YOUNGER_Mine Water Chemistry. the Good, The Bad and the Ugly

Research article

Environmental Geology 32 (3) October 1997 7 Q Springer-Verlag 157

Mine-water chemistry: the good,the bad and the uglyDavid Banks 7 Paul L. Younger 7 Rolf-Tore Arnesen 7 Egil R. Iversen 7 Sheila B. Banks

Received: 13 August 1996 7 Accepted: 3 January 1997

David Banks1 (Y)86 Holymoor Road, Holymoorside, Chesterfield, Derbyshire,S42 7DX, UK

P. L. YoungerDepartment of Civil Engineering, Cassie Building, University ofNewcastle, Newcastle-on-Tyne, NE1 7RU, UK

R.-T. Arnesen 7 E. R. IversenNorwegian Institute for Water Research (NIVA), Postboks 173Kjelsås, N-0411 Oslo, Norway

S. B. Banks1

Scott Wilson CDM, Bayheath House, Rose Hill West,Chesterfield, Derbyshire, S40 1JF, UK

Present address:1Norges Geologiske Undersøkelse, Postboks 3006 – Lade,N-7002 Trondheim, Norway

Abstract Contaminative mine drainage watershave become one of the major hydrogeological andgeochemical problems arising from mankind’s in-trusion into the geosphere. Mine drainage waters inScandinavia and the United Kingdom are of threemain types: (1) saline formation waters; (2) acidic,heavy-metal-containing, sulphate waters derivedfrom pyrite oxidation, and (3) alkaline, hydrogen-sulphide-containing, heavy-metal-poor waters re-sulting from buffering reactions and/or sulphate re-duction. Mine waters are not merely to be per-ceived as problems, they can be regarded as indus-trial or drinking water sources and have been usedfor sewage treatment, tanning and industrial metalsextraction. Mine-water problems may be addressedby isolating the contaminant source, by suppressingthe reactions releasing contaminants, or by activeor passive water treatment. Innovative treatmenttechniques such as galvanic suppression, applica-tion of bactericides, neutralising or reducing agents(pulverised fly ash-based grouts, cattle manure,whey, brewers’ yeast) require further research.

Key words Mine-water chemistry 7 Acid minedrainage 7 Mine-water pollution 7 Mine-watertreatment

Objective

The objective of this paper is to review the variety of hy-drochemical characteristics, environmental impacts andtreatment methods pertaining to waters discharging frommines or mine waste deposits. The paper will draw pre-dominantly on examples from the United Kingdom andNorway, many of which have not previously been re-ported in mainstream literature, to illustrate this variety.A list of references is provided for the reader wishing toobtain further details of the sites reviewed.

Introduction

In the “First World”, many coal and metal ore mineshave been closed, for economical or political reasons,during the past decade (Younger 1994; Coldewey andSemrau 1994; King 1995; Dumpleton 1995). Such closurescan lead to a variety of environmental impacts, some ofthe most acute being the effects on the hydrogeologicalenvironment due to cessation of pumping. Such negativehydrogeological impacts are described by Younger (1993,1994) and may include inundation of low-lying areas(particularly areas affected by subsidence which may beregional and up to several metres magnitude; Phillips andHellewell 1994; Yao and Reddish 1994), contaminant mo-bilisation and transport via rapid groundwater flow path-ways (i.e. mining fractures and openings), subsidence ofshallow workings and elevated mine gas emissions.All of these problems are potentially serious, but this pa-per will focus on the problem known as “acid mine drai-nage” (although it is not always acid, nor always a prob-lem) as it presents a particularly acute and urgent chal-lenge to many countries (NRA 1994; King 1995). Theproblem is not a new one, but has been present in someguise since the commencement of deep mining necessi-tated the dewatering of drainage water to surface water-courses. In fact, the names of some rivers, such as thenorth Cornish Red River, the Ochre Dykes of Derbyshireand South Yorkshire, the Yellow Stream of Lancashire,the Norwegian Raubekken and the Spanish Rio Tinto in-dicate the historical nature of mining-related water con-tamination.

Research article

158 Environmental Geology 32 (3) October 1997 7 Q Springer-Verlag

Fig. 1Simplified diagram illustrating reaction pathways for pyriteoxidation (after Stumm and Morgan 1981; Arnesen 1993b)

Chemical evolution of minewaters

Pyrite oxidationMining allows the introduction of oxygen to the deepgeological environment and thus the oxidation of miner-als which are in a reduced state. The same occurs whenreduced minerals are brought to the surface and deposi-ted in spoil tips (Wiggering 1993a, b). The most commonfamily of such minerals are the sulphides. Unlike mostgeochemical weathering processes, the oxidation of somesulphides (of the type MS2) leads to the production, rath-er than the consumption, of protons (i.e. acid). The irondisulphide, pyrite, is ubiquitous in most metal sulphideand coal deposits and may exist in potential associationwith other chalcophile elements such as As, Bi, Cd, Co,Cu, Ga, In, Hg, Mo, Pb, Re, Sb, Se, Sn, Te and Zn (Spearsand others 1994). Pyrite undergoes a complex cycle ofreactions during oxidation (Eq. 3; Fig. 1), but in simpli-fied form the following equation describes the net proc-esses:

2FeS2c2H2Oc7O2p2Fe2cc4SO42–c4Hc

(aq) (1)Pyritecwatercoxygenpferrous ironcsulphatecacid

Further partial oxidation of ferrous to ferric iron con-sumes some protons:

4Fe2cc4Hc(aq)cO2p4Fe3cc2H2O (2)

Ferric iron may act as an electron acceptor for furtherpyrite oxidation, or hydrolysis may occur, both processesreleasing further protons:

FeS2c14Fe3cc8H2Op15Fe2cc2SO42–c16Hc

(aq) (3)

Fe3cc3H2OpFe(OH)3c3Hc(aq) (4)

The overall sequence of reactions is acid-producing:

4FeS2c14H2Oc15O2p4Fe(OH)3c8SO42–c16Hc

(aq) (5)

Where mine-water pumping is constant and the mine-water level is stable, little pyrite oxidation occurs below

the water level, and few metals are leached from above,resulting in a relatively non-environmentally aggressivemine water. Active pyrite oxidation will, however, contin-ue to occur in the unsaturated zone and, if pumps areturned off, the rising water level will leach out a largeamount of heavy metals, resulting in a highly acid andcontaminating solution. This affect has been observed inthe Durham Coalfield by Younger (1993) and Youngerand Sherwood (1993), as well as at Wheal Jane tin mine.Other sulphide minerals will oxidise similarly to pyrite,releasing heavy metals and sulphate. It is interesting tonote, however, that oxidation of common sulphides ofthe form MS does not in itself release acid, e.g. sphaler-ite:

ZnSc2O2pZn2ccSO42– (6)

Subsequent hydroloysis of metal ions may release pro-tons, but equilibrium speciation modelling of sphaleriteoxidation-dissolution (using MINTEQA2 and/or WA-TEQ4F) indicates that pH is unlikely to be depressed be-low 5.5 due to partial hydrolysis of zinc ions.

