Environmental monitoring of landfills

By Philip Mulvey – Environmental & Earth Sciences and Wendi Trotter – Former Environmental & Earth Sciences

 

 Abstract

Groundwater monitoring is usually undertaken to detect the migration of contaminants or leachate from a landfill. Groundwater monitoring is now a requirement of all new, current and some old landfills in New Zealand, set down by the local or regional authority as a result of the Resource Management Act, 1991 (RMA). The positioning of landfill bore and the design of monitoring constituents should be structured to obtain the data necessary to detect and monitor leachate plumes. This requires an understanding of leachate chemistry, characteristics and behaviour to develop a monitoring program that is pro-active and site specific.

 Keywords: landfills, groundwater, piezometers, contaminants, liners, consents.

 

Introduction

The most significant potential for widespread environmental degradation from landfills comes from the potential to contaminate groundwater. Petts, 1993, in a review of landfill risks had four of the eight risks relating to leachate and three of those relate to groundwater. The US-EPA has estimated that 75% of the 55 000 landfills in the USA have contaminated aquifers (cited in Jones-Lee & Lee, 1993). However, vast amounts of the USA, Netherlands, Denmark, Iceland, Japan and New Zealand have ground conditions that do not readily support landfills. Therefore the type and nature of the landfill as well as the location and geology, affect the nature and frequency of groundwater monitoring undertaken.

This paper discusses the issues involved with why we need to monitor landfills, the positioning of monitoring bores and the frequency of monitoring, as well as the constituents monitored.

 

Types of landfills

There are, in regard to landfill disposal, two philosophies of waste disposal; disposal and storage. Within these philosophies there are several different landfill types. These are as defined by Westlake (1995) as;

Disposal

  1. Dilute and attenuate: Essentially previous practice of non engineered landfill.
  2. Simple containment: Defined by ISWA, 1992 as a landfill site where the rate of release of leachate into the environment is exceptionally low such that polluting species are at acceptable concentrations.
  3. Controlled seepage: Accepts that leakage will occur and designs a landfill to safely leak (Mulvey, 1992, Loxam, 1993).
  4. Accelerated bioreactor: Minimises compaction and night cover and optimises moisture content to increase gas generation.

 

Storage

  1. Dry tomb: Endeavours to prevent water entering and leaving waste.

The dry tomb type landfill type is practiced in the USA (Westlake, 1995), however, luckily this approach does not appear to be advocated in New Zealand with the CAE Landfill Guidelines. The approach in New Zealand is to create a controlled bioreactor where conditions optimise the degradation of waste and render it benign to the environment (CAE Landfill Guidelines, 1992).

In New Zealand there are now essentially three landfill types:

  1. Inert waste (includes demolition rubble),
  2. Solid wastes (food waste are not necessarily included at all solid waste landfills), and
  3. Hazardous waste.

Groundwater monitoring is necessary at all three types under the RMA. This paper will essentially address monitoring issues at landfills that contain putrescible materials which include the first two types. The principles also apply to hazardous waste landfills.

 

Why monitor?

The RMA states that landfills are subject to monitoring provisions and this applies to new, current and old landfill sites. The CAE Guidelines for landfills and internal Regional Authority policies regard monitoring as a necessity to ensure there is no measurable affect on ground and surface water resulting from landfilling activities. However, it needs to be understood that – ALL LINERS WILL FAIL at some time, possibly many decades after closure. Thus when designing landfills, waste retention should be seen as an effective measure to protect the environment during the period when pollutants are at their most concentrated state while allowing the natural degradation of waste. This way the leachate will begin to progressively degenerate enabling controlled release into the environment, without lowering environmental quality.

Table I shows leakage rates through membrane lines, compacted soil and composite liners under best average and worst cases. Clearly monitoring is necessary either to detect failure and correct it or to confirm the controlled seepage is behaving as predicted.

Groundwater can be defined as “any water contained in or occurring in an aquifer”, and an aquifer is defined as “a saturated permeable geologic unit that can transmit significant quantities of water under ordinary hydraulic gradients”.

