Hydrogeology in the Service of Man, Mémoires of the 18th Congress of the International Association of Hydrogeologists, Cambridge, 1985.

ASSESSING GROUNDWATER POLLUTION FROM LANDFILL SITES: RESULTS OF CASE STUDIES

G.M. WILLIAMS. Fluid Processes Research Group, British Geological Survey, Xeyworth, Nottingham.

ABSTRACT Predicting the extent of groundwater pollution from a proposed

landfill site depends on a thorough knowledge of the character­istics of groundwater flow, and of the chemical and biochemical interactions that occur between waste components, groundwater and aquifer minerals. Although certain parameters, eg permeability, porosity, hydraulic head, may be determined in the vicinity of a proposed landfill, the measurement of dispersivity over a large enough scale to be representative cannot be made within a short time scale. Similarly, reactions.especially biodégradation, which depend on the establishment of a bacterial population within the aquifer may be difficult to determine over short time scales, and prediction of contamination migration must rely heavily on quanti­fication of these processes at existing landfill sites in analogous hydrogeological environments where pollution plumes are already established.

Studies undertaken within the U.K. at long established hazardous waste landfills, eg Villa Farm, Hooton, will be used to illustrate the dominant controls on leachate attenuation in intergranular aquifers and the approach used to obtain quantitative data for pre­dictive use.

INTRODUCTION Predicting the migration of contaminants from a landfill site

requires an understanding of the physical factors which control the flow of water through the aquifer and the chemical and biochemical reactions that take place between waste components, groundwater and aquifer minerals. Once the physical properties of the aquifer and the chemical and biochemical reactions have been quantified they must be integrated into a solute transport model in order to pre­dict contaminant concentration distributions at various times for a range of possible operating conditions, waste types and input or leaching rates.

Obtaining information to describe the physical flow of water usually involves a field investigation of the site geology and measurement of the hydraulic properties of the principal formations or features which control groundwater flow. In its simplest form this involves the determination of hydraulic conductivity and porosity which, coupled with hydraulic head information, allow the velocity and direction of water flow in the saturated zone away from the landfill to be determined. When attempting to construct a model to describe contaminant migration rates, or their concen­tration distribution with time, a great deal more information is required and models become far more complex. For a conservative or non-reactive solute, data are required on the longitudinal and

- 106 -

transverse dispersivity of the aquifer, and on the extent of solute interaction between mobile and immobile pore water (matrix dif­fusion) which is particularly important in dual porosity and un­saturated media. For reactive solutes all geochemical and bio­chemical processes that affect their mobility and transformation of the compound in the aquifer need to be identified and quantified (e.g. sorption, precipitation, complexation, oxidation/ reduction biodégradation, etc).

Obtaining these extra data presents difficulties with regard to both physical and geochemical factors. Whereas certain physical properties such as hydraulic conductivity and porosity can be determined by conventional short term aquifer tests, the measure­ment of dispersivity cannot, at present, be evaluated without con­ducting a tracer test at the scale of the expected contaminant plume. Since plumes may be of the order of hundreds of metres, the tracer test must be carried out for a similar time to that needed to establish the pollution plume. Such a test is obviously impractical to conduct.

Similarly, reactions such as biodégradation, nitrate and sul­phate reduction which depend on microbial catalysis may be diffi­cult to simulate in the laboratory since microbial populations take time to develop and are adapted to a range of contaminants in the aquifer. Furthermore, if laboratory data are used they must be validated against similar measurements in field situations before their general use can be established.

At present, very little has been done to validate solute trans­port models by applying them to real contaminant plumes. In instances where models have been applied to contaminant plumes following a chemical spill, discharge of effluent or the migration of leachate from landfills, the value of dispersivity used to simu­late the observed contaminant distribution is usually an empirical factor with no physical significance since it results from averag­ing procedures in groundwater sampling (Anderson, 1979). Averaging procedures may arise from using the analysis of a pumped sample of water from a well which penetrates both clean and contaminated zones in the aquifer. The resulting sample is diluted and this leads to an over-estimation of the dispersivity and of the poten­tial for attenuation. Poor sample preservation prior to analysis or poor sampling techniques can also obscure the elucidation of attenuation mechanism; for example, if an anoxic contaminated groundwater sample is exposed to oxygen, contaminant concentrations can be significantly reduced due to precipitation or oxidation reactions.