Buffering reactionsThe acidity generated by such reactions only results in asignificantly depressed pH in the mine water when it ex-ceeds (on an “equivalents” basis) the alkalinity availablein the system. Such alkalinity may be available as bicar-bonate in groundwater (generated by normal weatheringprocesses or reduction processes) or in the form of mi-neral phases.The most prolific mineral sources of alkalinity are thecommon carbonate minerals, calcite and dolomite (Nilsenand Grammeltvedt 1993). These are not, however, alwayspresent in significant quantities: in the Coal Measures ofnorthern England, for example, the most common car-bonate minerals are the ferrous varieties ankerite[(Ca,Mg)FeII(CO3)2] and siderite (FeIICO3). Dissolution ofthese minerals will allow localised and temporary neutral-isation only, for when oxidation and hydrolysis of theirferrous iron finally occurs, there will be a release of pro-tons (see Eqs. 2 and 4), which will largely (if not com-pletely) negate the previous neutralisation (Morrison andothers 1990).In many coal mines, limestone dust is spread liberallyover mine surfaces to hinder the mobilisation of explo-sive coal dust and to act as a flame “barrier” in the caseof an explosion. This limestone dust may also serve toneutralise mine water from working mines. This may beone of the reasons why the pumped mine water at mine3, Longyearbyen, Svalbard (World Coal 1995) emerges ata pH of 8.2, despite a sulphur content of some 1% in theLongyear Coal (Amundsen 1994; Banks 1996; see Table 1).Additionally, it is believed that the alkaline mine waterhere is also due to the natural hydrochemistry of the sub-permafrost groundwater entering the mine at its deepestlevels.In the absence of readily soluble carbonate minerals, dis-solution of alumino-silicate minerals (mafic silicates, feld-spars and clays) can make a significant contribution to

Research article


Tabl

e 1

Com

pari

son

of d

isch

arge

wat

er q

ualit

y fr

om s

elec

ted

Bri

tish

and

Sva

lbar

dian

pum

ping

and

aba

ndon

ed m

ines

and

spo

il ti

ps (

sour

ces;

You

nger

199

3; M

adaw

ee 1

994;

Fug

e an

dot

hers

199

1; L

emon

199

1; B

ucha

n 19

62; B

anks

199

6; B

anks

and

oth

ers

1996

b)

Dis

char

ge(l

/s)

pHT

DS

(mg/

l)Fe (m

g/l)

Al

(mg/

l)M

n(m

g/l)

Zn

(mg/

l)C

u(m

g/l)

ClP

(mg/

l)SO

42P

(mg/

l)

Pum

ping

coa

l m

ines

Kib

bles

wor

th (

Dur

ham

)N

icho

lson

s (D

urha

m)

Kim

bles

wor

th (

Dur

ham

)T

ilman

ston

e (K

ent)

Moo

rgre

en P

iper

(E.

Mid

land

s)Lo

ngye

arby

en M

ine

3 (S

valb

ard)

a

740 90 105

c. 0

.06

7.1

7.1

7.3

7.7

6.9–

7.9

8.2

3185

3100

1800

2107

0.63

5.8

5.0

25~

0.1–

7.0

~0.

01~

0.02

0.00

4

0.05

60.

034

0.03

0

0.05

5~

0.00

5

795

3600

–10

800

236

690

1170 38

040

4 7.4

Aba

ndon

ed c

oal

min

esSt

ony

Hea

p (D

urha

m)

Dun

ston

(C

hest

erfi

eld)

R. H

ippe

r D

isch

arge

(C

hest

erfi

eld)

Duk

e’s

Leve

l, B

uxto

n (D

erby

shir

e)Y

nysa

rwed

(So

uth

Wal

es)

Hap

ton

(Lan

cash

ire)

c. 2

0c.

0.7

5c.

20

c. 3

5c.

9

6.3

6.3

3.6

6.3

4.2

7.9

630

2.2

10.6

101.

34.

918

0 0.09

7

~0.

045

17.3 0.07

8~

0.5

~0.

01

1.26

4.02

0.36

6.1

0.13

8

~0.

007

0.22

10.

048

0.06

1~

0.00

5

~0.

007

0.00

70.

005

0.00

07

26 29 18 32 21

148

210

1044 83

1554 17

6

Coa

l sp

oil

tips

Cro

ok (

Dur

ham

)Q

uaki

ng H

ouse

s (D

urha

m)

Oat

land

s (C

umbr

ia)

Thu

rcro

ft (

S. Y

orks

hire

)b

Long

year

byen

Min

e 3

(Sva

lbar

d)Sv

erdr

upby

en M

ine

1 (S

valb

ard)

c. 0

.1c.

0.2

5

3.5

4.1

5.5

6.8

3.7

2.7

1000

2314

70 15 287 18

.6 1.6

179

0.97

~0.

045

1.8

27.5

5.2

2.0

0.40

3.2

0.05

~0.

007

0.49

1.3

~0.

007

0.01

40.

168

511 4.

57.

0

810

1358 14

613

27 7710

77

Met

al m

ines

Cae

Coc

h (W

ales

, pyr

ite)

Cw

m R

heid

ol (

Wal

es –

Pb,

Cu,

Zn)

Alle

n H

ill S

paw

(M

atlo

ck, D

erby

shir

e)c.

0.1

5 l/

s

2.5

2.8–

3.0

6.5

1460 14

.9

84.2

113

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0.1

0.13

2

3.05

2.10

0.94

38–7

20.

029

0.16

0.03

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6882

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5110 44

1–84

612

4

Met

al m

ine

spoi

l ti

psC

wm

Rhe

idol

(W

ales

– P

b, C

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710

4–12

857

7–97

81.

2–9.

3579

1

EC l

imit

s fo

r dr

inki

ng w

ater

6.2–

8.5c

0.05

c

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0.10

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f th

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(Ban

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Research article


neutralisation. However, the rates of dissolution of theseminerals are often significantly slower than those of thecarbonates, and hence non-equilibrium rock-water inter-action can be expected where flow is reasonably rapid. Inparticular, more unstable (mafic) silicate mineral phasessuch as olivine, pyroxene or anorthitic plagioclase will bemore effective at neutralising acidic waters. This is illus-trated by the case of Nikkel og Olivine AS’s nickel sul-phide mine at Ballangen, Nordland, Norway (Arnesenand Iversen 1995), where the high content of olivinearound the ore body leads to a slightly alkaline mine wa-ter, with dissolved nickel and, to some extent, arsenic,being the only significant metallic contaminants. Itshould be emphasised, however, that simple acid-base ac-counting will not be able to predict mine-water quality.Relative rates of reaction and consumption of carbonate,silicate and sulphide phases will be decisive.In particularly aggressive (low pH) mine waters, evenclay minerals may undergo reaction and provide someneutralising capacity; for example, kaolinite:

Al2Si2O5(OH)4c6Hcp2Al3cc2H4SiO4cH2O (7)

This explains the high aluminium concentrations oftenfound in mine waters of pH 4 or lower.

Sulphate reductionIn some mine systems, particularly those overflowingfrom saturated workings (Fig. 2), opportunities for bacte-rially catalysed sulphate reduction may occur, generatingmetal sulphides and/or hydrogen sulphide and alkalinity(Younger and Bradley 1994; Younger 1995b). The pH ofthe water is thus elevated. In extreme cases (for example,at Hapton Colliery in the Lancashire coalfield; Table 1),pH values of around 8 can occur, in conjunction with ne-gligible quantities of toxic metals and considerableamounts of hydrogen sulphide. The diffusion of methaneinto workings may be a factor in promoting highly sul-phate-reducing conditions.

Ochre depositionThe iron discharged in coal mine drainage is typicallylargely in the form of ferrous iron (Hedin and others1994). Above a pH of 4 (most coal mines), ferric iron istypically absent, due to rapid hydrolysis and precipita-tion. Below a pH of 2.5 (e.g. in many spoil tips or sul-phide mines), hydrolysis proceeds only very slowly andferric iron can remain substantially in solution. On expo-sure to the atmosphere, ferrous iron will tend to be in-creasingly oxidised to ferric, accompanied by hydrolysisand at least partial precipitation as an iron oxyhydroxide(ochre):

4Fe2ccO2c4Hcp4Fe3cc2H2O oxidation (8)

Fe3cc3H2OpFeOOHcH2Oc3HcpFe(OH)3c3Hc hydrolysis (9)

Other heavy metals will co-precipitate with iron oxyhy-droxide or be adsorbed onto the iron oxyhydroxide mass.

Fig. 2Schematic illustration of two types of mine system,characterised by (Type 1) rapid unsaturated flow and intensepyrite oxidation, yielding net acidic waters and (Type 2) slow,saturated flow, buffering reactions and sulphate reduction,yielding net alkaline waters

In very low pH waters (pH~4), dissolution of aluminiummay occur from alumino-silicate minerals such as kaolin-ite. When aluminium-rich acidic mine waters are effluentinto higher pH surface waters, they typically precipitatewhite aluminium hydroxide (gibbsite).