  • Calculated flow rates through liners. flow rates are in litres (ha/day), (after street (1993), cited Westlake, 1995)

 

Type of liner Best case (L/ha/day) Average case (L/ha/day) Worst case (L/ha/day)
Geo-membrane alone 2 500 (2 holes/ha) 25 00 (20 holes/ha) 75 00 (60 holes/ha)
Compacted soil alone 115 (k=10-10 /s) 1 150 (k=10-9 m/s) 11 500 (k=10-8 m/s)
Composite 0.8 (2 holes/ha k=10-10 m/s poor contact) 47 (20 holes/ha k=10-9 m/s poor contact) 770 (60 holes/ha k=10-8 m/s poor contact)

 

The term contaminate is defined as to render impure whilst pollute is to affect the amenity of a user (UK Royal Commission on Environmental Pollution, 1984). Therefore a regulatory authorities environmental goal should be to recognise the feasible and emphases the protection of aquifers without ignoring potential adverse impacts from groundwater in aquitards. Such a goal may require a risk based approach rather than an empirical approach. This more appropriate goal is the basis behind the proposed monitoring strategy.

 

Location of monitoring bores

There are two types of subsurface water monitoring devices, suction lysimeters for the unsaturated zone and bores for the saturated zone. Lysimeters are usually used as part of a leak monitoring system in hazardous waste landfills and are placed beneath the liner. Their use is normally limited to clay liners as installing them under artificial liners is difficult.

It is important when planning an environmental groundwater investigation to assess permeability, gradient, geological medium and direction of groundwater flow to help determine piezometer installation details and location. Areas of natural preferred pathways can be established by defining lineaments from aerial photographs, geological structural mapping and, geophysical techniques. Most of this information can be obtained from quarries prior to landfilling when the landfill has been an old quarry. More monitoring bores are required down gradient than upgradient.

In landfills, interface drainage (perched water table) is often ignored and there is usually no surface expression of interface drainage. Bores should be placed into the interface between the soil layers or the natural rock at the topographical low. Water only collects in these bores during periods of prolonged rainfall or a seasonal wet period and can usually discharges into local water courses some distant down gradient. These interface drainage monitoring bores should be sampled immediately after rainfall.

Table 2 shows the time taken for leachate to reach a piezometer at given distance from the landfill. The landfill has been constructed with a 0.8 m compacted clay liner and collection system (0.5 m head ) compared with one without a collection system (10 m head) in a variety of geologic media and assumes there is no unsaturated follow. When sand, weathered sandstone and permeable volcanic aquifers are immediately beneath the landfill, composite liner systems are required and monitoring bores should be located within 5 m of the landfill and at about 50 m from the landfill. Landfills without a leachate collection system located in weathered shale, siltstone and fractured clays clearly should not be located close to an aquifer and monitoring bores would be ideally located at approximately 10 and 50 m from the edge of the landfill. A landfill in the same environment with leachate collection should have piezometers located at 5 m and between 15 and 20 m from the landfill edge.

  • Time required for leachate to flow various distances

 

  Distance from landfill
(m)
Liner only
h = 10m
Liner and collection
h = 0.5m
Liner alone 309 days 7 years
Sand/some volcanic’s
k = 10-4 m/sec
p = 0.4
5
10
50
1.6 hrs
6.4 hrs
6.7 days
13 hrs
53 hrs
56 days
Clay
k = 10-9 m/sec
p = 0.4
5
10
50
29 years
117years
2940 years
244 years
976 years
2440 years
Fractured Clay
k = 10-6m/sec
p = 0.5
5
10
50
13.4 days
54 days
3.6 years
111 days
1.2 years
30 years
Shale/siltstone
k = 3.1 x 10-9m/sec
p = 0.15
5
10
50
3.5 years
14 years
355 years
30 years
118 years
3000 years
Weathered shale/siltstone
k = 1.0 x 10-6m/sec
p = 0.3
5
10
50
8 days
32 days
2.2 years
67 days
267 days
18 years
Sandstone
k = 2 x 10-7m/sec
p = 0.1
5
10
50
13 days
54 days
3.7 years
111 days
1.2 years
30 years
Weathered Sandstone
k = 5 x 10-5 m/sec
p = 0.2
5
10
50
2.6 hours
10 hours
11 days
1 day
3.5 days
90 days

