2004:194 CIV
M A S T E R ' S T H E S I S
Groundwater Geochemistry in Remediated Sulfide-Rich Tailings
- Kristineberg, Northern Sweden
Distribution of dissolved organic carbon and metals
Helena Skoglund
Luleå University of Technology MSc Programmes in Engineering
Department of Chemical Engineering and Geosciences
Groundwater geochemistry in remediated sulfide- rich tailings – Kristineberg, Northern Sweden
Distribution of dissolved organic carbon and metals
Helena Skoglund
Department of Chemical Engineering and Geosciences Division of Applied Geology
Luleå University of Technology SE-97187 Luleå, Sweden
Abstract
The aim of remediation of impoundments with sulfide-rich tailings is to prevent oxygen transport to the tailings and thereby inhibit the oxidation of sulfides. Over time, remediation usually leads to decreasing contents of elements such as metals and sulfur in the groundwater of the impoundments. This has been the progress in most parts of impoundment 1 in
Kristineberg mining area in northern Sweden. The studied impoundment was exposed to weathering during 50 years before remediation. A protective cover of one meter of till was applied on the tailings, and by sealing off ditches the groundwater level was raised. Fresh groundwater flowing in from the surrounding hills has resulted in a wash out of released elements from most parts of the impoundment, and subsequently the concentrations of metals in the groundwater has decreased. The geochemistry in the northwestern part though, shows a rather different behaviour than the rest of the impoundment, with much higher concentrations of metals and sulfur in the groundwater.
The aim of this study is to find an explanation to the high content of metals in this part. The study evaluates data from sampling during five years, 1998 – 2002. The data consists of analyses of the groundwater, analyses of the solid tailings and measures of groundwater levels. The evaluation of the groundwater data in this study focuses on four groundwater pipes, two installed in the northwestern part (pipes H and I) and two pipes installed in the middle of the impoundment (pipes F and G). The pipes have been installed on different depths in the tailings.
The result shows that the concentrations of metals and sulfur in the groundwater are highest in the deeper pipe H in the northwestern part of the impoundment than in any other pipe. During year 2002 the average value for Fe in pipe H was 17 350 mg/l, while in pipe F the average value was 156.9 mg/l. The average value for Zn was 2 510 000 µg/l in pipe H, in pipe F the average value for Zn was 882 µg/l.
In pipe I, F and G, the pH of the groundwater has increased after the remediation, while the pH in pipe H has been stagnant around 4.5. The content of dissolved organic carbon (DOC) in the groundwater is higher in pipe H than in the other pipes. Peat and till underlie the tailings in the northwestern part and could explain the high concentrations of DOC in the
groundwater.
Depth profiles of the impoundment indicate that the tailings in the northwestern part of impoundment 1 are located in a depression. Pipe H is installed in the deeper parts of these tailings where the groundwater flow can be very slow or even stagnant. This results in that polluted water is not transported away, and the expected washout of dissolved metals will not occur. At depth where pipe H is situated, the groundwater will continue to have high
concentrations of elements for a long time.
The intention of remediation with a raised groundwater level was to prevent further diffusion of oxygen into the tailings. Studies of groundwater levels though, show that in some places the groundwater table does not reach above the tailings. An oxidation of sulfides could therefore still occur at these locations.
CONTENTS
1 INTRODUCTION ... 2
2 SITE DESCRIPTION ... 4
3 BACKGROUND... 7
4 THEORY... 8
4.1 SULFIDE OXIDATION... 8
4.2 MINERAL DISSOLUTION / BUFFERING PROCESSES... 9
4.3 ATTENUATION PROCESSES... 10
4.4 MICROBIAL PROCESSES... 11
4.4.1 Anaerobic microbial processes ... 11
4.5 INFLUENCE OF ORGANIC MATTER... 12
5 METHOD... 13
5.1 SAMPLING... 13
5.2 IN SITU ANALYSIS... 14
5.3 ANALYTICAL METHODS... 15
6 RESULTS... 16
6.1 DISSOLVED ELEMENTS IN GROUNDWATER... 16
6.2 DEPTH PROFILES... 23
6.3 GROUNDWATER LEVEL... 24
7 DISCUSSION... 28
7.1 DISSOLVED ELEMENTS IN GROUNDWATER... 28
7.2 DEPTH PROFILES... 29
7.3 GROUNDWATER LEVEL... 30
8 CONCLUSIONS... 32
9 ACKNOWLEDGEMENTS ... 33
10 REFERENCES ... 34 APPENDICES
1 Introduction
Mining is an important industry in Sweden, where it has occurred for more than 1000 years.
Sweden is now one of the largest producers of metals in Europe, with Europe’s largest iron and copper mines (the iron mine in Kiruna and Aitik copper mine in Gällivare). However, mining also produces large quantities of tailings and waste rock. Tailings are produced during mineral extraction and processing when sulfide ores are crushed and milled. Tailings often contain large amounts of different sulfides, such as pyrite, pyrrhotite and other minerals.
When the tailings are exposed to atmosphere it causes oxidation of the sulfides.
The interaction between hydrological and geochemical processes in the tailings is illustrated in figure 1.1. Sulfide-oxidation reactions occur in the vadose zone of the impoundment where oxygen diffusion through the partly water-saturated tailings is rapid. Oxidation of iron
sulfides is a source of acid mine drainage (AMD) with subsequent release of metals from sulfides, such as sphalerite (ZnS), chalcopyrite (CuFeS2) and galena (PbS), and other minerals. AMD is characterized by a low pH, high sulfate content, and the presence of dissolved metals such as iron, zink and copper.
Figure 1.1. Hydrological and geochemical processes controlling metal transport in mine wastes, (Blowes and Ptacek, 1994).
The initial oxidation of iron sulfides needs oxygen and water (Stumm and Morgan, 1996).
Remediation of mining waste is therefore usually directed towards limiting oxygen transport to the waste by applying soil cover or water cover over the tailings. One common method is to construct water-remaining dams to keep the tailings water saturated. Another method is to add a sealing layer over the tailings and a protective cover of till on top of the sealing layer.
In this report the groundwater in an impoundment with sulfide bearing tailings at Kristineberg
years before remediation. The impoundment was then isolated from the atmosphere to prevent further oxidation by a combination of methods. One part of the impoundment was covered with a sealing layer of compacted clayey till overlain by a protective cover of till. The other part was remediated by a cover of unclassified till, and by raising the groundwater level. A raised groundwater level would result in dissolution of secondary minerals, such as oxides and hydroxides, and a wash out of weathering products that have been retained in the tailings.