To overcome the time scale related problems of measuring dis­persivity and difficulties in simulating biogeochemical conditions in the environment surrounding a landfill, detailed studies have been undertaken at a variety of landfill sites in different hydro-geological environments where pollution plumes have developed. By careful sampling and sample preservation (Stuart and Hitchman,1985) it is possible to determine transport parameters for the aquifer and gain an insight into the mechanisms controlling contaminant migration.

In 1974, the British Geological Survey, (formerly the Institute of Geological Sciences) commenced a major programme of research into the behaviour of hazardous wastes in landfill sites. This

- 107 -

work, funded by the UK Department of the Environment, involved the investigation of 20 landfill sites chosen for the types of wastes deposited and for a range of hydrogeological environments. Attempts were made to identify leachate attenuation reactions and thereby determine which hydrogeological environments are more suited to waste disposal This research included studies of both containment sites and sites where leachates are allowed to migrate, in order to study conditions in the landfill and in the unsaturated and the saturated zones. These investigations were supported by a series of laboratory and small scale controlled field experiments including the use of lysimeters (Anon 1978).

Facilities to monitor, collect and dispose of leachates are usually constructed at containment landfills in low permeability formations such as mudstones or clay, otherwise, in areas where precipitation exceeds évapotranspiration, the site may saturate causing operational difficulties, and eventually leachate will overflow at the surface. Such sites favour study of the landfill and, by measurement of the elements of a water balance (residual rainfall, quantity of leachate removed, and changes in leachate storage in the landfill), the pollutant source term can be defined.

Where landfills are situated on permeable deposits the source term is difficult to determine but if the landfill has been estab­lished for some time, such sites provide the opportunity to study the long term behaviour of contaminants and determine real scale dispersion parameters. The following summaries include the results of site investigations and indicate the direction in which research is progressing.

The Hooton Landfill, Cheshire. The Hooton Landfill is situated in disused brickpits, excavated

in a 10m thick deposit of glacial clay. As much as 3m of clay remains in places as a containment medium. Unsaturated silts and sand up to 12m thick underlie the clay and these rest on 6m of weathered unsaturated Sherwood Sandstone above the water table. Because of access constraints3monitoring the saturated zone down hydraulic gradient of the site has been limited and the study has primarily focussed on conditions in the landfill and in the un­saturated zone (Williams and Harrison, 1983).

A variety of commerical and industrial wastes have been deposi­ted, including ammonium salts and nitrates from the manufacture of fertilizers, and, in places, the waste is saturated to a depth of 6m. Leachate, of similar composition to that of domestic refuse, is pumped from a sump at one end of the site and discharged to a sewer. A water balance for the landfill shows that at least 33% of the rainfall over the landfill area contributes to leachate pro­duction although this ignores the small volume that has been esti­mated to be lost to the aquifer.

Boreholes drilled through the landfill have provided information on the variation in pore water composition with depth both in the landfill and in the unsaturated zone below the residual clay. Within the landfill, pore water composition is relatively constant through the unsaturated waste, but in the saturated landfill there is a gradual decrease in the concentration of labile constitutents with depth (fig. 1). Chloride and ammonia remain relatively con­stant but there is a gradual reduction in the concentration of

- 108 -

sulphate, total organic carbon, and phenol with depth. Methane at a concentration of at least 14% (vol/vol) was detected in the borehole whilst drilling through the unsaturated waste which sug­gests that conditions are sufficiently reducing to allow sulphate reduction and methanogenesis to occur.