Al3cc3H2OpAl(OH)3c3Hc (10)

Role of bacteria in acid mine drainageThe oxidation of pyrite in mine drainage waters is cata-lysed by the action of acidophilic sulphide-oxidising bac-teria, in particular, Thiobacillus ferrooxidans, whichthrives at a pH range of 1.5–3.0. Thiobacillus ferrooxi-dans, a chemoautotroph, derives energy for its metabolicprocesses from the oxidation of reduced sulphur and ironcompounds and utilises CO2 as a carbon source. By cata-lysing the oxidation of ferrous sulphide to ferric sulphate,this bacterium greatly accelerates the otherwise slowchemical oxidation of iron sulphide.Although Thiobacillus ferrooxidans is probably the mostwell-known and predominant bacterium associated withacid mine drainage, a range of other iron/sulphur-oxidis-ing bacteria exist which have been isolated from coalspoil heaps and drainage waters (McGinness and Johnson1993). These include Leptospirillum ferrooxidans, Thioba-cillus thiooxidans and Sulfobacillus thermosulfidooxidans.Spoil heaps and mine drainage waters can be expected to

Research article


contain a mixed population of these and other acido-philic bacteria, the exact species present and their relativeactivities depending on the prevailing conditions, includ-ing temperature and pH (Norris 1990).Bacteria are known which promote and/or catalyse theoxidation, hydrolysis and precipitation of dissolved iron(ochre formation), although their role in mine waterchemistry has yet to be fully understood (McGinness andJohnson 1993; Hedin and others 1994). It is thought thatabiotic oxidation of dissolved ferrous to ferric iron is fas-test at elevated pH levels and, hence, in net alkaline wa-ters with pH 6–7, this process should dominate. At lowerpH values (2–3) bacterial oxidation is at its fastest anddominates in substantially acidic waters. The use of me-tal-immobilising bacteria has been discussed as a poten-tial treatment method for heavy-metal contaminated ef-fluent (Gadd 1990; Banks 1992).

Other quality problems associated with minedrainage

The oxidation of sulphide minerals to release heavy me-tals, sulphate and acid is the fundamental reaction char-acterising acid mine drainage. The quality of mine drai-nage water may, however, be detrimentally affected byother parameters than these, of which the most promi-nent are probably the following:1. Salinity-related parameters – deep mining may en-

croach upon bodies of stagnant, or even fossil,groundwater of high salinity. Dewatering of coastalmines may also cause the intrusion of modern seawat-er. In all these cases, a high salinity may be expectedin the mine drainage water, potentially resulting incontamination of fresh surface water or groundwaterwith chloride. This was the case at Tilmanstone collie-ry in the Kent Coalfield, U.K. (Headworth and others1980), where mine drainage water containing around318,000 tonnes of chloride at concentrations of up to5,000 mg/l was discharged into the Chalk aquifer be-tween 1907 and 1974. In deep saline brines in UKcoalfields (Andersen 1945; Downing and Howitt 1969),elevated concentrations of ammonium in dischargewater may also pose a pollution threat (NCB 1982;NRA 1995). In some deep saline mine waters, reduc-tion of sulphate occurs, permitting the solution ofhigh concentrations of Ba and Ra in the absence ofsulphate solubility controls (Lebecka and others 1994).

2. Nitrates – where N-based explosives have been exten-sively used for mining or quarrying, residues of explo-sives in the spoil or in the mines themselves may beoxidised to nitrates and mobilised in groundwater.This problem resulted in the closure of public watersupply boreholes down-gradient of a tunnel drift atKopperå, Norway (T. Moseid, personal communica-tion). In Ballangen mine, Norway, suspended solids,drilling mud, lubricating oil and residues from explo-sives are cited as significant contaminants in the minewater (Arnesen and Iversen 1995).

3. Organic parameters – the organic quality of coal minedrainage water has not been studied in detail, but

trace contamination by organic compounds derivedfrom coals is possible. One study performed in the lig-nitiferous regions of Romania, Serbia and Bulgariafound a tentative correlation between (as yet uniden-tified) refractory fluorescent organic substances ingroundwater, possibly derived from contact with lig-nite strata, and the incidence of Balkan endemic neph-rosis, a kidney complaint (Goldberg and others 1994).

The good: useful mine waters

Potable mine waters: lead mining, Derbyshire, UKThe Carboniferous limestone of Derbyshire has beenmined extensively for many centuries for Pb and fluoride.The minerals fluorspar and galena, with associated sphal-erite, barite, calcite, pyrite and chalcopyrite occur togeth-er in hydrothermal veins. Sub-vertical veins or “rakes”were mined at several underground levels to impressivedepths. The mining often penetrated below the water ta-ble and the mines were drained by free-flowing adits or“soughs” to local rivers. It is perhaps surprising thatthese soughs (and also lead mining shafts) serve as high-quality, large-scale drinking water supplies today, oper-ated by regional water undertakings (Edmunds 1971).Table 2 demonstrates the generally wholesome nature ofthe water from these soughs. As might be expected for alimestone groundwater, the water is hard, rich in Ca andslightly alkaline. Sulphate levels are somewhat elevatedwith respect to EC guideline levels and other nearby Car-boniferous lithologies such as the Millstone Grit (Ed-munds and others 1989), indicating that some sulphideoxidation or sulphate dissolution is occurring. AlthoughZn and Ba are present at slightly higher levels than aredesirable, heavy metals generally, and Pb in particular,pose no major water quality problems. There are severalpossible reasons why the fluoride and toxic metals arenot present in excessive concentrations (Albu and others1997):1. Kinetic factors – the relatively slow rate of dissolution

and oxidation of sulphide minerals compared with therapid flux of limestone groundwater through the mineconduits.

2. The solubility of heavy metals is suppressed by thehigh alkalinity of the water and the solubility of fluo-rides by the high Ca concentrations. Geochemical spe-ciation modelling indicates that Ba concentrationsseem to be limited by barite solubility. Barite (BaSO4)and calcite are the only two common minerals (apartfrom the aluminium minerals) with respect to whichthe waters are saturated. Minerals such as fluorite andotavite approach saturation in some waters and, forZn and Pb, the respective carbonates yield saturationindices (SI) in the vicinity of –1. These figures may beregarded as consistent with saturation of waters withrespect to fluorite and heavy metal carbonates occur-ring in the immediate vicinity of the ore body (andthus limiting solubility), the water later being diluted

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Table 2Water quality from lead mine soughs in the Carboniferous Limestone of Derbyshire (data from Edmunds and others 1989)

pH EC(mS/cm)

Ca(mg/l)

HCO3P

(mg/l)SO4

2P

(mg/l)F(mg/l)

Ba(mg/l)

Cu(mg/l)

Cd(mg/l)

Fe(mg/l)

Pb(mg/l)

Zn(mg/l)

EC limit

Meerbrook SoughHillcarr SoughMandale SoughMagpie SoughWaterloo Inn Sough

6.2–8.5a

7.27.27.17.27.0

400a

580540530520420

100a

8688989898

262261257261222

25a

250b

5736273330

1500b

680750610640320

100a

70280327336229

100c

0.90.60.40.40.5

5b

1.6~0.5

0.80.8

~0.5

50a

200b

~0.510.2

~0.5~0.5

0.8

50d

17~6

16~6~6

100c

338297

8574

7

a Guide limitb Maximum permissible concentration

c At discharge from pump or treatment plantd At point of consumption

by normal limestone groundwater during its passagealong the sough.

3. The surfaces of metal sulphides being coated with in-soluble oxides or carbonates, preventing further oxida-tion.

4. The fact that a hydraulic equilibrium situation haslong been in place, with limited water-table fluctuationand thus limited flux of oxygen into the veins to pro-mote sulphide oxidation.