Fresh groundwater from the surrounding hills would lead to a flush out of released elements from the impoundment and the concentrations were expected to decrease after a few years.
The raised groundwater level would also keep the tailings water saturated and limit oxygen diffusion.
Various geochemical studies have been made at the remediated impoundment. Geochemical investigations were done in 1998 by characterizing the tailings mineralogy and the chemical composition of both tailings and pore water (Holmström et al., 2001). Geochemistry of the groundwater was evaluated during the years 1998 until 2001 to study the effects of applying dry cover with a combination of raising groundwater level above tailings (Corrège, 2001).
The groundwater in the impoundment generally shows a clear decrease over time in the concentrations of most of the metals. Since 1998 the pH has increased and the redox decreased. The geochemistry in the northwestern part of the impoundment though, shows a rather different behaviour than the rest of the impoundment. The groundwater in this part has higher concentrations of metals and metalloids, such as Fe, S, Mg, Al, Mn and Zn, and with a pH remaining low. The aim of the present study is to give an explanation to the high metal content in the groundwater in the northwestern part of the impoundment.
2 Site description
The Kristineberg mining area is located in the western part of the Skellefte ore district, approximately 175 km southwest of Luleå (figure 2.1). The bedrock consists of 1.9 Ga massive pyrite rich ore bodies intercalated within volcanic rocks and overlain by sedimentary rocks. The largest ore body in the area is the Kristineberg Zn-Cu deposit, which was
discovered as early as 1918. The most common ore minerals present in the Kristineberg mine is pyrite, chalcopyrite, sphalerite and rutile (du Rietz, 1951; Malmström et al., 2001). Gangue minerals present are mostly quartz, sericite, chlorite and talc (Ekstav and Qvarfort, 1989).
Figure 2.1. The location of the Kristineberg mining area is situated in northern Sweden. The mining area consists of five impoundments with sulfide-rich tailings from different mines.
Mining started in 1940 by Boliden AB and is still in progress. A concentrator plant was also opened which was supplied with ores from several mines in the area, so the impoundments contain a mixture of different tailings. The unoxidised tailings consist of about 10 to 30 % sulfide minerals such as pyrite (FeS2), pyrrhotite (Fe1-xS), sphalerite (ZnS), chalcopyrite (CuFeS2) and galena (PbS) (Holmström et al., 2001). Pyrite is by far the most abundant sulfide mineral.
Five tailings impoundment are located within the Kristineberg mining area (figure 2.1). In this report the groundwater geochemistry in impoundment 1 is studied. It is the oldest
impoundment and was used until the early 1950’s. Impoundment 1 is located in a valley and covers approximately 0.10 km2 (Boliden Mineral, 1995).
Figure 2.2. Impoundment 1 at Kristineberg mining area with location of monitoring points. The north-western part of impoundment 1 is studied in this report. This part is enlarged in the upper picture.
The average thickness of the tailings in impoundment 1 is about 6 to 8 meters and the tailings are underlain by peat and till. The impoundment was remediated in 1996 by a combination of methods. In the north-eastern part the tailings were covered with a 0.3 meter sealing layer and on top of that a protective layer was applied with 1.5 meter of till. This part is referred to as the zone with dry cover in figure 2.2. The other part of the impoundment is referred to as a zone with raised groundwater. In this part a protective layer of 1 meter of till was applied over the tailings. The groundwater was shallow in this area and by sealing off intercepting and drainage ditches the groundwater table could be raised. Figure 2.3 shows a schematic profile over the layers in the northwestern part of impoundment 1. For a detailed description of the remediation, see Lindvall et al. (1999).
Figure 2.3. Schematic illustration of the depth profile in the northwestern part of impoundment 1, Kristineberg, mining area.
Hydraulic conductivity for the tailings in impoundment 1 is ranging from 2·10-5 to 5·10-5 m/s (Axelsson et al., 1986). The porosity is ranging from 35 % to 45 % on samples of tailings taken from point P4 (figure 2.2) (Carlsson, 2000). The tailings were discharged into the impoundment with a water flow originating from a spigotting point located about 100 meters southwest of pipes F and G. The sedimentary processes resulted in deposition of more coarse grains in the area near the spigotting point and more fine grains further away.
3 Background
This study focuses on the northwestern part of impoundment 1 (figure 2.2), an area in which the groundwater table has been raised and the tailings has been covered with till. The purpose of raising the groundwater table was to keep the tailings water saturated to prevent further oxidation of sulfides. Before remediation the oxidation front had reached down to depths of 0.1 to 1.15 meter. The oxidation appeared to have been intense with high concentrations of dissolved metals in groundwater and pH-values as low as 2.6 (Ekstav and Qvarfort, 1989;
Holmström et al., 2001).
Buffering reactions occurred and buffering minerals was mostly Mg-bearing silicates and talc.
Below the oxidation front there was an enrichment zone due to precipitation of secondary minerals and adsorption of dissolved ions on different mineral surfaces (Corrège, 2001;
Holmström et al., 2001). After the remediation these secondarily retained elements, mainly Fe, S, Mg and Zn, have been remobilized as a result of the raised groundwater level (Corrège, 2001). A first washout effect shows increasing concentrations in the groundwater of elements secondarily retained below the oxidation front (see figure 3.1). A second washout with clean groundwater from the western till slope washes away the mobile elements and the
concentrations decrease.
Figure 3.1. Illustration of the washout effect.
Studies of the groundwater geochemistry in impoundment 1 show that the quality of the groundwater varies a lot over the impoundment (Corrège, 2001). For example the pH varies between 3.9 and 7, the concentration of Fe has a range between 30 and 19 000 mg/l and Zn- concentrations between 16 to 2 500 000 g/l. The second washout is marked by a clear decrease of elements in the groundwater due to an increased influence of fresh groundwater washing away the pollutants. The northwestern part of the impoundment though does not give any clear signs of remediation. In pipes H and I the concentrations of most elements in the groundwater are still very high, with higher dissolved amounts registered in the deeper pipe H (Corrège, 2001). The tailings in this area are underlain by till which has been affected by the tailings and the acid mine drainage (Holmström et al., 2001). Iron precipitations occur, creating a secondary acidification and a secondary buffering, possibly by micas (Corrège, 2001).