Loge (concentration) mg/1

5 6

925mg/l 2460mg/l 3356mg/l

/

' V ' / 1

-Chloride

[Phenol] t = 925e" 0 " 1 ' ' "

t '/2 = 1-57 years

[TOC] 2460e"

t ' /2 = 3-09years

[S04l i = 3356e~0 '2 6 2 1

t '/2 = 2-65 years

Fig. 1. Variation in leachate composition with depth in the Hooton Landfill.

Records show that wastes were deposited directly over a steep face to nearly the full thickness of the landfill. This explains the relatively homogeneous character of the waste in certain boreholes and the small variation in the concentration of con­servative components in pore water. The rate of rise in water level in the waste has been estimated at aproximately lm/year so that the length of time during which the waste has been saturated can be related to depth in the fill. Assuming that labile con­stituents are attenuated predominately in saturated conditions a

- 109 -

rate of "attentuation" which may be due to biodégradation or sorption can be determined (fig. 1). The observed data can be fitted to an exponential decay function giving apparent half lives of 3.09 years for TOC, 2.64 years for sulphate and 1.57 years for phenol. These data are fundamental in defining the source term for contaminant release from the landfill and if regarded as reflecting biodégradation can also be used to assess the rate at which gas (methane and carbon dioxide) is generated from dissolved organic carbon in the leachate. Under these conditions the rate of gas production will be:-

A V R r C (1-e " ) for t ̂ H/r

Where A V R r C e - A t ( e W r - l )

o for t > H/r

unit volume

3,

of landfill A = area of landfill (m ) V = volume of leachate per (dimensionless) Co = initial leachate composition (kg/rrO X = half life for organic decay (years) R = Volume of gas produced by conversion of unit mass of organic carbon (m /kg) r = rate of rise in water in the landfill (m/year) H = maxiumum height to which water is permitted to rise in the landfill (m)

Values typical of the Hooton Landill can be substituted in the equations to predict the rate of gas production in the landfill (fig. 2).

3 0

;

2 0 r = Rate of rise in water table 1m/year

H = Maximum rise 6m

10

TIME (years) •

20

Fig 2. Theoretical variation in gas production in the Hooton landfill with time.

- 110 -

» 103

m < 102

• Unsaturated landfill x Saturated landfill • Unsaturated zone with depth

in metres below base of landfi Perched water Groundwater

1-0 TOC

P 10

CHLORIDE

Fig 3. Correlation between chloride and total organic carbon (mg/1 )-Hooton landfill

® Unsaturated waste x Saturated waste ©3 Unsaturated zone with depth in

metres below base of landfill t s p Perched water table t ° Groundwater

< I Q.

10

1.0 S04 0-1 SO, 0.01 SO/,

10 10 2_ ' 103

CHLORIDE

Fig 4. Correlation betwaen chloride and sulphate(mg/l)-Hooton Landfill

- 111 -

In spite of the landfill being located in clay excavations, boreholes drilled into the unsaturated glacial sands and silts beneath the landfill have shown that leachates have migrated to the water table in the Sherwood Sandstone. However, in any unsaturated zone study to be able to relate attenuation to depth5 the flow must be vertical and the rate of flow through the formation must be known. These conditions are not usually satisfied in studies at existing sites and proper evaluation of unsaturated zone processes are better studied in controlled experiments with the use of lysimeters where a range of instrumentation can be installed to quantify contaminant flow and compositon (Ross, 1985). However, confirmation of processes leading to leachate attenuation in the unsaturated zone must still be attempted in field studies for validation of controlled experiments.