 

Note(s):

  1. Permeability from Fetter 1988, AGC, 1984, Porosity from Fetter.
  2. Assume conditions are static for the interval.

Landfills located on or above a thick deposit of well compacted clay or tight mudstone/siltstone in the absence of preferred pathways, requires piezometers within 1 m of the landfill edge. As drilling this close may cause fracture pathways, the closest would be 3 m for auger or core drilling rigs and 5 m for air, rotary percussion rigs. Bores beyond 5 m away will be a waste of money for early detection unless located in preferred groundwater pathways. In low permeability ground a positive response is often due to surface water or interface drainage water flowing down the annulus of the hole. Bentonite seals must be used above the test zone and at the surface. In such an instance, all constituents of plume arrive at once, rather than due to attenuation.

 

Sampling intervals

In New Zealand there is no standard set of guidelines specifying sampling intervals, constituents and techniques, as occurs in other countries such as found in some states in Australia. The approach is regulated by the individual regional authorities who generally design a site specific approach. The development of monitoring requirements on new, current and old landfills is a relatively new process to New Zealand regulators and has in the past been conducted on a trial and error basis with a significant involvement from the US EPA documentation. By reference to Table 2 sampling times for different geological conditions can vary significantly. Landfills located above sand, weathered sandstone and permeable volcanic aquifers essentially require semi-continuous monitoring equipment, and therefore the approach in New Zealand will provide more room for variation and site specific monitoring program design.

Continuous monitoring probes are essentially limited to electrolytic conductivity (EC) which is only really useful when the total dissolved solids in the groundwater are below 2 500 mg/L, but can never-the-less be a useful indicator of leachate.

If semi-continuous monitoring is not possible then pH, redox potential (pe), EC, odour and colour should be measured in the field weekly or fortnightly and appropriate constituents ideally monthly, at worst quarterly. In all other instances except stiff clay and unweathered shale, the location of the inner bores should be placed to allow quarterly monitoring. A detection at the inner monitoring bores should be designed to take approximately 3 years to reach the next level of bores and consequently these are monitored infrequently. By using quarterly monitoring an early breakthrough can be detected and an appropriate strategy put into place prior to the groundwater reaching the site boundary. In a controlled seepage landfill remediation should be unnecessary and monitoring is used to collaborate the landfill conceptual model and confirm successful attenuation.

For weathered sedimentary rocks and a thick deposit of stiff clay the inner bores, and bores on preferred pathways, should be monitored quarterly for pH, pe, EC, odour and colour and six monthly or even annually for appropriate constituents. Outer bores should only be added for background sampling or if there is a response in the inner bores.

 

Leachate chemistry

The nature of the reactions in a landfill, and the type of putrescible material partially decides the constituents of the leachate. Typical domestic refuse will contain between 50 to 70% carbohydrate which is readily degraded. Interaction of the by-products of the landfill bioreactor with the night cover and underlining soil/rock also has a significant impact on the nature of the leachate.

Landfill leachate undergoes three decompositional stages;

  1. Immediately after wet deposition, decay begins (heavy compaction of the landfill with night cover only delays the onset of decay). Aerobic microbes use oxygen to convert cellulose and sugars to energy, water and carbon dioxide in a process known as respiration. By another pathway organic acids such as oxalic acid are also readily produced. This creates an initial leachate that is high in organic acids, intermediate fulvic and humic compounds and dissolved carbon dioxide, which is also known as carbonic acid. The presence of these organic acids and the un-disassociated carbonic acid lowers the pH to between 4 and 5. The leachate is also high in chloride sodium sulphate and potassium.
  2. The initial stage is usually very short lived and all available oxygen becomes rapidly consumed as heat is generated. This creates a fermentation vat, providing the acidity is neutralised by the waste and all aerobic bacteria die. This stage converts easily decomposed materials (food, greenwaste and paper) into organic acids, carbon dioxide and water.
  3. Anaerobic fermentation takes over which converts organic compounds to ultimately produce methane and water.