4 Theory
The rates of sulfide oxidation are dependent of several parameters such as air and water transport, temperature, pore water pH, particle size distribution, organic and mineral content, microbial processes, and secondary mineral formation. This chapter describes some of the processes that take place in the sulfide tailings.
4.1 Sulfide oxidation
Pyrite is by far the most common sulfide mineral in the Kristineberg ore bodies. The unoxidised tailings in impoundment 1 have a content of approximately 26 % pyrite (Holmström et al., 2001). Oxidation is a complex process involving several steps. Pyrite- oxidation is commonly described by the reaction given in equation 1; dissolved ferrous iron (Fe2+) is released, sulfides (S22-) oxidise to sulfates (SO42-) and acidity in the form of H+ is formed.
) ( 2 ) ( 2 ) ( )
( )
2 ( ) 7
( 2 2 2 42
2 s O aq H O l Fe aq SO aq H aq
FeS + + + + + + (1)
The formed Fe2+ may oxidise to ferric iron (Fe3+), (equation 2). The ferric iron hydrolyzes to form ferric hydroxide, releasing more acidity as given in equation 3.
) 2 (
) 1 ( )
( ) 4 (
) 1
( 2 3 2
2 aq O aq H aq Fe aq H O l
Fe + + + + + + (2)
) ( 3 ) ( ) ( ) ( 3 )
( 2 3
3 aq H O l Fe OH s H aq
Fe + + + + (3)
Equation 1, 2 and 3 can be summarized to equation 4.
) ( 4 ) ( 2 ) ( ) ( )
2 ( ) 7 4 (
) 15
( 2 2 3 42
2 s O aq H O l Fe OH s SO aq H aq
FeS + + + + + (4)
Ferric iron can also be reduced by pyrite itself (equation 5), then the sulfide in pyrite is oxidised and H+ is again released along with additional ferrous iron, which may re-enter the reaction cycle again via reaction 2. Ferric iron is a strong oxidant, such that at low
pH (< 4.5) and under sterile conditions, it is generally accepted that Fe3+ oxidises pyrite much faster than O2 (equation 1). At low pH, ferric iron is the dominant oxidant of pyrite by
reaction 5 (Alpers et al., 1994).
) ( 16 ) ( 2 ) ( 15 ) ( 8 ) ( 14 )
( 3 2 2 42
2 s Fe aq H O l Fe aq SO aq H aq
FeS + + + + + + + (5)
This reaction also proceeds more rapidly than the abiotic oxidation of dissolved Fe2+ to Fe3+
in equation 2 (Herbert Jr., 1999). For this reason, the oxidation of ferrous iron is considered the rate-limiting step in abiotic oxidation. At neutral to alkaline pH the abiotic rate of Fe2+
oxidation rises rapidly. Although, the concentrations of Fe3+ and hence the rate of pyrite oxidation decreases greatly due to the precipitation of ferric hydroxides (equation 3). Findings
(Moses et al., 1987). The major function of O2 is to oxidise ferrous iron to ferric iron (equation 2) and thus maintain the pyrite oxidation cycle (Herbert Jr., 1999). As the pH increases, pyrite oxidation becomes more dependent on the Fe-recycling reaction and less dependent on the reservoir of aqueous Fe3+ (Moses et al., 1987).
Oxidation of other sulfides may also release metals and sulfate, although no acid is formed in the primary step. The oxidation of the sulfide minerals sphalerite and galena is shown in equation 6 and 7.
) ( )
( )
( 2 )
(s O2 aq Zn2 aq SO42 aq
ZnS + + + (6)
) ( )
( )
( 2 )
(s O2 aq Pb2 aq SO42 aq
PbS + + + (7)
However, if ferric iron is the oxidant, acid is produced in the form of H+ according to equation 8.
) ( 8 ) ( )
( 8 ) ( )
( 4 ) ( 8 )
(s Fe3 aq H2O l Zn2 aq Fe2 aq SO42 aq H aq
ZnS + + + + + + + + + (8)
4.2 Mineral dissolution / buffering processes
Pyrite weathering does not always produce acid mine drainage because the pH is buffered by dissolution of other minerals, such as carbonates, silicates and oxides/hydroxides. Buffering by dissolution of carbonate-minerals can be represented by the reaction for calcite (equation 9 and 10) (Blowes and Jambor, 1990).
) ( )
( )
( )
( 2 32
3 s H aq Ca aq HCO aq
CaCO + + + + pH > 6.5 (9)
) ( )
( )
( 2 )
( 2 2 3
3 s H aq Ca aq H CO aq
CaCO + + + + pH 6.0 (10)
Equation 10 added to equation 4 gives equation11.
+
+
+ ( )
2 ) 7 4 (
) 15 ( 2
)
( 3 2 2
2 s CaCO s O aq H O l
FeS
) ( 2
) ( 2 ) ( 2 ) ( )
(OH 3 s SO42 aq Ca2 aq H2CO3 aq
Fe + + + + (11)
Subsequently, two moles of carbonates are required to neutralize the acid generated by the oxidation of one mole of pyrite.
Silicate minerals have higher buffering capacities than carbonates, but the dissolution rates of silicates are slower than that of calcite (Paktunc, 1999). The reaction between chlorite and sulfuric acid (H2SO4) shows that one mole of chlorite neutralize eight moles of sulfuric acid, equation 12.
+8 ( )
) ( )
( 8 2 4
10 3 2
5Al SiO OH s H SO aq
Mg
) ( 3
) ( 6 ) ( 8 ) ( 2 ) (
5Mg2+ aq + Al3+ aq + SO42 aq + H2O l + H4SiO4 aq (12)
Equation 12 added to equation 4 shows that one mole of chlorite neutralizes the acid generated by the oxidation of four moles of pyrite, see equation 13.
+
+
+ ( ) ( ) 15 ( ) 8 ( )
) (
4FeS2 s Mg5Al2Si3O10 OH 8 s O2 aq H2O l
) ( 3
) ( 8 ) ( 2 ) ( 5 ) ( ) (
4Fe OH 3 s + Mg aq + Al3+ aq + SO42 aq + H4SiO4 aq (13)
In conditions with low pH common secondary minerals, such as kaolinite and oxides of iron and aluminium, can become unstable and dissolve, see equation 14, 15 and16.