At Hooton, pore water profiles show considerable variation in composition with depth below the landfill and suggest that lateral migration of leachates in perched water tables has occurred. Plot­ting the concentration of labile constituents against conservative species such as chloride however gives an indication of potential attenuation reactions other than dilution. In figures 3 and 4 the lines marked 1.0 TOC and 1.0 SO indicate the range in composition achieved by mixing leachate with groundwater with no attendant chemical reactions. The 0.1 TOC and 0.01 T0C lines represent the range in composition that would result if during mixing the T0C were to be attenuated by a factor of 0.1 or 0.01 respectively. Similar curves are shown for SO . The plots show that although there is a reduction in T0C and SO in pore water within the un­saturated zone, the variation is not related to depth beneath the landfill. Some perched water tables however show very low TOC:Cl and SO : CI ratios which may be interpreted as evidence of atten­uation (sulphate reduction, biodégradation?) in relatively stagnant saturated conditions where gaseous exchange is probably minimal. Overall, however, leachates are not attenuated significantly in migrating through the unsaturated zone of sands and silts although it is possible that organic compounds are undergoing biotrans­formations which cannot be detected by the measurement of T0C. Contrary to previous indications of the effectiveness of the unsaturated zone beneath a landfill for attenuating leachates (Anon, 1978) until more studies are completed, little reliance should be put on the unsaturated zone for attenuating organic components of leachate.

VILLA FARM LAGOONS Migration of contmainants in the saturated zone of an uncon­

solidated sand aquifer has been studied at the Villa Farm Lagoons where, unlike Hooton, access beyond the boundary of the site is relatively good (Williams et al., 1984). The site has been in use for approximately 30 years for the disposal of liquid wastes including heavy metal sludges, pickle liquors, alkalies, phenolic solvents and oil/water mixtures which for a short time attained an input of approximately 7000m /year.

The lagoons, which cover an area of approximately 2000m , were originally excavated through 4m of surface clay to extract sand from a Quaternary lacustrine deposit. The sand aquifer is up to 12m thick and is composed of fine to medium ferruginous quartz sand

- 112 -

with 10-20% carbonate and subordinate clay minerals (illite and kaolinite) and allocthonous coal fragments. The sand is underlain by a poorly permeable clay which has an eroded undulating surface.

The disposal lagoons form a recharge mound some 3m above natural groundwater levels and wastes move radially and downwards into the aquifer under this hydraulic head when the lagoons are dredged to remove the accumulated sludge. The wastes have stratified at the base of the aquifer and regional groundwater flow has advected the contaminants up to 500m away from the site although their distrib­ution is controlled locally by the elevation of the lower clay/sand interface (fig 5).

BH30 BH29

Clay

Running sand

Clay

Rest water level

100 m > 100 mg/ l CI

> 1000 > 3000

> 4000

> 5000

Fig 5. Cross-section through aquifer showing chloride (mg/l)

Natural groundwater is of the calcium sulphate/bicarbonate type with up to 10% (of saturation) dissolved oxygen. Within the pollution plume, ground water is anoxic and three geochemical zones have been identified on the basis of redox reactions in the trans­ition from anaerobic to aerobic conditions:-

(1) A highly polluted zone with low redox potential at the base of the aquifer and near the lagoon in which sulphate and nitrate are reduced. (2) A transition zone of intermediate redox potential in which iron and manganese are mobilised in solution and where only nitrate is reduced. (3) A relatively unpolluted zone characterised by the presence of nitrate and rising levels of dissolved oxygen.

Contaminants migrating from the lagoons are therefore subject to chemical and biochemical transformations depending on the physico-chemical conditions within this range of environments.

The behaviour of contaminants within each zone has not yet been studied in detail but general observations of the attenuation of particular components can be made by comparison with the con­servative species chloride.

- 113 -

In figure 6 the chloride plume is extensive. Phenol (fig 7) is attenuated fairly rapidly probably by biodégradation and/or sorp­tion but the relative importance of either is not known at present. Total organic carbon (fig 8) is not as strongly attenuated as phenol and in the centre of the plume (zone 1) is not significantly attenuated with respect to chloride. TOC:Cl ratios fall in the transition into aerobic conditions at the periphery of the plume suggesting that aerobic conditions favour biodégradation reactions.

1 1

» ~ £ ^

*5>s.r

© «

\ \ \ \

©

—--J* (—\>-^ ®~- 500 - * ^

© © ^ \ \ I /\/xr~

\ \ v Tw—

\ \ M

ipi"0"01 ^a^T'74-v.