5. The metals ores having been extremely efficiently re-moved from the veins by largely manual working.

The lead mines of Derbyshire thus provide a clear dem-onstration of the importance of the host rock lithology(rather than the ore mineralogy) for the quality of themine drainage water.Virtually identical conditions obtain in the former leadmines of the Greenhow district, North Yorkshire, where amajor nineteenth century drainage adit (the Eagle Level)is exploited for public water supply by Yorkshire Waterplc. Still further north, in the Pb-Zn orefield which strad-dles the Northumberland/Durham border, water from theabandoned Hunstanton mine was formerly used for watersupply in and around the village of Blanchland. However,other mine waters in that orefield are rather less whole-some than those listed in Table 2, with major ferruginousand/or zinciferous discharges occurring in upper Wear-dale and South Tynedale.

Other uses of mine watersBanks and others (1996a) have described in some detailthe other uses and positive functions of mine waters inCentral and Eastern England (Fig. 3). These include:1. The use of mine waters in sewage treatment as floccu-

lating agents. Ochreous mine waters from the Buxtoncolliery in Derbyshire were, from shortly after 1886,piped some 4 km to Buxton treatment works wherethey were used as flocculating agents. Currently, minewater from the copper mine complex at Falun inSweden is mixed with municipal sewage at the Främbytreatment works as a form of flocculating agent, al-though the residue does admittedly need to be deposi-ted at a special waste site due to content of heavy me-tals (some 44 mg/kg of Cd, 52.5 g/kg of Zn, 1.0 g/kg of

Fig. 3Map of Great Britain, showing locations of coalfields andselected named sites

Cu in 1990). In 1995, some 109,000 m3/year mine waterwas blended with 5,900,000 m3/year sewage (G. Strålin,Falu kommune, pers. comm.).

2. Potable water treatment – Dudeney and others (1994)proposed that ferruginous mine drainage from Bull-house and Handbank mines near Stocksbridge, in the

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South Yorkshire Coalfield, could be employed in theflocculation process for treatment of surface waterfrom the nearby Langsett Reservoir.

3. Mine drainage waters have been employed as mineralwaters or spas in historical times (Allen Hill Spaw of1824 in Matlock, Derbyshire, UK; Naylor 1983) and arestill considered for this purpose in some nations at thepresent day (at coal mines in China; Jianli and others1993). Currently, it is reported that mine water (ofrather good quality), derived from the Zlaté Horý goldmines, near Ostrava in the Czech Republic, is used toproduce both Zlatá Vodá (Golden Water) mineral wa-ter and also carbonated soda waters.

4. In some cases, coal mine waters (particularly relativelyuncontaminated pumped mine waters) have repre-sented an important source of baseflow to otherwiseoverused watercourses. Such mine waters will often befree of contamination by nutrient parameters such asnitrate and phosphate, so characteristic of many ef-fluent- or agriculture-impacted surface waters or shal-low groundwaters. For example, when pumping ofmine water from Strafford colliery into the sewage-im-pacted upper reaches of the River Dove, near Barnsleyin Yorkshire, was stopped, action groups actually cam-paigned for the resumption of minewater pumping tosupport the amenity value of the watercourse.

5. In the East of England Coalfield, hypersaline brinesoften occur at depth as formation water in the CoalMeasures strata and are encountered in deep coalmines. In the Tyneside area these were pumped indus-trially for their salt content. The waters formed the ba-sis of alkali production via the LeBlanc process atFriars Goose Pumping Station from 1827 to 1927.Many of these brines were also deficient in sulphate(due to their highly reducing nature) and hence couldcontain large quantities of Ba. Industrial quantities ofbarium sulphate were extracted from pumped minewaters at Backworth Eccles colliery up until 1978.

The bad: acid mine drainage

UK examplesThe British Coal Measures consist of a largely Upper Car-boniferous sequence of deltaic or swamp-depositedshales, mudstones, sandstones and coals. The Coal Meas-ures have typically been mined for coal at several levels,resulting in an “anthropogenic aquifer”, a network ofhigh transmissivity pathways stretching over long dis-tances, sometimes across several counties, as in the EastMidlands (Banks and others 1996b). Evidence suggeststhat, if pumped, the workings are responsible for induc-ing high vertical head gradients in the overlying strata(Banks and others 1997). Additionally, the impact ofmankind’s mining activities is seen in the widespread oc-currence of ferruginous (iron-rich) mine water discharges(Younger 1993, 1995a, 1995b; Robins 1990; Banks andothers 1996b).

The Great Northern Coalfield of Durham and Northum-berland was once the world’s most prolific source of coal.Over the last three decades, however, exploitation of thecoalfield has contracted almost to zero. In the western-most and southernmost areas of the exposed coalfield inCounty Durham, ferruginous mine water discharges fromabandoned deep mines first emerged in the late 1970s,with marked impacts on some of the tributaries of theRiver Wear. Twenty years on, more than 20 discrete out-falls still emit polluted waters, typically with total iron inthe range 2–40 mg/l, and with pH usually between 5 and6.5 (Younger and Bradley 1994, Younger 1995b).In the counties of Tyne and Wear and Northumberland,cessation of dewatering in the last decade has led to ris-ing mine-water levels, and polluting discharges to surfacewaters are anticipated early in the next century (Younger1995c). There are as yet no contingency plans for com-batting this pollution and consequently serious degrada-tion of river quality is anticipated in the Tyne, Blyth andWansbeck catchments. The peculiar legal situation pre-vailing until very recently in Britain, whereby mine-waterdischarges from abandoned mines were specifically ex-empted under the Water Resources Act of 1991 from pol-lution legislation, has left a regulatory vacuum in whichno specific party has accepted full responsibility for tack-ling mine-water-related problems.

Copper and zinc mining, NorwayThe Caledonian mountain chain and, in particular, thebase metal mineralisation province of Trøndelag (mid-Norway) hosts Norway’s richest strata-bound sulphidebodies of Fe, Cu, S and Zn, and also some of Europe’sworst mine drainage-related pollution (Iversen and Joha-nessen 1984; Fig. 4). The region also boasts some of thecountry’s finest salmon fishing rivers and the two activi-ties have not been able to avoid a degree of conflict. In-deed, the Orkla, Gaula, Orva, Hitter, upper Glomma,Stallvika, Skorovatn/Grøndal and Bleikvass Rivers haveall been classified by the State Pollution Prevention Agen-cy as being “significantly” or “highly” polluted (grade 3or 4) as a result of mining activities releasing Cu, Zn, Fe,Pb and Cd (Schartau 1992).The metals are derived from natural exposures of sul-phide minerals, mine drainage, tailings, spoil and slag de-posits, and processing works. In areas where mining hasa long history, such as the Røros area (commercial min-ing for copper since 1644), the main source of pollutionis usually subaerial mine-waste tips, with a high contentof sulphide minerals due to inefficient extraction prac-tices. Later, towards the end of the 1960s, the practice ofunderwater disposal of mining tailings from selective flo-tation became widespread (Arnesen and Iversen 1993),with the aim of limiting access of oxygen to the waste.Seven mines have practised underwater disposal of tail-ings in the recent past; at Folldal (Hjerkinn), Grong, Løk-ken, Røros, Skorovas and Sulitjelma, with sulphide con-tents in the tailings ranging from 4% to 45%.The mines have typically been worked historically (somesince the seventeenth or eighteenth century) for Cu or

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Fig. 4Map of Norway, showing locations of selected Norwegian metalsulphide mines

Zn. More recently, FeS2 (for sulphuric acid production)has been the economic primus motor in many mines,particularly in the late nineteenth and early twentiethcenturies, although falls in S prices in the 1970s caused apartial reversion to metal production. Many of the mineshave become uneconomical over the last two decadesand, despite determined government subsidy to sustainemployment, have been gradually closed down.