) ( )
( 2
) ( 2 ) ( 6 ) ( )
( 4 3 4 4 2
5 2
2Si O OH s H aq Al aq H SiO aq H O l
Al + + + + + (14)
) ( 3 ) ( )
( 3 ) ( )
(OH 3 s H aq Fe3 aq H2O l
Fe + + + + (15)
) ( 2 ) ( )
( ) ( )
(OH 3 s H aq Al3 aq H2O l
Al + + + + (16)
Dissolution of silicates and secondary minerals can explain high concentration of dissolved elements, such as Mg, Al and Si.
4.3 Attenuation processes
The discharge of metals in groundwater and into receiving waters not only depends on weathering processes, it is also affected by processes that affects the mobility of metals. The mobility depends on factors such as pH, redox, groundwater movement and the soil´s content of e.g. clay minerals and organic material. Metals from weathering can be retained in the tailings or in the water dams by precipitation of secondary minerals or by adsorption onto surfaces of mineral particles or organic material. Clay minerals and organic matter have a high affinity to adsorb metals (Ingri, 2002).
When H+ released by sulfide oxidation reaches carbonate-, hydroxide-, and other base- containing solids, a sequence of pH-buffering reactions occurs. Examples of these reactions are given in chapter 4.2. Acid-neutralization reactions result in an increase in the pH. This increase in pH is frequently accompanied by the precipitation of metal-bearing hydroxide and hydroxysulfate minerals that remove dissolved metals from the water within the
impoundment. Some examples of hydrolysis that remove metals from solution is given in equation 17-19 (Hellier, 1999).
+ ++ H OFeOOH s +H O+ H
Fe3 3 2 ( ) 2 3 (17)
+
++ H O Al OH s + H
Al3 3 2 ( )3( ) 3 (18)
+
++ H OMnO s + H
Mn4 2 2 2( ) 4 (19)
Secondary hydroxides and oxyhydroxides of Fe, Mn or Al are of particular importance when it comes to coprecipitation and adsorption processes. For example elements such as Si, Al, S, As, Ni and Cr have been observed to have a strong affinity for goethite, FeOOH (Jambor, 1994). Adsorption and coprecipitation also control the Zn and Cu concentrations, the extent of
metal removal by adsorption onto metal oxide surfaces increases with increasing pH (Blowes and Jambor, 1990).
For certain trace elements such as Cu, Pb and Zn, adsorption on sulfide mineral surfaces, e.g.
pyrite, seems to be a mechanism for the attenuation of elements from the tailings porewater (Müller et al., 2002). Acid-neutralization and mineral-precipitation reactions that occur at the location of sulfide oxidation can decrease the rate of oxidation through the formation of inhibitory mineral coatings (Blowes and Ptacek, 1994). These processes also decrease the rate of transport of H+ and dissolved metals away from the source area.
4.4 Microbial processes
Mine waste deposits are far from sterile and contain a large quantity of bacteria, including iron- and sulfide-oxidising bacteria. The autotrophic thiobacilli are able of oxidising metal sulfides to produce sulfuric acid and dissolve metals. (Gould et al., 1994). Thiobacillus ferrooxidans is the best characterised of acidophilic thiobacilli, which means that they will grow only at a low pH. It has a pH range of 1.0 to 3.5, and an optimum near 2.0 (Gould et al., 1994).
T.ferrooxidans is an autotrophic bacteria. To gain energy for the reduction of carbon dioxide (CO2 fixation) T.ferrooxidans oxidises reduced inorganic compounds, including ferrous iron and sulfides, using oxygen as an electron acceptor (Gould et al., 1994; Herbert Jr., 1999). The presence of T.ferrooxidans and other bacteria can accelerate the rate of Fe2+ and sulfide mineral oxidation by O2 by several orders of magnitude (Herbert Jr., 1999).
The primary pyrite oxidant in both abiotic and biotic systems is ferric iron rather than
molecular oxygen (Herbert Jr., 1999). Below about pH 3 the oxidation of pyrite by ferric iron (equation 5) is about ten to a hundred times faster than by oxygen (equation 1) (Ritchie, 1994). The oxidation by Fe3+ though is also dependent on the rate for the oxidation of Fe2+ to Fe3+ (equation 2). Under abiotic conditions the rate of oxidation of pyrite by ferric ion is controlled by the rate of oxidation of ferrous ion (Fe2+), which decreases rapidly with decreasing pH. Iron-oxidising bacteria though, principally T.ferrooxidans, catalyze the oxidation of ferrous iron and increase the rate of this reaction by a large factor that has been estimated at about 105 (Ritchie, 1994). The rate of pyrite oxidation proceeds about as fast as the aqueous Fe3+ can be produced from Fe2+ through microbial catalysis (Herbert Jr., 1999).
4.4.1 Anaerobic microbial processes
Sulfide oxidation, as mentioned, requires both oxygen and moisture, but there is some
evidence for the oxidation of reduced-sulfur compounds under anaerobic conditions (Gould et al., 1994).
Anaerobic microbial processes are of importance for metal mobilisation in remediated mine tailings. Besides O2, metal sulfides can be oxidised by oxidants such as Fe(III), Mn(IV) and nitrate (Shippers, 2003). Anaerobic conditions under a cover may not stop bacterially mediated sulfide oxidation. There is increasing evidence that iron-oxidising bacteria such as T.ferrooxidans, under anaerobic conditions can fix CO2 and use a variety of metabolic pathways including the oxidation of ferrous iron and sulfide minerals for the gaining of energy (Herbert Jr., 1999; Pronk and Johnson, 1992).
T.ferrooxidans are able of growing under low oxygen conditions. It has been generally accepted that T.ferrooxidans activity stops when oxygen is depleted, however, bacterial activity can continue even under extremely low oxygen tension (Herbert Jr., 1999). Studies have shown that T.ferrooxidans can couple ferric iron reduction to sulfide oxidation in the absence of oxygen (Pronk et al., 1992; Ritchie, 1994). Although T.ferrooxidans has
traditionally been considered to be an obligate aerobic organism it has been shown that it can survive anaerobically by using ferric iron as an electron acceptor (Pronk and Johnson, 1992;
Pronk et al., 1992; Ritchie, 1994).
It has been shown that T.ferrooxidans is capable of leaching a Zn-Fe sulfide in the absence of O2, provided CO2 was available (Goodman et al., 1983). Ferric iron was presumably the oxidant in this case. Studies have also shown that bacterial growth increases with CO2
concentration indicating that controlling the levels of CO2 may be almost as important as controlling the levels of oxygen (Herbert Jr., 1999).