1; Î

f^^^^N^' ^"^wii

" \ . \

©

7^

<& I

^ p

'•;P0(ic

f~ !

Scale, metres ViLLA FARM DISPOSAL SiTE

0 25 50 75 WO

Fig 6. Chloride Concentration in Groundwater (mg/1)

VILLA FARM DISPOSAL SITE

0 25 50 75 HO

Fig. 7. Concentration of Phenol in Groundwater mg/1)-1982

- 114 -

Characterisation of the dissolved organic compounds in samples taken from the anaerobic zone (zone 1) at the base of the aquifer with increasing distance from the lagoons nevertheless suggests that organic biotransformations are taking place ultimately with the formation of methane and carbon dioxide which ocurtrs at up to 45% vol/vol in the unsaturated zone above the pollution plume predominantly around 160m from' the lagoons (table 1). The lack of apparent attenuation of TOC should not therefore be taken as implying that no change is taking place.

Distance from lagoon (m)

5

32

164

300

% of TOC identified

5.5

21.4

5.4

1.65

Organic compounds detected (in decreasing concentration)

4-methyl phenol 2,3-dimethyl phenol benzoic acid 4-ethyl phenol 3-methyl benzoic acid 4-methyl-2-pentanone 3,5-dimethyl phenol aromatic carboxylic acid phenol unknown compound

3-methyl benzoic acid 1,2-benzene dicarboxylic unknown compound unknown compound aromatic carboxylic acid 4-methyl phenol 1,4-benzene dicarboxylic 4-methyl-2-pentanone 2,3-dimethyl phenol 1,3-benzene dicarboxylic

3-methy! benzoic acid unknown compound

acid

acid

acid

2-hydroxy-3-methyl benzoic acid unknown compound unknown compound aromatic carboxylic acid 1,3-benzene dicarboxylic aliphatic carboxylic acid 2-methyI benzoic acid unknown compound

3-methyl benzoic acid 1,4-benzene dicarboxylic 1,3-benzene dicarboxylic unknown compound

acid

acid acid

2-hydroxy-3-methyl benzoic acid 2-methyl benzoic acid 1,2-benzene dicarboxylic 1,2,4-benzene tricarboxyl unknown compound

acid ic acid

aliphatic carboxylic acid (heptanoic?)

Estimated concentration

mg/l

8.13 4.66 4.20 3.95 3.06 2.97 2.14 2.14 1.98 1.84

14.99 14.62 10.99 8.80 8.13 4.97 3.54 3.28 1.90 1.78

1.13 0.92 0.83 0.49 0.39 0.34 0.28 0.26 0.25 0.19

0.66 0.40 0.26 0.24 0.24 0.18 0.15 0.13 0.12 0.12

Table 1. Dominant components in groundwater identified by GC/MS(all samples taken from base of aquifer, see Fig. 8 for locations)

Heavy metals typified by nickel (fig. 9) do not migrate far from the lagoons but are precipitated as sulphides and carbonates following sulphate reduction.

This semi-quantitative assessment of the relative attenuation of major contaminant groups is of value but for quantitative evalu­ation of attenuation mechanisms which can be used as a data base

115

VILLA FARM DISPOSAL SITE

•&- SAMPLE POINTS GC/MS ANALYSIS 3o GAS PROSES

0 25 50 75 100

Fig. 8. Maximum concentration of TOC in Groundwater (mg/1

VfLLA FARM DISPOSAL SîTE

0 25 50 75 WO

Fig 9. D is t r ibu t ion of Nickel in Groundwater (rng/1

- 116 -

for evaluating contaminant transport elsewhere it is necessary to define the groundwater flow regime. This is being undertaken using a 3-D numerical model based on field measurements of hydraulic head, and the determination of porewater velocity and hydraulic conductivity with depth in the aquifer. The distribution of chloride is then used to derive longitudinal and transverse dis-persivities. Once these physical transport parameters are known simple geochemical reactions such as biodégradation or non­reversible sorption (exponential decay with time) or retardation (for reversible sorption) which often can be calculated from laboratory measurements of the distribution coefficient, Kd, can be integrated into the model. In simple situations attenuation mechanisms can be quantified by fitting values in a model to simulate observed contaminant distributions.

