metals (Pb iCu - Cr >Zn >Cd > Ni) was found to be the inverse of the order of mobility established in small-scale column experiments (Boreham et al, 1976).

More detailed analysis of core samples using sequential, selective chemical extraction procedures (Ross, 1978, 1980, Ross et al, 1981) showed that the metals were retained by a combination of adsorption and precipitation processes. The order of retention could be related to the stability of the phases responsible for attenuating the metals.

Thus lead and copper, which are strongly retained, are associated with recalcitrant sulphide and hydroxide phases (Solubility constants, heavy metal sulphides 10-²⁴ - 10-⁴² Mol ²kg^-2; hydrxides 10-¹⁰ - 10-²⁶ Mol ³ kg-³). More mobile elements, such as nickel and cadmium, are associated with labile exchangeable and carbonate fractions (Solubility constants heavy metal carbonates 10^-10 - 10^-14 Mol Kg-2). However the distribution of metals among these phases does not remain constant with continued irrigation. Figs.5. qand 6 show the distribution of cadmium and lead in two cores taken at different times during the period of heavy metal leachate irrigation. The total concentrations of each metal with depth are shown in the line graphs to the right of Figs. 5 and 6 while the bar charts show the percentage of this total concentration attributable to various phases discriminated by sequential extraction. In the early stages of irrigation (Figs. 5a and 6a) high concentrations of both cadmium and lead are found in the upper 100mm of each core sample. In the second core sample (Figs5b and 6b) taken towards the end of the 1140 day period of heavy metal irrigation, the peak concentration of cadmium has fallen to about 250tig g 1 and the concentration profile is nearly constant with depth, suggesting near equilibrium conditions. In contrast the concentration profiles for lead (Fig 6b) have remained almost unchanged, apart from a slight increase in concentration at greater depth.

The difference in the concentration profiles is explained by the distinct difference in the distributon of the two metals between various phases in the sediment. Cadmium is retained largely by the carbonate and exchangeable fractions. Initially, while the leachate pH is buffered through the dissolution of calcite/aragonite in the sediment, cadmium carbonate is precipitated and high concentrations of cadmium are retained at the surface of the lysimeter (Fig 5a).

As the sediment loses this pH buffering capacity, the pH of the pore fluid will fall to the pH of the leachate (pH5.0) and the cadmium carbonate will start to dissolve. This leads to a redistribution of the cadmium retained by the sediment and a reduction in the total amounts retained at the surface of the lysimeter. Lead is associated with the less soluble hydroxide and sulphide fractions in the sediment and is therefore less vulnerable to remobilization when the pH buffering effect of the calcite disappears. There is therefore little change in the concentration or distribution of lead between the two core samples (Figs. 6a, b).

Fig. 5. Distribution of cadmium in the Lower Greensand: (a) after approximately 700 and (b)
1000 days irrigation with heavy metal leachate.

Fig. 6. Distribution of lead in the Lower Greensand: (a) after approximately 700 days and (b)
1000 days irrigation with heavy metal leachate.

Fig 7. Calcite/aragonite pH buffer control model; heavy metal migration with continuous input of leachate. Equilibrium (=/m) and preliminary saturation concentrations will vary for individual metal.

pH buffering - Equilibrium distribution model
he changing distribution of the heavy metals with continuing irrigation gives rise to a modified model for heavy metal migration which is summarised in Fig. 7. This shows the change in sediment concentration/depth profile which might be expected at equal time increments during continuous irrigation with the heavy metal leachate. Fig. 7a shows the initial condition with the saturation of the first layer of sediment; the saturation equilibrium concentration represents the summation of all fractions contributing to attenuation at pH7.0-7.5. Fig 7b shows the profile developed after a further (equal) period of irrigation; the pH buffering capacity of the first layer has been exhausted and a new 'lumped' equilibrium is established at the lower input pH of 5.0. The new equilibrium position again refers to all processes contributing to attenuation, but will be shifted further from the saturation equilibrium position for those elements retained principally as carbonates, than for elements associated with hydroxide or sulphide fractions. At greater depth in the sediment the pH is still buffered by calcite, and the high preliminary saturation concentration is still observed; the metal concentration front is also displaced to greater depth, as a result of leaching of the metal from the surface layers where the pH is no longer calcite buffered.

The pH buffered 'saturation front' gradually moves through the sediment as shown in Figs. 7c - e, which represent the sediment concentration profiles after three further equal periods of irrigation.

The model can easily be extended to take account of the effect of removing the heavy metal input as shown in Fig. 8. Removing the heavy metal input does not have the effect of immediately remobilizing all the metals absorbed by the sediment. Instead the upper layers of sediment which have reached equilibrium with the acid metal leachate will act as a source of metals for the (chemically) undersaturated layers beneath. Fig. 8a represents the first stage of leaching of the profile developed in Fig. 7e. Layer 1 of the sediment reaches a new equilibrium with the 'blank' leachate producing essentially an acid heavy metal leachate. Layers 2,3 arid 4 were previously at equilibrium with the acid heavy metal leachate and the 'new heavy metal leachate' will pass through these layers with little effect.