 

Although this is an overly simplified explanation, every stage of the processes is catalysed by micro-organisms, usually bacterium.

During the fermentation process (stage 2), all chemically bound oxygen is consumed, firstly from soluble ions such as nitrate and sulphate and then from minerals or solids containing nitrate, sulphate, iron and manganese minerals particularly goethite (rust). These minerals could be sourced from the rubbish or the soil used as night cover. As the chemically bound oxygen is consumed, increasing amounts of methane (CH4) are produced; and ammonium, iron, and manganese appear as ions in the leachate at the expense of organic acids (they contain organically bound oxygen). The pH increases by the consumption of hydrogen in the formation of ammonium, methane, bicarbonate and hydrogen sulphide. Hydrogen sulphide will not form in the presence of iron and heavy metals, which are precipitated as sulphide resulting in 99.9% metals being confined in the landfill (Belevia Baccini, 1989).

A landfill that becomes rapidly fermenting will maximise gas production and reduce the organic acid and biological demand in the leachate. However ammonium and bicarbonate will increase. It is important therefore to recognise these stages during the landfill monitoring program.

 

Leachate constituents

In summary;

  1. early stage leachate is dominated by sodium, potassium, chloride sulphate, nitrate and organic acids. It has a low pH (4.5 to 6.0) and has a high biological oxygen demand.
  2. Later stage leachate is dominated by sodium potassium chloride and bicarbonate with sulphate and ammonium being co-dominant or subdominant, the pH is neutral. Iron and manganese are minor constituents.
  3. both cases heavy metals, petroleum organics and halogenated organics comprise less than 1% of the leachate constituents.

 

Theoretical movement of contaminants in groundwater

If there was no physical and chemical mechanisms of attenuation, the leachate would arrive at any point at maximum concentration (plug flow), however various mechanisms result in the decrease in concentration of the leachate.

Landfill leachate normally contains high concentrations of ions and discharges into groundwater or surface water containing a lower ion concentration. Initially, and when the leachate discharge is low compared to that of the receiving waters, dilution occurs. This results in a reduction in concentration of the ions which usually has little effect on the comparative dominance of ions in the groundwater (i.e. their ratios to one another). Dispersion, like dilution, results in a reduction in concentration, normally without alteration of the ionic ratios. Alteration of ionic ratios is the result of the preferential removal of one ion compared to the others or the appearance of new ions. Ions are reduced in concentration by precipitation, co-precipitation, ionic substitution and ionic exchange. The latter two also cause the appearance of new ions, as does bacterial action and change in the oxidation/reduction status.

Bacterial decay, initiated within the landfill continues as the leachate moves through the clay liner, unsaturated soil and into the surrounding saturated zone (rock or unconsolidated sediments). Essentially they continue the degradation process of organic acids started in the landfill. Due to their activity, bicarbonate concentrations increase ( at neutral pH ) and sulphate and nitrate reduction continues. The redox activity of the water becomes, if not already, reducing (i.e. pe is negative ).

 

Analysis of data

The analysis of monitoring data needs to focus on the understanding of how leachate evolves and degrades. Monitoring programs should be designed to recognise the degree of degradation of a leachate plume and its known and potential impact on the ground and surface water.

 

Leachate plumes

The nature of the plume both in concentration and how it is attenuated depends on the initial concentration and local ground conditions. Therefore different geological media produce difference plumes.

A heavily clayey soil or rock with clay minerals and low permeability will attenuate and fragment the plume, resulting in widely separated distinct phases of arrival. A non-reactive porous medium will result in a poorly attenuated plume – the plume arrives without distinct phases and is essentially a dilute leachate. A positive detection in a bore may not be due to the groundwater thought to be monitored. If a plume in an attenuating environment arrives as a pulse, rather than in distinct phases or the phases are only slightly separated, leakage down the hole from interface drainage or surface runoff may be the cause. Normally in such a bore, standing water levels show greater variation than other bores.