Skorovas mineThe Skorovas mine commenced activity in 1952/1953with the ore containing initially 3–4% Cu. Later the com-position of the ore stabilised at 45% S, 1.3% Cu, 1.7% Znand 0.05% As (Arnesen and Iversen 1995). The sulphideore was initially worked for S, but a fall in S prices led toCu and Zn being the primary products since 1976. Themine closed in 1984, and the environmental consequencesof Skorovas’s mining history are concisely summarisedby Kopperstad (1992). Up until November 1975, miningwaste from ore separation was deposited in and aroundthe shores of Dausjø Lake, whereas, after this date, un-derwater disposal in the lake of tailings from selectiveflotation took place. In total, some 200,000 tonnes ofwaste were produced, containing around 30% S (Arnesenand Iversen 1995).The mine area drains to two catchments. The mine itselfdrains into the Stallvik River catchment, whereas the run-off from the mine waste tips and the natural outcrop ofthe ore drains via the Dausjøbekk Stream and DausjøLake into the Skorovas and Grøndal Rivers. The former

catchment has received some 8–10 tonnes Cu and 17–18 tonnes Zn per year, while the latter has receivedslightly more. Of the metal run-off to the Skorovas catch-ment, some 10% is thought to be due to natural weather-ing of the ore outcrop. Indeed, the name Dausjø (DeadLake) is perhaps indicative of long-standing natural “con-tamination” in the Skorovas catchment. Mining-inducedcontamination is thought responsible for the lower Stall-vik River being empty of fish. The mining history of thearea has left a spectrum of contaminant sources for to-day’s environmental geologists to tackle, including themine itself, the mine waste tips and the contaminated de-posits lying at the bottom of Dausjø Lake.

Killingdal mineThe two ores at Killingdal, the North and South Ores,were found in 1674 and 1791, respectively (Table 3). Thegenesis and mineralogy of the ores are described by Rui(1973). Working was only sporadic prior to 1850. Themine’s production was never large; by 1984, 3 milliontonnes of ore had been extracted, but the mine wasclosed in 1986. The mine is over 1 km deep, consisting oftwo shafts, the Killingdal shaft (along the ore body) andthe Bjørgan shaft (a transport shaft constructed, in 1965,at right angles to the ore). Prior to the 1980s, highly Cu,Zn, Fe and SO4

2– contaminated run-off from spoil heapsat the top of the Killingdal shaft was allowed to flow intostreams draining into the Gaula River (Iversen and Jo-hannessen 1984). The mine was relatively dry, being con-structed in low-permeability phyllites, and no deeppumping was necessary during working. The smallamounts of seepage were removed partly by shallowpumping, while a certain amount (some authorities sug-gest 6 l/min, although the real figure is likely to be less)was removed by evaporation caused by elevated tempera-tures and extremely powerful air currents through themine Tentative water fluxes prior to closure are shown inFig. 5. Following closure of the mine, the spoil heapswere, in 1991, collected around the top of the Killingdalshaft and covered by low-permeability till. This was cap-ped with coarse stones to hinder erosion. Drainage waterfrom the tips was collected and diverted into the Killing-dal shaft. It was anticipated that the amount of waterdraining from the shaft to the mine would be sufficientlysmall to be evaporated away within the mine system (Ar-nesen and Iversen 1995). Unfortunately, due to errors inestimating the mine’s water balance, the till capping didnot exclude sufficient water from the shaft area and wa-ter was found to be running into the mine at a rate of40 l/min (Fig. 5).The rate at which the mine was filling up was only un-covered when the Norwegian government authorities de-cided that a deep mine like Killingdal would be ideal fordisposal of low- and intermediate-level radioactive waste.The team chosen to assess this proposal visited the mineto report that the mine was filling with water with a pHof 2.5, over 10,000 mg/l SO4

2– and an impressive array ofheavy metals, at a rate of some 13 m per month (NOTE-BY 1992). The quality of the water draining down the Kil-

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Table 3Composition of Norwegian Cu/Zn sulphide ores

Killingdal (North Ore)a Killingdal (Main Ore)a Skorovas (late production)b Løkken Orec

S (%)FeSiO2

CuZnMnCoPb

42.1434.15

1.0310.61

0.40

48.1540.99

1.905.89

45

1.31.7

41.437.513.7

2.21.90.070.070.02

Ni ppm (g/tonne)SeAsCdAgAu

1200 500

8050401016

0.2

a after Rui (1973)b after Arnesen and Iversen (1995)

c after Arnesen and others (1994)

Fig. 5A cross section throughKillingdal mine, illustratingwater fluxes before and afterremedial action on spoil tipsin 1991 (after Banks 1994).The points marked 1–5represent the samplinglocations on which Fig. 6 isbased (mODpmetres abovesea level)

lingdal shaft was seen to decline with increasing depth inthe mine (Banks 1993, 1994), due to either (or both) fur-ther oxidation of ore minerals in the mine or concentra-tion increase due to evaporation (Fig. 6). By June 1994,the Central Station level was submerged and air circula-tion (and hence evaporation) within the mine was cur-tailed (Arnesen and Iversen 1995). Overflow at a rate ofsome 40 l/min from the Bjørgan shaft was predicted with-in 40 years. Even if the drainage water were to be div-erted away from the Killingdal shaft, overflow would oc-cur within some 400 years, though this rate of overflowwould be unlikely to represent a major problem. Theplans for radioactive waste disposal appear to have beendropped (Henriksen and others 1992), but mine watercontinues to accumulate, and a new solution to the run-off from the spoil tips at the mouth of this shaft may ev-entually have to be found.

Løkken mine, NorwayThe Løkken area consists of a complex of interconnectedmines in a westwards dipping FeS2–CuFeS2–ZnS ore body(Table 3) in greenstone bedrock. The mines are of vary-ing age, active from 1664 to 1987 (Fig. 7) and are all nowclosed. The Løkken mine system is some 4 km long, con-sisting of four main shafts, the deepest being some1,000 m deep. Until 1952, the highly contaminated,pumped mine water was discharged to the Raubekken(Red Stream), a tributary of the River Orkla. This had amajor impact on salmon fishing in the Orkla and from1952, the mine water was piped 30 km to the town of Or-kanger, at the mouth of the Orkla in Orklafjord, wherethe water was chemically treated at the Thamshavnsmelter. After 1962, however, treatment ceased and theraw mine water was discharged to the fjord. The deepAstrup shaft was opened in 1972. The water quantity

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Fig. 6Variation with depth ofhydrochemistry of free draining waterrunning into mine along Killingdalshaft (after Banks 1994). Samples takenAutumn 1992; for sampling locations,see Fig. 5 (mODpmetres above sealevel)

Table 4Composition of a variety of mine waters from Norwegian sulphide mines

MineMain ore

Ballangena

(Ni/Olivine)Kjølib

(Cu/Zn)Løkkenc

(Cu/Zn)Løkkend,Astrup Shaft(Cu/Zn)

Stowartz,Rørose

(Cu/Zn)

Skorovasf

(Cu/Zn)Flåt, Evjeg

(Ni/Cu)Killingdalh

(Cu/Zn)

pHCond. (mS/cm)

7.22457

2.722158

2.3 2.71 6.28445

2.365922

3.68267

2.7

SO42P (mg/l) 93 1564 1671 7291 68 26 500

Fe (mg/l)Al (mg/l)Cu (mg/l)Zn (mg/l)Ni (mg/l)

5.9

441

248

25 5001080

1529

530 000870 000

21020.1

35 600131 000

5.06

2003300

2284

129 400256 500

2.58776151

2150

3680885

574 0005640 000

Cd (mg/l)As (mg/l) 192

404 100 519 12 800280

a Mine water after sedimentation basin, average 1993, Arnesenand Iversen (1995)b Collected run-off from Kjøli mine, average for 1984, Arnesenand Iversen (1995)c Mine water, average for period 1961–1975, Arnesen and Iver-sen (1995)d Average for water samples from the Astrup shaft, March1988–October 1989, Arnesen and others (1994); Arnesen andIversen (1995)

e Mine water, average for 1990/1991, Arnesen and Iversen (1995)f Mine water from Gråbergstollen, average for 1990, Arnesenand Iversen (1995)g Mine water from Gruvebekken, Arnesen and Iversen (1995)h Mine water from base of Killingdal mine, after Noteby (1992)and Banks (1994)

from this part of the mine was pumped into the Wallen-berg complex but comprised only some 5–10% of the to-tal quantity pumped from the latter (Arnesen and Iversen1995). Up until 1983, the water pumped from the minecomplex was the largest single emission of Cu and Zn inNorway. The water contained 0.5–1 g/l Cu and a similarquantity of Zn, and was discharged to Orklafjord at some500,000–600,000 m3/year. In 1983, the old part of themine around Wallenberg was closed and sealed off fromthe deeper Astrup complex with concrete plugs. The Wal-lenberg complex was allowed to fill up and water fromthe Astrup complex was also pumped at some 55,000 m3/year into the Wallenberg mine until 1987, when the As-trup mine was closed and also started to fill up.In 1989, when the Wallenberg mine was still filling, themajor source of mining-related pollution was assessed bythe Norwegian Institute for Water Research (NIVA) as

being the numerous old spoil tips along the Løkken val-ley sides, which were responsible for 90% of the total Curun-off. In 1992, the Wallenberg shaft had filled andpumping had to be recommenced. A solution to run-offfrom both the spoil tips and the mine itself was thus re-quired (Arnesen 1993a).