4.5 Influence of organic matter
High amounts of Fe(III) are available in oxidized tailings. At neutral to alkaline pH, Fe(III) is precipitated as Fe(III)(hydr)oxides and are almost insoluble. In the tailings cover, or in
organic bottom sediments, organic acids and compounds may be present. Organic compounds can contribute to the dissolution of ferric oxides resulting in an increase in the concentration of Fe(III) in solution by forming aqueous Fe(III)complexes (Luther et al., 1992).
These organic complexes can react with sulfide in the unoxidised tailings. Microorganisms may couple ferric reduction with sulfide oxidation in the anoxic zones. In the presence of organic ligands, an increase in the rate of pyrite oxidation has been observed with an increase in the concentration of soluble organic Fe(III)complexes (Luther et al., 1992). At acidic pH the concentration of complexed Fe(III) is high enough to make FeS2 dissolution possible (Shippers, 2003).
When Fe(III)- and Mn(IV)(hydr)oxides are biologically dissolved in the presence of organic compounds adsorbed metals will be released. Reduction of Fe(III) and Mn(IV) are an important mechanism for the oxidation of organic compounds in aquatic environments (Lovley, 1991). Equation 20 illustrates the oxidation of glucose to CO2 under anaerobic conditions with Fe(III)as the oxidising agent (Lovley, 1991; Shippers, 2003).
+ +
+ +
+
+ Fe III H O HCO Fe H
O H
C6 12 6 24 ( ) 12 2 6 3 24 2 30 (20)
In the presence of sulfate-reducing bacteria H2S (g) is produced when organic material is oxidised, as shown in equation 21 (Stumm and Morgan, 1996).
) ( 2 ) ( 2 ) ( )
( 2 ) ( )
(
2CH2O s +SO42 aq + H+ aq H2S g + CO2 g + H2O l (21)
Some of this gas may react with reduced iron and form secondary sulfides as shown in equation 22 (Hellier, 1999).
) ( 2 ) ( )
( )
( 2
2 aq H S g FeS s H aq
Fe + + + + (22)
5 Method
The data from samplings and analyses were collected during the period August 1998 to April 2003 and some of the data has been published in other articles (Holmström et al., 2001;
Corrège et al., 2001; Corrège and Öhlander, 2001).
5.1 Sampling
Sampling of solid tailings was performed in 1998. Five profiles were drilled in impoundment 1 using a drill-rig. All profiles were extended down to the underlying peat or till. The drill cores were split into 20-cm subsamples, which were placed in polyethylene plastic bags immediately (Holmström et al., 2001). Only data from samples in profile P3 (figure 2.2) is used.
Sampling of groundwater began in August 1998 and has been performed across the
impoundment (figure 2.2) on 38 occasions until April 2003. The groundwater sampling has been performed by the use of BAT®-pipes which results in a minimal sample exposure to atmosphere before the analysis. Data from five different pipes are used; F, G, H, I and P (figure 2.2). Pipe P, installed in August 2000 provides information about the background geochemistry in the groundwater running from the western till slope. Data from F and G is used for comparing the extent of remediation in the impoundment. The depths of the BAT®- pipes are shown in table 5.1.
Table 5.1. Depths of the BAT®-pipes F, G, H and I.
Pipe Depth (m)
F 8.02 G 3.37 H 8.205 I 3.545
The water samples were collected in glass bottles (DURAN®), which were acid-washed in 5 % HNO3 for three days. The bottles were then emptied and refilled with argon several times, so that the expected volume of oxygen left inside ranged between 0.1 and 0.01 %.
Before sampling, the groundwater in the BAT®-pipes was renewed 3 to 4 times. During the entire sampling, the sampled water was protected from light and heat.
During the period August 1998 until April 2003 the groundwater level were measured in piezometers across the impoundment, using an electronic probe. Data from 13 different piezometers are used. In figure 5.1 the installation depths of the H-pipes in the northwestern part of impoundment 1 are shown.
Figure 5.1 Installation of piezometers in the northwestern part of impoundment 1, Kristineberg mining area.
5.2 In situ analysis
Redox and pH were measured with a Metrohm® Pt-electrode and a Metrohm® combined pH electrode. The pH was calibrated using two Titrisol® pH 4 and pH 7 buffers. The redox electrode was checked using two Ag/AlCl Reagecon® standards (124 mV and 358 mV). All redox values were adjusted to the standard hydrogen electrode. The conductivity was measured with a Hanna® conductivity meter. Alkalinity was determined by titration with a decimolar Titrisol® HCl solution, using a portable Hach® digital titrator. All measurements were made within a few minutes after opening each bottle. The water left in the bottle was then divided in three subsamples for the analyses performed in laboratories: one for anions analyses (currently chloride, sulphate and nitrate), one for metals and main elements by ICP techniques and one for dissolved organic content.
The water samples for anion analyses were filtered directly into plastic bottles and placed immediately in a refrigerator. The water samples for main elements and metal analyses were filtered directly into acid washed 60 ml Azlon® HDPE bottles, acidified with 1 % redistilled suprapur HNO3.
The water samples for organic content were filtered, using a metal filter-holder and a glass container, through Whatman® glass fibre filters (0.7 m pore size and 47 mm), which had been burnt before use at 450° C for 24 hours. The water was then poured into Falcon® plastic tubes acidified with 1 % 2 M HCL and placed in the refrigerator immediately.
Blank analysis using the same methodology with milliQ water instead of groundwater verified that the conditions of manipulation and analyses were satisfactory, with less than 2 % average relative error on all pipes and elements.
5.3 Analytical methods
The samples, which were kept cold and in darkness, were brought to a laboratory within 24 hours. Fe, Al, As, Ba, Cd, Co, Cr, Cu, Mn, Mo, Ni, P, Pb and Zn were measured using ICP- SMS (high-resolution ICP-MS), and Ca, K, Mg, Na, S, Si and Sr were analysed by ICP-AES (Atomic Emission Spectroscopy). The data were reported after three to four runs on each instrument. The quality of the analyses was checked using synthetic standards. The precision determined in such conditions as ± one standard deviation was generally better then 5 %. The dissolved organic carbon content was determined using high-temperature combustion and potassium phthalate as a standard. The anions content was determined by ion
chromatography.
6 Results
6.1 Dissolved elements in groundwater
Data concerning the average values for groundwater samples in the northwestern part of impoundment 1 are shown in table 6.1. The average values are calculated on data from 2002 except the average values on data from pipe P. These values are calculated on data from sampling during the period October 2000 until October 2002. This is because pipe P is a reference pipe located in the hill slope outside the impoundment and should not be expected to change too much over time. Subsequently, taking an average value covering all three years will have the benefit to include more samples. All data from the sampling occasions are shown in appendix 1-5.