 

Identifying leachate breakthrough

Table III presents the groundwater monitoring results for a general landfill (no household or domestic refuse) situated in sandy clayey sub-base with saline groundwater. Groundwater has been monitored during and after the operational phase of landfill.

  • BH1 is a leachate sample taken just after the landfill had been capped.
  • BH7 was initially sampled just after the first phase of a pollution plume, and before the actual closure of the landfill. The bore is located within 10 m of the landfill edge.
  • BH13 is a background bore located 25 m up gradient from outer landfill edge and the results are representative of background water quality.

Monitoring began five months before closure of the landfill.

Leachate is less saline than the natural water, but it is dominated by ammonium, sodium and potassium, while the natural water is dominated by sodium, calcium and magnesium. To highlight the breakthrough of displaced native cations and leachate breakthrough, a special ratio known as the L/N ratio has been derived.

The L/N ratio is defined as the addition of the dominated cations in the leachate divided by the addition of the dominant cations in the groundwater multiplied by a factor. When a cation is dominant or co-dominant in both waters, it is included in the divisors only. Usually the difference in the L/N ratio of leachate and groundwater is at least two orders of magnitude.

Equation 1                         *L/N = (K + NH4) / (Mg + Ca + Na) x 100

The results in Table 3 indicate contaminant concentrations are at or just below background, thus only by the use of ratios can the characteristic of the plume be defined. Unfortunately, monitoring commenced as the initial phase of a leachate plume was occurring however, compared to the upgradient background bore BH13, BH7 has a classic breakthrough signature.

A leachate breakthrough signature is recognised by;

  1. The first phase has relative increases in bicarbonate and sulphate (to chloride) and a reduction in nitrate have occurred with calcium + magnesium + sodium (native cations, N) have increased relative to potassium + ammonium (leachate cations, L) indicating the displacement of native cations has occurred in advance of the main front.
  2. The second phase is apparent in the next sampling, BOD rises sharply, ammonium and iron start to rise and relative to chloride, sulphate is reduced, and potassium + ammonium rise, relative to calcium + magnesium + sodium.
  • water monitoring data for a non domestic and an industrial refuse landfill

 

  BH1 BH7 BH7 BH7 BH7 BH7 BH7 BH7 BH7 BH13 BH13
Months after closure 0 -5 5 5 8 14 18 20 24 11 17
pH 7.2 6.8 6.8 6.8 7 6.7 6.6 6.6 6.8 7.3 6.9
TDS 5780 5490 6200 5900 6200 6550 7450 6870 6920 8456 8420
BOD 290 5 40 10 22 20 35 30 20 <2 9
Iron 0.62 0.06 1.4 0.15 0.1 0.7 0.09 0.1 0.07 0.1
Sodium 930 1760 2090 1860 1860 2000 2330 2090 2190 2810 2630
Calcium 51 92 20 135 140 130 150 125 130 190 170
Potassium 590 15 38 41 50 46 44 49 49 27 28
Magnesium 125 180 210 220 260 280 270 280 280 340 285
Ammonia 1050 0.6 4.9 7.9 12 34 8.5 14 19 1.7 1.1
Chloride 1290 2350 2480 2250 2570 2660 3250 2940 2880 4240 4120
Nitrate 39 <0.1 <0.1 0.13 <0.1 <0.1 3.5 <0.1 0.63 3.1 1.6
Sulphate 100 420 120 130 130 140 150 1150 120 380 390
Bicarbonate 5260 1430 2470 2520 2370 2380 2480 2400 2430 1720 1530
C1/SO4 13 5.6 21 17 20 19 22 19 24 11 11
C1/HCO3 0.25 1.64 1.00 0.89 1.08 1.12 1.31 1.23 1.19 2.47 2.69
Ca/K 0.09 6.1 0.53 3.3 2.8 2.8 3.4 2.6 2.6 7.0 6.1
C1/TDS 0.22 0.43 0.40 0.38 0.41 0.41 0.44 0.43 0.42 0.50 0.49
Na/K 1.6 117 55 46 37 43 53 43 45 104 93
L/N* 104 0.76 1.84 2.20 2.74 3.31 1.9 2.52 2.6 0.85 0.9