The ugly: extreme visual impacts

Wheal Jane, Cornwall, UKWheal Jane, between Redruth and Truro, was one of themore modern Cornish tin mines, having only been exten-sively worked since the 1960s. However, recent workingsintersect older stopes and roadways dating backhundreds (if not thousands) of years. The host rocks are

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Fig. 7Schematic section through Løkken mine,Norway

Table 5Water quality in Wheal Jane Mine, Cornwall, following refilling of mine (after Bowen and others 1994)

Sample depth pH Cond. Cd Fe Zn Cu

(m below water level)Shallow 10 mMiddle 90 mDeep 180 m

3.72.62.5

(mS/cm)164073908440

(mg/l)0.132.64.8

(mg/l)94

18462162

(mg/l)98

13791541

(mg/l)4.5

4044

Killas shales intruded by quartz porphyry dykes whichcarry veins bearing cassiterite, chalcopyrite, pyrite, wol-framite, arsenopyrite, galena and silver. While the tin mi-nerals were always the most important ores, some of themore extensive old workings at Wheal Jane were amongthe most prolific producers of pyrite (as a by-product) inCornwall. Additionally, arsenopyrite was worked for ars-enic from the Nangiles mine (also connected to Jane).The modern workings reach depths of 450 m and wereexceptionally wet by UK standards, with routine dewater-ing around 107 l/day, and peak pumping rates up to sixtimes this amount. Recent data strongly suggest thatthese high flow rates arise from the huge extent of theassociated old workings, rather than from any unusualpermeability in the Killas host rocks (Bowen and others1994). During mining, pumped water was discharged tothe River Carnon after partial treatment. In March 1991,a UK government grant to support dewatering was with-drawn, resulting in an immediate cessation of pumpingand consequent closure of the mine (Hamilton and oth-ers 1994a).The subsequent mine water rise was monitored, both interms of level and quality. On 5 July 1991, the waterquality at three different depths was recorded (Table 5),revealing incipient stratification of the water in the mine,with the poorer water quality occurring at depth. Waterlevels in the mine rose extremely quickly (Fig. 8), with arecovery of more than 400 m in less than 8 months. On17 November 1991, the mine water overflowed into theRiver Carnon via Jane’s adit at 14.5 m above sea level.The owners had provided a treatment lagoon for theoverflow, which soon proved to be embarrassingly inade-

Fig. 8Rate of rise of water levels (mODpabove sea level) in WhealJane tin mine, shaft No. 2, following pump shutdown in March1991 (reproduced with permission from J. Wright, EnvironmentAgency, UK)

quate. The initial discharge was of water with a pH of 2.8and a total metals loading of over 5 g/l. The NationalRivers Authority took emergency action by lime dosingthe water in Jane’s adit, plugging the adit mouth andpumping water from the adit to the Wheal Jane tailingsdam. In late December 1991, pumps failed and arrange-ments began to be made for a more sustainable pump-and-treat operation at the No. 2 shaft. Before the new op-erations could commence, however, a further rise of headin the mine system of around 4 m led to a catastrophicrelease of an estimated 50!106 l of highly polluted minewater when a plug in the portal of the old Nangiles aditunexpectedly burst.

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The main consequence of the Nangiles plug failure wasrapid and spectacular discoloration of the Fal Estuary. Infull daylight, and beneath the cameras of national newshelicopters, an orange plume spread throughout the es-tuarine waters, past bayfront residences, marinas and ho-tels. It was certainly the single most spectacular pollutionincident in UK living memory. The potential damage thata repeat of such unsightly pollution could do to the localtourist industry caused justifiable alarm to residents ofthe area, and the national news coverage ensured that theUK government was quick to authorise emergency pollu-tion prevention activities by the National Rivers Author-ity (Hamilton and others 1994b). Accordingly, pumpingwas initiated from the Wheal Jane No. 2 shaft, and activetreatment commenced (aeration, lime dosing, flocculationand settlement in the tailings dam). The active treatmentsystem was rapidly optimised, and now has a typical me-tals removal efficiency in excess of 98%. In the following4 years of sustained pumping and treatment, dischargerates have varied between 5,000 and 40,000 m3/day, whileFe contents in the pumped water have fallen steadilyfrom some 1,600 mg/l to 400 mg/l, indicating an approachto a dynamic equilibrium in the system, as gradual flush-ing of the products of pyrite oxidation (which were large-ly taken into solution during groundwater rebound)reaches completion.Biological surveys of the Fal Estuary after the initialmine-water outburst surprisingly revealed little or noacute impact on marine flora and fauna. This was espe-cially good news, given that the Fal area contains one ofonly two Maerl colonies in the United Kingdom (an in-ternationally important marine habitat, resembling coralreefs, with exceptionally rich biodiversity). It thus tran-spires that the principal impact of the extreme pollutionfrom Wheal Jane was visual – the ultimate in ugly minewaters.

Living with the bad and the ugly– treatment of mine drainage

Several options exist to ameliorate the effects of acidmine drainage from disused workings:1. Continue pumping in order to retain control over

drainage pathways. This may eventually evolve into apolicy of controlled refilling of mines (e.g. DurhamCoalfield, U.K.).

2. Plug drainage adits and allow the mine to refill (e.g.Skorovas, Norway).

3. Treat drainage water (e.g. Wheal Jane, U.K.).4. Dispose of drainage water to a lower sensitivity reci-

pient: for example, seasonal storage of mine water forlater release to a recipient during periods of high flow(e.g. spring flood).

Similarly, a number of options exist to minimise the con-taminated run-off from spoil tips:1. Isolate waste tips in situ by, for example, low-permea-

bility membranes (e.g. Kjøli, Norway).

2. Move waste tips to a more suitable place of disposal,for example, underwater (e.g. Skorovas, Norway).

3. Recirculate contaminated water through mine work-ings (e.g. Løkken, Norway).

4. Collect contaminated water for treatment or re-infil-tration.

Conventional active treatmentAs the Wheal Jane case study illustrates, conventional ac-tive treatment by alkaline dosing, aeration, flocculationand settlement can provide an efficient response to minewater pollution (Best and Aikman 1983). In most treat-ment plants for sulphide mine drainage in Canada andSweden, calcium hydroxide is the preferred alkali for dos-ing (Arnesen 1993b). Such treatment processes are aero-bic, with heavy metals typically being precipitated as hy-droxides.Alternative active treatment strategies employing anae-robic processes have also been successful, as at Laisvall inSweden, where lead is precipitated as sulphide or atThamshavn in Norway where mine waters from the Løk-ken mine were treated with hydrogen sulphide producedat the Thamshavn smelting works, with copper beingprecipitated as a sulphide (Arnesen 1993b).Other treatment methods which have been used for re-moval of heavy metals from non-ferruginous waters in-clude the addition of scrap iron to galvanically reduceand precipitate copper from solution, ion-exchange tech-niques, electrochemically assisted sorption onto dead bio-mass (Hancock and others 1995), reverse osmosis (Arnes-en 1993b) and electrolysis to recover copper (under trialat Kongens mine, Røros, Norway).