However, in highly mineralised pollution plumes common to land^ fills simple geochemical models are of limited application and should be regarded as only the first step in biogeochemical model­ling. Chemical models in which the theory of electrolytes is used to predict thermodynamic properties of a multicomponent solution are currently being developed (Jenne, 1979). By these means the equilibrium speciation of chemical components within the system can be predicted although kinetics of the reactions are not usually considered. Coupling aqueous chemical models with groundwater flow is an important area which requires good quality field data and a sound theoretical understanding of aqueous chemistry.

CONCLUSIONS These studies are not comprehensive and considerably more work

is required to quantify the processes that control the attenuation of contaminants in a landfill and in unsaturated and saturated formations. Landfill site investigations are of limited value by themselves but the results of detailed investigations hold impor­tant clues to the processes that control contaminant behaviour which can then be verified and quantified under controlled labora­tory or field conditions. It is considered that future landfill research depends on the development of models in which groundwater flow and (bio) geochemical reactions are coupled in order simulate the observed distribution of contaminants with time around existing well characterised landfills.

Groundwater flow models must be constructed from comprehensive field data of aquifer morphology, aquifer properties and the distribution of contaminants, so that models can be constrained to real field conditions and factors such as dispersivity which are derived from the model are not distorted to account for inappro­priate or poorly collected field data. Biodégradation rates or sorption capacities also deduced from fitting a transport model incorporating reactive solutes to field data should wherever possible be validated against results in independent measurements from laboratory or controlled field experiments to seek out inconsistencies and verify whether microcosm work can be used in real scale predictions„

Unless it can be shown that existing well characterised contaminant plumes can be quantitatively described by models using reliable data, then landfill site assessment will continue to be associated with uncertainty and therefore with a corresponding

- 117 -

degree of risk. Since landfill has come to be regarded as an environmentally acceptable, low cost option for municipal waste, and, in Britain, for the co-disposal of a range of hazardous wastes, the scientific frame-work and basic data for rational site assessment is surprisingly scant.

Acknowledgements

This paper is based on the results of landfill research funded by the Department of the Environment and is published by permission of the Director of the British Geoloigcal Survey (NERC).

REFERENCES

Anderson, M.P., 1979. Using models to simulate the movement of contaminants through groundwater flow systesm. C.R.C. Critical Rev. in Environm. Control, 9(2):pp 97-156.

Anon., 1978. Final Report by the Policy Review Committee on the Behaviour of Hazardous Wastes in Landfill Sites. Department of the Environment. London. Her Majesty's Stationery Office, 169 pp.

Jenne, E.A. (Editor) 1979. Chemical modelling in aqueous systems. ACS Symposisum Series 93, Washington, 913 pp.

Ross, C.A.M., 1985. The unsaturated zone as a barrier to groundwater pollution by hazardous wastes. Mémoires of the I.A.H. 18th Congress: Cambridge, England.

Stuart, A., Hitchman, S.P., 1985, Borehole sampling techniques and field analysis of groundwater in landfill pollution studies = Submitted to Engineering group of the Geological Society, London. 21st regional conference "Groundwater in Engineering", Sheffield, England, 16th - 19th Sept.

Williams, G. M. and Harrison, I»B. 1983. Case study of a containment site - the Hooton landfill, Cheshire. Rep. Fluid Processes Unit, Inst. Geol. Sci., FLPU 83-8. 65pp.

Williams, G.M., Ross, C.A.M., Stuart, A., Hitchman, S.P., Alexander, LoS., 1984. Controls on contaminant migration from the Villa Farm lagoons. Q.J. Eng. Geol. London, Vol. 17, pp 39-55.

- 118 -