Table 6.1. Average values in the groundwater in the year 2002 for pipes located in the northwestern part of impoundment 1. The average values for pipe P are from October 2000 until October 2002.
Pipe P
(Background) (2000-2002)
Pipe F
Sampling depth:
8.02
Pipe G 1
Sampling depth:
3.37
Pipe H 2
Sampling depth:
8.21
Pipe I
Sampling depth:
3.55 pH 6.0 ± 0.3 5.8 ± 0.1 5.7 ± 0.8 4.5 ± 0.2 5.7 ± 0.2 Redox (mV) 416 ± 41.2 268 ± 12 272 ± 47.1 340 ± 2.8 268 Cond (mS/cm) 0.054 ± 0.017 0.811 0.373 ± 0.326 0.661 ± 0.9 2.97 ± 0.1 Ca (mg/l) 3.03 ± 0.26 18.7 ± 1.5 17.3 ± 3.21 380 ± 5.7 432 ± 10 Fe (mg/l) 0.39 ± 0.66 156.9 ±64.3 69.0 ± 66.8 17350 ± 919 787 ± 276 Mg (mg/l) 1.33 ± 0.13 13.8 ± 1.6 4.17 ± 0.67 1540 ± 85 28.7 ± 0.9 Na (mg/l) 1.93 ± 0.19 3.02 ± 0.43 3.39 ± 0.88 9.3 ± 0.6 5.78 ± 0.21 S (mg/l) 3.79 ±0.98 152 ± 19.7 56.8 ± 50.4 13850 ± 212 1030 ± 42.4 Si (mg/l) 5.27 ± 0.39 1.02 ± 0.03 3.56 ± 1.96 7.55 ± 0.66 6.64 ± 0.09 Al (µg/l) 22.5 ± 11.0 251 ± 35.7 3116 ± 2716 251500 ± 20506 679 ± 133 As (µg/l) 0.11 ± 0.07 2.67 790 ± 685 64.7 ± 8.56 18.4 ± 7.85 Ba (µg/l) 7.54 ± 0.92 9.05 ± 0.78 14.1 ± 8.04 6.56 ± 0.05 3.87 ± 0.21 Cr (µg/l) 0.19 ± 0.11 1.80 ± 0.24 47.5 ± 78.4 1245 ± 92 5.19
Cu (µg/l) 0.92 ± 0.56 <2 7.63 ± 10.4 <20 <10
Mn (µg/l) 17.8 ± 10.0 698 ± 67.9 340 ± 192 43350 ± 495 4650 ± 99 Pb (µg/l) 0.18 ± 0.14 <0.2 2.23 ± 1.84 11.6 ± 0.6 <1 Sr (µg/l) 25.7 ± 1.88 98.1 ± 16.4 40.9 ± 5.3 900 ± 17.7 375 ± 5.7 Zn (µg/l) 12.6 ± 6.19 882 ± 132 6113 ± 925 2510000 ±
212132 32200 ± 4101
Cl- (mg/l) 1.47 ± 0.86 3.00 ± 0.57 2.80 ± 1.32 30 5
SO42- (mg/l) 215 ± 263 507 ± 163 182 ± 171 40500 ± 12021 2800 ± 849 DOC (mg/l) 3.4 ± 1.7 3.9 ± 1.0 2.65 ± 0.81 26.4 ± 2.3 4.7 ± 0.85
1 The data shows large differences in concentrations of elements in pipe G, especially during the last two occasions during 2002.
2 Without the sampling in October 2002 when it was remarkably low content of most of the elements in pipe H.
Pipe H has concentrations of most of the elements that are higher than in the rest of the impoundment. The same result was found during the years 2000 and 2001 (Corrège et al., 2001; Corrège and Öhlander, 2001). The concentration of Fe, Mg, S, Al, Cr, Mn, Zn, Cl-, SO42- and dissolved organic matter (DOC) are much higher in pipe H than in any other pipe.
The pH-values are lower in pipe H than in the other pipes.
Data concerning average values for pipe H over the years 1998 until 2002 are shown in table 6.2. The data do not show any significant decrease in redox nor do the concentrations of metals, sulfur or dissolved organic matter decrease. The pH does not show any sign of increasing. The As-concentration though shows a tendency of increasing.
Table 6.2. Average values in the groundwater for the years 1998 - 2002 for pipe H located in the northwestern part of impoundment 1. The missing data correspond to some concentrations below the detection limits.
Regarding the data for 2002 the values from October have been ignored due to low reliability.
Pipe H
Sampling depth: 8.21
1998 1999 2000 2001 2002 pH 4.4 ± 0.09 4.4 ± 0.12 4.5 ± 0.07 4.5 ± 0.1 4.5 ± 0.2
Redox (mV) 350.7 ± 5.1 346.2 ± 7.5 339.8 ± 10.6 349.7 ± 7 340 ± 2.8 Cond (mS/cm) 26.0 ± 2.8 23.8 ± 0.7 14.7 ± 10 23.6 ± 0.5 Unrealistic Ca (mg/l) 394.7 ± 30.1 404.3 ± 22.3 387.6 ± 22.7 390.8 ± 21.8 380 ± 5.7 Fe (mg/l) 17733 ± 1736 18267 ± 983 18800 ± 2053 19350 ± 930 17350 ± 919 Mg (mg/l) 1630 ± 103 1618 ± 72 1702 ± 103 1680 ± 51 1540 ± 85 S (mg/l) 14483 ± 1091 14267 ± 726 15040 ± 829 15200 ± 1410 13850 ± 212 Si (mg/l) 11.2 ± 0.9 8.8 ± 0.6 8.2 ± 0.2 8.9 ± 0.3 7.55 ± 0.66 Al (µg/l) 207167 ±
13378 280000 ±
23307 260200 ±
14446 284500 ±
14900 251500 ± 20506 As (µg/l) 48.1 ± 8.7 33.8 ± 11.2 41.6 ± 2.8 46.6 ± 13.8 64.7 ± 8.56 Ba (µg/l) 10.2 ± 1.9 7.8 ± 1.4 9.75 ± 5.15 6.56 ± 0.05 Cr (µg/l) 1140 ± 83.7 1283 ± 127 1326 ± 159 1170 ± 140 1245 ± 92
Cu (µg/l) 12.3 31.6 ± 2.1 87.9 94
Mn (µg/l) 46233 ± 2813 48850 ± 2352 45680 ± 1669 48250 ± 4840 43350 ± 495 Pb (µg/l) 19.4 ± 3.14 17.6 ± 2.97 13.69± 4.4 12.7 ± 2 11.6 ± 0.6 Sr (µg/l) 942 ± 111 1064 ± 105 952 ± 29.8 900 ± 17.7 Zn (µg/l) 2546667 ±
178960 2396667 ±
152272 2352000 ±
358287 2552500 ±
143800 2510000 ± 212132 Cl- (mg/l) 23.8 ± 29.0 39.0 25.57 ± 35.6 30 ± 13 30
SO42- (mg/l) 57600 ± 11632 29500 ± 12477 70850 ± 21932 40500 ± 12021 DOC (mg/l) 31.0 ± 5.7 25.0 ± 3.9 22.4 ± 5.7 23.9 26.4 ± 2.3
Table 6.3 shows the average concentrations of elements in pipe I over the years 1998 until 2002. This pipe is located more shallow than pipe H, but in the same area.