Passive mine-water treatmentActive treatment may be a cost-effective long-term solu-tion to mine-water pollution at sites such as Wheal Janewhere loadings of metals and acidity are especially high.However, where discharges from abandoned mines areless extreme in their quality, the ongoing revenue costs ofactive treatment can become less attractive, and passivetreatment (primarily using constructed wetlands technol-ogy) is increasingly favoured (Hedin and others 1994).In the United States, acid drainage from abandoned me-tals mines poses such acute problems in some areas thatseveral localities have been accorded “Superfund Status”by the federal government (King 1995). One such area isColorado, where gold, silver and other metals have beenmined since 1800. The mines are now largely abandoned.Excess concentrations of Cu, Zn, Cd, Pb and As affectsome 13,000 miles of the USA’s rivers, 1,400 miles ofthese in Colorado alone (CDM unpub. report). One Su-perfund site, the Clear Creek/Central City site, lying some35 km west of Denver, covers some 400 square miles andhas been the subject of intense research on mine watertreatment. Methods assessed included conventional activetreatment, but the study concluded that “natural” or“passive” treatment of water was the only technically andeconomically feasible long-term alternative. In a jointproject between Camp Dresser and McKee (CDM) and

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Fig. 9Time weighted annual transport of sulphate, copper, zinc andiron (t/aptonnes/year) from the Kjøli mine area (after datafrom Arnesen and Iversen 1995)

the Colorado School of Mines (CDM unpub. report), a pi-lot-scale treatment plant, based on reed bed technologywas built and tested. The treatment plant consisted of a3-m!20-m!1.3-m deep concrete trough, divided intothree compartments, containing a layer of river rock, alayer of compost material or substrate planted with local-ly transplanted rushes and cattails. In later versions, baf-fles were added to increase contact with the substrate andfinally an aerobic “polishing” cell was added. Under opti-mum conditions, the plant operated very successfully, re-moving up to 99% of Cu, 98% of Zn, 94% of Pb and 86%of Fe, while increasing the pH from 3.0 to 6.5.Numerous examples of the application of constructedwetlands to drainage from abandoned coal mines in Ap-palachia have been reviewed by Hedin and others (1994),who conclude that predictable chemical reactions are re-sponsible for the observed improvements in water quali-ty, and that the most frequent cause of failure of con-structed wetlands is under-sizing.In the United Kingdom, passive treatment of mine watershas been initiated only since 1994, with the constructionof an experimental wetland system at Wheal Jane, a se-ries of wetlands in the Pelenna Valley of South Wales andat Woolley Colliery in West Yorkshire (Coal Authority/IMC, unpub. report), and a compost wetland for treat-ment of acidic, Al-rich colliery spoil drainage at QuakingHouses, County Durham (Younger 1995c). The long-termviability of such passive treatment systems does not,however, at present, seem to have been adequately ad-dressed.

Other control/remediation techniques

Continued pumpingIn some large, interconnected coalfields, the most cost-ef-fective means of dealing with mine water pollution maybe to continue regional dewatering after mining hasceased (Younger and Harbourne 1995). In Durham, mine-water levels in the extensive coal mine complex underly-ing the county are being held at bay by continued dewa-tering. Thus, the breakout of mine waters to surface reci-pients is being prevented, at least as long as central gov-ernment continues to provide funding for this activity tothe Coal Authority. Cessation of dewatering in this areawould probably result in major mine-water pollution inthe River Wear catchment. Elevated Fe levels wouldprobably lead to either the closure or expensive majorupgrade of the 26!106 l/day public water supply worksat Lumley, based on the River Wear. In West Yorkshire,the Coal Authority operate a similar scheme at the form-er Woolley Colliery (Coal Authority/IMC, unpub. report).Pumping at this shaft (with associated passive settlementand wetland treatment) ensures that uncontrolled surfacedischarges of highly ferruginous mine water will not oc-cur, and also prevents flooding of underground workingsat the National Museum of Mining at Capstone. A furtherscheme along similar lines is planned for the Dysart-Lev-en Coalfield, Fife, Scotland (Younger and others 1995),where currently rising mine waters are being monitored

until they reach a pre-set threshold some tens of metresbelow sea level. When this level is reached, a pump-and-treat scheme will be initiated to prevent polluting dis-charges in the valleys of the Rivers Ore and Leven.

Isolation – Kjøli mine, NorwayKjøli mine is another Cu and Zn mine in the vicinity ofKillingdal which contributed heavily to pollution in theRiver Gaula. One million tonnes of ore were mined be-tween its opening in 1766 and its abandonment in 1941.Some 200,000 tonnes (or 80,000 m3) of spoil have beendumped in tips around the mine area and these havebeen responsible for a yearly run-off of 4.2 tonnes Cu,0.24 tonnes Zn, 38.3 tonnes Fe and 198 tonnes sulphate(Østmoe 1991; Iversen 1992). In 1981, the tips were col-lected in one area and were limed with 100 tonnes of cal-cium hydroxide (the quantity being that calculated toneutralize the run-off for 1 year). This failed to provideany long-term improvement in run-off. Indeed, followingliming the situation appeared to deteriorate (Fig. 9). Thisis likely to have been due to the physical disturbance andhence oxygenation of the spoil tips. After a careful as-sessment, isolation of the spoil heaps (to hinder access byoxygen and water) was found to be the most promisingcourse of remedial action. The isolation, which took placein 1989, employed spoil tip capping using syntheticgeomembranes overlain by a natural till cover. The op-tion of using till alone was not considered to be suffi-ciently impermeable to hinder access of water. Also in1989, the mine adit was plugged by concrete. The waterlevel within the mine rose some 17 m but, by 1990, a newoverflow had occurred at a higher level (Arnesen and Iv-ersen 1995). Nevertheless, the remedial action appears tohave been largely successful. Although the yearly trans-port of Cu and Zn from the mine area is still around600 kg/year Cu and 90 kg/year Zn (1992 values; Arnesenand Iversen 1995), these still represent a significant de-cline in Cu and Zn compared with 1985.

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Fig. 10Variation in water chemistrywith depth in the water-filledWallenburg shaft, Løkken,during filling of the mine(after data from Arnesen1993b)

Recirculation – Løkken mine, NorwayIn 1989, NIVA assessed the major part of the Cu contam-ination at Løkken as being generated by the numerousspoil heaps. The immediate recipient of the contamina-tion is the highly polluted Raubekk Stream and, ultimate-ly, the Orkla River. Several forms of action have beentaken to reduce the release of heavy metals to the envi-ronment, but the most novel has been the full-scale at-tempt at recirculating drainage water from the spoil tipsthrough the now water-filled mine complex (KommunalTeknikk 1994; Lundberg 1994), using the 6!106 m3 ofmine space as a giant copper adsorption basin.NIVA carefully monitored the quality of mine water inthe Wallenberg mine during its filling (Fig. 10). The waterdeveloped a clear stratification, with relatively uncontam-inated “surface water”, derived from run-off throughmine openings and open fractures, lying over deeperhighly contaminated mine water. Although sulphate levelswere as high as they had ever been in the deep mine wa-ter, pH had increased relative to the original mine water,Cd had decreased and the concentrations of Cu had re-duced by up to 99.9%, to less than 1 mg/l in some cases.Zinc, conversely, showed elevated concentrations, of upto 4 g/l at the greatest depths (Arnesen and others 1994;Arnesen and Iversen 1995). The conditions did not ap-pear suitable for sulphate reduction, and it was thus sug-gested that Cu adsorption and metal exchange for Zn wastaking place on the surface of sulphide minerals, and thatthis could be used to treat spoil tip waters by circulationthrough the mine. As Zn is considered to be a lesser en-vironmental toxin than Cu, the overall effect is an im-provement in quality.The treatment system commenced between April and Au-gust 1992. Spoil tip drainage is introduced at a rate of600,000 m3/year to the mine complex at the tourist mineGamlegruva (Old Mine) and, after a journey of approxi-mately 1.2 km, taking several years, is pumped up fromthe newer Wallenberg shaft (Kommunal Teknikk 1994).The workings contain a considerable quantity of basic li-thologies, such as greenstone and calcite, which contri-bute towards raising the water’s pH. The other parametermost radically affected is Cu. Whereas the original watercontains 75–200 mg/l Cu, the pumped effluent water con-tains a mere 0.5–10 mg/l Cu and around 20 mg/l Zn (Ar-

nesen and others 1994; Arnesen and Iversen 1995). It isperhaps a little too early to judge the results, however, asit is not certain that full breakthrough of influent waterhas been achieved in the Wallenberg shaft. There are alsosome doubts as to the long-term sustainability of the ef-fect, as exposed pyrite is eventually expected to becomesaturated with Cu, but the process is imaginative, prom-ising and deserves further research.Longer term plans may involve attempts to promote sul-phate reduction, and hence metal sulphide precipitation,within the mine. This might be achieved by adding anorganic substrate to promote the growth of sulphate-re-ducing bacteria. Laboratory-scale research is being car-ried out on the use of surplus dairy whey, cow manureand brewing yeast as “food” for bacterial communities(Lundberg 1994; E. Iversen, pers. comm.; Arnesen andothers 1994).