In layer 5 calcite will start to dissolve and the heavy metals in this layer of sediment will begin to re-equilibrate to equilibrium position A, for the acid heavy metal leachate. Below layer 5 the pH is still fully buffered, and these layers of sediment will act as a sink for the metals displaced from the surface layers. After a second equal period of irrigation (Fig. 8b) the concentration of heavy metals in layer 2 has reached the new equilibrium position and further redistribution of the metals occurs. This redistribution process is seen clearly in Fig. 9 which shows the concentrations and relative speciation of cadmium and lead at the end of the experiment, after 500 days irrigation with blank leachate. No calcium was found in the carbonate fraction in

Fig.8 pH-buffering model for heavy metal redistribution
during irrigation with blank leachate

Implications for further model development

The model of pH dependent equilibriation and, essentially chromatographic, redistribution of heavy metals obviously oversimplifies both the chemical reactions and the physical processes involved in attenuation. Although one or two fractions usually dominate the retention of any element, the total amount retained reflects the summation of a complex interplay of processes.

Possible reaction pathways are summarised in Fig. 10 and discussed in more detail by Ross (1978, 1980) and by Ross et al. (1981).

Fig. 10. Possible reaction pathways of heavy metals in the Lower Greensand 

However the pH dependent 'lumped equilibrium' model can be used as the basis of a simple mathematical model. A preliminary attempt to do this, using measured concentrations from core samples to estimate 'saturation equilibrium', 'acid leachate equilibrium' and 'blank leachate equilibrium' concentrations for nickel, sLiggested that at a more realistic input concentration of l0mg 1 , with secondary 'blank' leachate irrigation after 1140 days, it would take nearly 11 years for all of the nickel to migrate through t? upper 400mm of sediment, with a peak concentration of 7mgl being observed after 2600 days (Ross et al, 1981).

This type of model relies on the measurement of metal concentrations in the sediment after irrigation with leachate, and is specific to the particular lithology under study, as well as the leachate composition. Results from other formations (Stuart, 1977; Newman and Spears, 1984) particularly the Sherwood Sandstone, which showed much less heavy metal uptake than the Greensand, support the critical importance of pH buffering in heavy metal retention, but do not enable reliable predictive modelling of contaminant migration in lithologies which have not been studied in this way. Similarly it is impossible to predict the impact of leachates of differing composition, although the synthetic leachate used was intended to reproduce a 'worst case' situation.

Despite the over-simplification of the mathematical modelling and the limited range of the experimental work, it is clear that there is no real risk to aquifers, even the Sherwood Sandstone, by controlled disposal of heavy metals. This has made the development of more rigorous mathematical models, coupling fluid transport with thermodynamic controls on solute speciation, unnecessary. However, the application of such models, which are currently being developed within the more sensitive area of radioactive waste disposal research (see e.g. Isherwood and Wolery, 1984), would allow the lysimeter data to be exploited fully. Only a coupled solute transport/thermodynamic model can be used ultimately as a reliable basis for the extension of these results to the evaluation of a potential landfill site. Conversely the data from these experiments offer an excellent means of validating such models.

Chemical (thermodynamic) modelling can be applied equally well to the other major ions common in leachates from both domestic and industrial landfills, since there is no evidence of a kinetic control on the migration of these elements (Ross 1978, 1980).However microbiologically catalysed reactions, such as nitrate and sulphate reduction and the breakdown of carboxylic acids, will be kinetically constrained and have considerable impact upon the speciation of labile elements (ie. S, 0, C, N, H, P. Fe? Mn?). These reactions will, in turn, control processes such as sulphide precipitation (c.f. Fig. 10) once the concentrations of inhibitory substances (in this case, heavy metals) have been reduced sufficiently to allow the development of an appropriate microbial population.

The lysimeter experiments have shown the potential for microbial catalysis of carboxylic acid degradation, but it is unclear if those results can be applied more generally. The same is true of redoxcontrolled speciation, which does appear kinetically limited.

There is a need for much more work on microbial catalysis, and field data need to be obtained to determine the range of reaction rates and the environmental factors which control them.

Conclusions

Extensive experimental lysimeter studies have been conducted in the Lower Greensand, on the unsaturated zone migration of contaminants common in landfill leachates. These studies, in conjunction with limited investigation of other lithologies, have shown considerable potential for attenuation, particularly for heavy metals. It is suggested that, for acid leachates, the primary control on the migration of major canons and heavy metals is pH buffering which controls their solubility.

A simple conceptual model is presented to explain the impact of pH buffering on 'lumped equilibrium' sediment concentrations, which is supported by heavy metal distribution profiles taken at various stages in the experiments. The relative mobility of the metals can be related to their pattern of attenuation in various fractions of the sediment.

Mobile elements such as nickel and cadmium which are retained by cation exchange and carbonate precipitation, are remobilised to a greater extent, when the pH buffering capacity of the sediment is exhausted. In contrast less mobile elements such as lead and copper which are attenuated as sparingly soluble sulphides and hydroxides, are much less affected by a reduction in pH. The more soluble carbonates, calcite and aragonite, are sacrificed to pH buffering, and hence more calcium and magnesium are eluted than added. The alkali metals are controlled by cation exchange, and compete effectively with heavy metals for exchange sites.

The experiments have also demonstrated the potential for microbial breakdown of carboxylic acids in the unsaturated zone, but much more field data need to be obtained before the effects of microbial catalysis can be predicted confidently.

Acknowledgements

This research project is part of a larger programme sponsored by the Department of the Environment. The author would like to extend sincere thanks to the many members of staff of both the Harwell Laboratory and the British Geological Survey, as well as Miss R Newman and Dr D A Spears of Sheffield University, who contributed to this work.

This paper is published by permission of the Director of the British Geological Survey (Natural Environment Research Council).

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