Table 6.3: Average values in the groundwater for the years 1998 – 2002 for pipe I located in the northwestern part of impoundment 1. The missing data correspond to some concentrations below the detection limits.
Pipe I
Sampling depth: 3.55
1998 1999 2000 2001 2002 pH 4.6 ± 0.1 5.0 ± 0.1 5.1 ± 0.15 5.5 ± 0.2 5.7 ± 0.2
Redox (mV) 347.3 ± 7.31 302.4 ± 6.29 296.4 ± 12.9 242 ± 8.3 268 Cond (mS/cm) 15.1 ± 1.2 14.8 ± 0.7 12.6 ± 8.4 5.5 ± 1 2.97 ± 0.1 Ca (mg/l) 369.7 ± 24.2 384.4 ± 15 372.6 ± 19.8 377 ± 21 432 ± 10 Fe (mg/l) 8909 ± 814 9949 ± 498 10202 ± 608 2760 ± 730 787 ± 276 Mg (mg/l) 830.8 ± 54.4 863.3 ± 23.4 848.4 ± 18.8 155 ± 71 28.7 ± 0.9 S (mg/l) 7021 ± 611.3 7447 ± 328 7520 ± 343 2280 ± 584 1030 ± 42.4 Si (mg/l) 8.09 ± 1.97 10.4 ± 0.96 10.9 ± 0.6 6.6 ± 0.8 6.64 ± 0.09 Al (µg/l) 92256 ± 5585 45600 ± 2364 52020 ± 3779 7970 ± 3350 679 ± 133 As (µg/l) 8.2 ± 1.9 7.9 ± 3.5 6.4 ± 2.4 16.9 ± 4 18.4 ± 7.85 Ba (µg/l) 3.98 ± 0.45 3.76 ± 0.89 3.58 ± 5.26 3.87 ± 0.21 Cr (µg/l) 96.27 ± 14.6 71.03 ± 12.5 88.16 ± 11.4 9.3 ± 6.2 5.19
Cu (µg/l) 58.10 ± 11 2.81
Mn (µg/l) 23200 ± 1655 26071 ± 1075 25620 ± 1076 8460 ± 2240 4650 ± 99 Pb (µg/l) 12.54 ± 7.16 4.87 ± 0.62 4.32 ± 1.1 1.26 Sr (µg/l) 592 ± 36.8 677 ± 32.4 582 ± 27.7 375 ± 5.7 Zn (µg/l) 655556 ±
41228 632571 ±
33261 617167 ±
38964 163750 ±
42571 32200 ± 4101
Cl- (mg/l) 14.0 ± 8.98 12.2 ± 11.1 22 22 5
SO42- (mg/l) 30111 ± 16221 23400 ± 5771 29800 ± 7634 2800 ± 849 DOC (mg/l) 7.1 ± 2.1 6.2 ± 1.8 7.5 ± 2.6 5.1 4.7 ± 0.9
The pH-value in pipe I shows a tendency of increasing while the redox and conductivity seems to decrease. The concentrations of most of the metals and of sulfur is decreasing, except for As which is increasing.
Figure 6.1 illustrates the average values of pH, redox, dissolved organic carbon (DOC) and concentrations of Fe, S, Zn, Mg and Al in the groundwater from pipes H and I during the years 1998 until 2002.
Figure 6.1. Average values of pH, redox, DOC and concentrations of Fe, S, Zn, Mg and Al in pipe H and I during the years 1998 to 2002. The data from sampling in October 2002 is not taking into the diagrams.
In figure 6.2 groundwater data from pipe H during August 1998 until April 2003 are plotted.
In October 2002 pipe H shows a significant decrease in concentrations of elements, such as S, Zn, Mn, Fe, Mg, Cr and Al. The pH does not seem to increase however. It is uncertain if the data from October 2002 is accurate though. If the data is ignored, there is no significant decrease in concentrations during 2002. Figure 6.2 also shows that in April 2003 the
concentrations were at the same levels as earlier. The data for pipe H from October 2002 has been ignored in the rest of this chapter.
Meanvalues for Al and DOC during 1998- 2002
0 80000 160000 240000 320000
1997 1998 1999 2000 2001 2002 2003
Al (µg/l)
0 10 20 30 40 DOC (mg/l)
PipeH, Al Pipe I, Al Pipe H, DOC Pipe I, DOC Meanvalues for Fe and S during 1998- 2002
0 5000 10000 15000 20000
1997 1998 1999 2000 2001 2002 2003
Fe (mg/l)
0 5000 10000 15000 20000 S (mg/l)
Pipe H, Fe Pipe I, Fe Pipe H, S Pipe I, S
Meanvalues for Zn and Mg during 1998- 2002
0 700000 1400000 2100000 2800000
1997 1998 1999 2000 2001 2002 2003
Zn (µg/l)
0 500 1000 1500 2000 Mg (mg/l)
Pipe H, Zn Pipe I, Zn Pipe H, Mg Pipe I, Mg Meanvalues for pH and redox during 1998-2002
4 4,5 5 5,5 6
1997 1998 1999 2000 2001 2002 2003
pH
0 100 200 300 400 Redox [mV]
Pipe H, pH Pipe I, pH Pipe H, redox Pipe I, redox
Concentrations of Zn and Mg in pipe H
500 780 1060 1340 1620 1900
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Mg (mg/l)
600000 1200000 1800000 2400000 3000000 3600000 Zn (µg/l)
Mg (mg/l) Zn (µg/l)
Concentrations of Al and DOC in pipe H
120000 160000 200000 240000 280000 320000
Apr-98 Oct-98 Apr-99 Oct-99Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Al (µg/l)
0 10 20 30 40 50 DOC(mg/l)
Al (µg/l) DOC (mg/l)
pH and redox in pipe H
200 250 300 350 400
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 redox
4 4,25 4,5 4,75 5 pH
Redox (mV) pH Concentrations of Fe, Mn and S in pipe H
0 5500 11000 16500 22000
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Fe (mg/l)
S (mg/l)
10000 30000 50000 70000 90000 Mn (µg/l)
Fe (mg/l) S (mg/l) Mn (µg/l)
In pipe I the concentrations of metals have decreased. They dropped significantly during 2001. The pH has increased and the redox has decreased, figure 6.3.