Hydraulic control – Killingdal mine, NorwayOne option which is currently being discussed (Henriks-en and others 1992) is the possibility of “plugging” theBjørgan shaft with a concrete block, thus preventingmine water overflow. In order for this to work, however,several conditions must be satisfied:1. No additional adits should exist. In older mines, accu-

rate plans may not exist and it may not be possible tolocate all old adits and workings.

2. The plug must be engineered to withstand the 200 mhead difference between the Bjørgan and Killingdalshafts and be constructed of a material capable ofwithstanding the potentially acid and sulphate-rich na-ture of the drainage.

3. The rock should be of sufficiently low permeabilitythat mine water may not penetrate natural groundwat-er flow paths (e.g. fractures) to emerge as contami-nated springs.

4. The question of how to deal with the continuing pro-duction of acid drainage by the spoil heaps must betackled.

Alternatively, the spoil tips could be re-capped to minim-ise the amount of leachate entering the mine. The re-duced flux of water could be allowed to flow through theworkings to exit at a low rate from the Bjørgan shaft. Itmight be expected that the quality of any water which is

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Fig. 11Water quality at the exit from Dausjø Lake, Skorovas mine(after data from Arnesen 1993b)

allowed to exit from the Bjørgan shaft will be superior tothat entering the mine, due to exchange or precipitationof heavy metals taking place at depth in the mine (seesection on Recirculation – Løkken above). A first flush ofhighly polluting water may yield problems.

Combined solutions – Skorovas mine, NorwayImprovements to the Dausjø catchment commenced in1976, with the underwater disposal of tailings from selec-tive flotation. Lime is often incorporated in the flotationprocedure and tailings have an elevated pH (at least tem-porarily until pyrite oxidation exhausts the buffering ca-pacity of the lime). The impact of the disposal of the al-kaline tailings on heavy metals and pH in Dausjøbekk isseen in Fig. 11. This has been responsible for re-estab-lishing fish life in the Grøndal River.As discussed above, the Skorovas mining area containsseveral sources of contamination, all of which have hadto be dealt with in a clean-up programme (Kopperstad1992). Concerted environmental rehabilitation began in1984 when the contaminated deposits lying in DausjøLake were capped with relatively inert ground rock waste.Acid mine drainage from spoil heaps was diverted pastDausjø to encourage a rise in pH and thus hinder mobil-isation of metals from the contaminated sediments. Inaddition, a liming station was established on DausjøStream to raise the pH of the water.These remedial works were only partially successful. Tocontrol contamination in the watercourse, 400 tonnes oflime per year were required, while the metal hydroxidesprecipitating out downstream of the liming station werecausing a significant problem in themselves. In 1987–1988, it was therefore decided to negate the need for lim-ing by removing the major source of the acid drainagecontamination to the Skorovas catchment, namely thespoil tips. The favoured location for their removal wasthe bed of the 20 m deep Dausjø Lake. The lake was par-tially emptied and the spoil emplaced and covered with

till material. The lake was then refilled, and it was hopedthat metals would be immobilised in the low-throughflowand probably chemically reducing environment of thelake bed. The lake was refilled in 1991 and results so farappear promising. Heavy metals in the catchment are be-ing held at satisfactory levels without the need for liming(Kopperstad 1992; Arnesen 1993b).As regards the mine itself, the chosen solution was toseal off, in 1990, all the lower adits and passageways withconcrete plugs and allow the mine to refill. Currently,now that the mine drainage no longer discharges, im-provements are noted in river water quality and fish havereturned to Stallvik River. The mine is expected to fill upand overflow from an upper adit in 1995–1997. It is an-ticipated that reducing conditions will be reestablished atthe deeper levels of the mine, hindering further sulphideoxidation and that the rate of throughflow will be smallcompared with the original active mining condition. It isthus anticipated that, although some continued dischargeof acid mine drainage from the mine will occur, the re-filling of the mine will represent a substantial environ-mental improvement.

Innovative technologiesSeveral authors have proposed the use of imaginativetechniques, which probably require considerable furtherresearch before they can be applied reliably “in-mine”.Shelp and others (1995) have shown at laboratory scalethat the use of sacrificial anodes can suppress the oxida-tion of metal sulphide ores. It is also speculated whether,in nations which are energy-rich, such as Norway, a po-tential difference could be applied to the ore body, op-posing the natural galvanic couple generated by ore oxi-dation (Logn and Bølviken 1974; Bølviken and Logn1975) and thus suppressing the oxidation reaction.The oxidation of sulphides is typically bacterially cata-lysed. Rastogi (1994) has demonstrated that the applica-tion of bactericides can hinder pyrite oxidation in spoiltips. The application of bactericides in deep mines hasyet to be demonstrated, however.The uses of bacterial bioreactors to treat heavy-metal-containing industrial wastewaters has been suggested byGadd (1990), although such techniques could be equallyapplicable to mine waters (Banks 1992). These reactorsmay be based on heavy metal-immobilising bacteria (e.g.Gallionella) or on sulphate-reducing bacteria.

Conclusion

This paper has reviewed a variety of circumstances inwhich mine-water drainage can arise and the similar vari-ety in mine-water chemistry that has been observed.Mine waters from apparently similar mine types can behighly acidic or alkaline, depending on the complex in-terplay of hydraulic, chemical and biological processesoccurring. Mine waters may be valuable in terms of highquality baseflow or drinking-water sources; they may be

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regarded as sources of raw materials; they may be highlyecotoxic and be responsible for devastating watercoursesor they may be highly visible, defacing our estuaries andrivers. Unfortunately it is often only the “ugly” mine-wa-ter occurrences that wake the attention of the public andpoliticians. The benevolent mine waters may not be ade-quately exploited, nor the dangers from the less-visible,but highly ecotoxic, mine waters be adequately appre-ciated.A quantitative understanding of the processes governingacid mine drainage remains elusive. These are particular-ly complex to tackle, not least because the chemistry ofmine waters is seldom controlled purely by thermody-namic equilibria between mineral and dissolved phases. Asatisfactory quantitative model needs to address the fol-lowing issues:1. Three-phase interaction between mineral, gas and wa-

ter phases, often on a seasonally fluctuating basis (e.g.in the zone of water-table fluctuation).

2. The importance of kinetic factors; the rates of reactionand consumption of carbonate, sulphide and silicatephases (Strömberg and Banwart 1994, Strömberg andothers 1994), diffusion-limited steps.

3. The role of bacteria as biocatalysts in sulphide oxida-tion.Nevertheless, a thorough understanding of the mecha-nisms and controls on the geochemistry of mine-waterevolution remains a priority task. There is an acuteneed to be able to predict the quality of pumped minewater when a new mine is commissioned or of drai-nage water when a mine is abandoned. Additionally,the chemical nature of a mine water will determinewhich treatment techniques are likely to be successful.

Acknowledgements The authors wish to thank Scott WilsonCDM and Mike Cooper of the University of Sheffield for assist-ance in producing the figures. Paul Younger and David Banksalso wish to thank Barney Banwart of Bradford University, UK,for very fruitful discussions on the theme of mine drainage hy-drochemistry.

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YOUNGER_Mine Water Chemistry. the Good, The Bad and the Ugly

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