 

Earthworks for closure, seasonal effects or two distinct pulses rather than a single continuous source are other variations that may influence the monitoring data and explain variable behaviour of leachate plumes.

 

Conclusion

Landfill monitoring should be undertaken to confirm controlled seepage of leachate is behaving as predicted and that it is having little or no measurable effect on the natural ground and surface water. Groundwater monitoring, bore location and the frequency of monitoring depends on the nature of the surrounding soil or rock within which a landfill is located. Location and frequency of monitoring should not be haphazard or uniformly applied, instead they should be tailored for local site conditions. A simplistic local conceptual model should be derived to highlight the monitoring requirements. Generally regional authorities in New Zealand appear to be taking this approach.

When interpreting the monitoring data focus should be centred on recognising the stage and nature of leachate breakthrough fronts. In general the initial breakthrough front will see an alteration to the L\N ratio, and a rise in chloride, bicarbonate and sulphate. This is usually followed by a second front that is high in biological oxygen demand, chloride, bicarbonate, ammonium and potassium and lower in nitrate and sulphate. An increase in iron and manganese normally follows. Therefore constituents monitored should be pH, electrical conductivity and redox potential in the field as well as chloride, bicarbonate, sulphate, BOD, ammonium, nitrate, calcium, manganese, sodium and potassium analysed in the laboratory. Only after the second front has passes should heavy metals be monitored. Ratios, particularly the L/N cation ratio, should be used to highlight differences in the anions and cations.

 

References

  1. Australian Groundwater Consultants. (1984). Effects of Coal Mining on the Groundwater Resources in the Upper Hunter Valley. NSW Coal Association.
  2. Belevi, H. and Baccini, P. (1989). Long term assessment of leachates from municipal solid waste landfills. In Sardinia 89, Proceedings of the Second International Landfill Symposium (Vol. 1, pp. 1-8). CIPA, Milan, XXXIV
  3. Fetter, C. (1988). Applied Hydrogeology Merrill Publishing Co.
  4. Jones-Lee, A. and Lee, G. F. (1993). Groundwater pollution by MSW landfill: leachate composition, detection and water quality significance. In Sardinia 93: Proceedings of the Fourth International Landfill Symposium 11-15 October. S Margherita di Pula: Cagliari, Italy.
  5. Lee, G. F. and Jones-Lee, A. (1993). Landfill and groundwater pollution Issues: “Dry tomb” vs Wetcell landfills. In Sardinia 93: Proceedings of the Fourth International Landfill Symposium.
  6. Loxham, M. (1993). The design of landfill sites – some issues from a European perspective. In Landfill tomorrow – Bioreactors or storage. In Proceedings of seminar held at Imperial College of Science. Technology and Medicine: London.
  7. Ministry for the Environment. (1992). Landfill Guidelines. Set by Center for Advanced Learning.
  8. Mulvey, P. (1992). Proactive Assessment: science derived not recipe driven. In RACI Second Environmental Chemistry Conference Abstracts. Melbourne.
  9. Petts, J. I. (1993). Risk management, risk management and containment landfill. In 1993 Hartwell Waste Management Symposium: Options for landfill. In AEA Technology. Harwell: Oxfordshire, K.

Westlake. (1995). Landfill Waste Pollution and Control

 

© Copyright 1999 – 2004 Environmental & Earth Sciences International
All rights reserved. No part of this paper may be reproduced or copied in any form or by any means without the written permission of the publisher.

Related Projects:

Request a call back

If you'd like us to give you a call or provide you with a free quote please fill in your details below and we'll be in touch.