Figure 6.3. Plots for Fe, Mn, S, Zn, Mg, Al, DOC, pH and redox in pipe I from August 1998 until April 2003.
In pipe F the concentrations of metals dropped significantly during the years 1998 and 1999.
The pH has increased from around 4.5 to over 5.5, figure 6.4.
Figure 6.4. Plots for Fe, Mn, S, Zn, Mg, pH and redox in pipe F from August 1998 until April 2003.
Concentrations of Zn and Mg in pipe F
0 100 200 300 400 500
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Mg (mg/l)
0 100000 200000 300000 400000 500000 Zn (µg/l)
Mg (mg/l) Zn (µg/l) Concentrations of Fe, Mn and S in pipe F
0 1000 2000 3000 4000 5000
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Fe (mg/l)
S (mg/l)
0 3000 6000 9000 12000 15000 Mn (µg/l)
Fe (mg/l) S (mg/l) Mn (µg/l)
Redox and pH in pipe F
200 250 300 350 400
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Redox
4 4,5 5 5,5 6 pH
Redox (mV) pH
Concentrations of Mg and Zn in pipe I
0 250 500 750 1000
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Mg (mg/l)
0 250000 500000 750000 1000000 Zn (µg/l)
Mg (mg/l) Zn (µg/l)
Concentrations of Al and DOC in pipe I
0 30000 60000 90000 120000
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Al (µg/l)
0 3 6 9 12 DOC (mg/l)
Al (µg/l) DOC (mg/l) Concentrations of Fe, Mn and S in pipe I
0 3000 6000 9000 12000
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Fe (mg/l)
S (mg/l)
4000 16000 28000 40000 52000 Mn (µg/l)
Fe (mg/l) S (mg/l) Mn (µg/l)
Redox and pH in pipe I
200 250 300 350 400
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Redox
4 4,5 5 5,5 6 6,5 pH
Redox (mV) pH
Also in pipe G the concentrations of metals have decreased. They have decreased with another pattern than pipes F and I, approximately showing a constant decrease, figure 6.5.
Figure 6.5. Plots for Fe, Mn, S, Zn, Mg, pH and redox in pipe G from August 1998 until April 2003.
Plotting the dissolved organic carbon (DOC) in the pipes F, G, H and I illustrates that the content of organic matter is between two to seven times higher in H than in the other pipes, figure 6.6.
Dissolved organic carbon in groundw ater
0 10 20 30 40
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 DOC [m g/l]
H I F G
Figure 6.6. Plots of dissolve organic carbon (DOC) in the pipes F, G, H and I during the period 1998 until 2002.
Plots with pH and Stot in pipes F, G, H and I illustrate that pH have increased in all pipes except in pipe H, and that the total concentration of sulfur has decreased in the pipes except in pipe H, figure 6.7.
Concentrations of Fe, Mn and S in pipe G
0 1000 2000 3000 4000
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Fe (mg/l)
S (mg/l)
0 2000 4000 6000 8000 Mn (µg/l)
Fe (mg/l) S (mg/l) Mn (µg/l) Linear (Fe (mg/l))
Concentrations of Zn and Mg in pipe G
0 100 200 300 400
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Mg (mg/l)
0 75000 150000 225000 300000 Zn (µg/l)
Mg (mg/l) Zn (µg/l) Linear (Mg (mg/l)) Redox and pH in pipe G
200 250 300 350 400 450 500
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 Redox
3 4 5 6 7 pH
Redox (mV) pH
Figure 6.7. pH and concentrations of Stot in the pipes F, G, H and I from August 1998 until April 2003.
Plotting pH and the concentrations of Zn in pipes I and F shows that when pH is increasing the content of dissolved zinc in groundwater is decreasing, figure 6.8.
Figure 6.8. pH and concentrations of Zn in pipe I and F from August 1998 until April 2003.
pH and concentrations of S in pipe H
4000 7500 11000 14500 18000
Mar-98 Nov-98 Jul-99 Mar-00 Nov-00 Jul-01 Mar-02 Oct-02 Jun-03
S [mg/l]
4 4,2 4,4 4,6 4,8 pH
S (mg/l) pH pH and concentrations of S in pipe F
0 1000 2000 3000 4000
Mar-98 Nov-98 Jul-99 Mar-00 Nov-00 Jul-01 Mar-02 Oct-02 Jun-03
S [mg/l]
4 4,5 5 5,5 6 pH
S (m g/l) pH
pH and concentrations of S in pipe I
0 3000 6000 9000
Mar-98 Nov-98 Jul-99 Mar-00 Nov-00 Jul-01 Mar-02 Oct-02 Jun-03
S [m g/l]
4 5 6 7 pH
S (m g/l) pH pH and concentrations of S in pipe G
0 800 1600 2400 3200
Mar-98 Nov-98 Jul-99 Mar-00 Nov-00 Jul-01 Mar-02 Oct-02 Jun-03
S [m g/l]
3 4 5 6 7 pH
S (mg/l) pH
pH and concentrations of Zn in pipe I
3 4 5 6 7
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 pH
0 250000 500000 750000 1000000 Zn (µg/l)
pH Zn (µg/l)
pH and concentrations of Zn in pipe F
4 4,5 5 5,5 6
Apr-98 Oct-98 Apr-99 Oct-99 Apr-00 Oct-00 Apr-01 Oct-01 Apr-02 Oct-02 Apr-03 pH
0 125000 250000 375000 500000 Zn (µg/l)
pH Zn (µg/l)
