Sedimentation Processes and Nutrients Turnover in Acidic Water Bodies during Experimental Treatment for Lake Remediation

Bachelor Thesis , 2006 , 75 Pages , Grade: 1,7

Environmental Sciences

The objective of the study is charactering of the primary flux in an enclosure experiment of chemical and/or biological neutralisation in one of the numerous acidified lakes in the Lusatian mining district, (Germany) and tracing the consequences for the water chemistry, reps. water quality.

Sedimentation as one of the most important physical process in the water controls the energy flow through aquatic ecosystems thus normally transporting nutrients to the lake bottom. The rate of seston deposition has great importance for understanding the circulation of organic and mineral matter within a lake and can be regarded as a measure of the nutrient recycling there.

Acidity removal amendments (soda ash, organic mater and phosphate) added to the enclosures in different combinations resulted in components specific seston fluxes and in changes in the biological and chemical parameters of the water column. It was found that the addition of neutraliser and organic amendments (ENePo) resulted in a flux of authochthonous particulate matter of 6 - 4g/(m²d) and enhanced at most and in a long term (in the course of the experiment) the nutrients (P, C, N) cycling and the water quality (as increasing from the bottom pH gradient was observed).

The addition of organic matter to the acidic water column (EPo) resulted in similarly enhanced gross sedimentation but with a small time lag compared to the same treatment of the neutralized water as the assimilation of the nutrients was hindered by the low pH (retaining P in the sediment) and the initially low primary production. The addition of phosphates (ENePo, EP) did not result in a long term and self-sustained enhancement of the primary flux as after P being assimilated and transported to the sediment it remained immobilized there, thus the P elemental flux as well as the flux of particular matter dropped down to the usual for the lake values.

Häufig gestellte Fragen zu: Charakterisierung des Primärflusses in einem Versuch mit chemischer und/oder biologischer Neutralisierung eines sauren Bergbausesees

Was ist das Thema der vorliegenden Arbeit?

Die Arbeit untersucht die Sedimentationsprozesse und den Nährstoffumsatz in einem sauren Bergbausesee (Grünewalder See, Lausitz, Deutschland) im Rahmen eines experimentellen Versuchs zur Sanierung des Sees. Der Fokus liegt auf der Quantifizierung des Primärflusses (Sedimentation) und seiner Beziehung zum Nährstoffumsatz (insbesondere Phosphor, Kohlenstoff und Stickstoff) unter verschiedenen Behandlungsbedingungen (Neutralisierung, Düngung, organische Substanzzugabe).

Welche Methoden wurden angewendet?

Es wurde ein In-situ-Enclosure-Experiment durchgeführt. Zwölf enclosures mit einem Durchmesser von 1 m wurden im See installiert und mit verschiedenen Behandlungen versehen (Kontrollgruppe, Neutralisierung mit Soda, Düngung mit Phosphat, Zugabe von Kartoffeln als organische Kohlenstoffquelle, Kombinationen davon). Sedimentationsfallen wurden eingesetzt, um die Sedimentationsraten und die Nährstoffflüsse zu messen. Zusätzlich wurden regelmäßig physikalisch-chemische Parameter (Temperatur, pH, gelöster Sauerstoff, Redoxpotential) und die Konzentrationen von Gesamtphosphor (TP), Gesamteisen (TFe), Gesamtkohlenstoff (TC), Gesamt-Stickstoff (TN), gelöstem organischem Kohlenstoff (DOC) und gelöstem anorganischem Kohlenstoff (DIC) gemessen.

Welche Ergebnisse wurden erzielt?

Die Zugabe von organischer Substanz, Neutralisierungsmitteln und/oder Düngemitteln, einzeln oder kombiniert, erhöhte den Partikelstrom zu den Sedimenten und veränderte die Wasserchemie und die Sedimentmikrobiologie. Die stärkste Steigerung der Sedimentationsrate wurde bei der Kombination von Neutralisierung und organischer Substanz beobachtet. Die Zugabe von Phosphat führte zwar anfänglich zu einem starken Anstieg der Sedimentationsrate durch Algenwachstum (Algenblüte), dieser Effekt war jedoch nicht langfristig stabil. Die Neutralisierung allein hatte einen kurzfristigen Effekt auf die Sedimentationsrate. Es wurden Zusammenhänge zwischen Sedimentationsraten und den Konzentrationen verschiedener Nährstoffe gefunden, insbesondere im Zusammenhang mit der organischen Substanzzugabe.

Welche Schlussfolgerungen lassen sich ziehen?

Der Partikelstrom zu den Sedimenten im sauren Bergbausesee Grünewalde ist gering und wird stark durch die Zugabe von Kohlenstoffquellen und Phosphor beeinflusst. Die Zugabe von organischer Substanz erwies sich als besonders effektiv für eine langfristige Steigerung des Sedimentationsflusses. Die Kombination von Neutralisierung und organischer Substanz führte zur Bildung von Eisensulfid und zu einer Verbesserung der Wasserqualität. Die Studie unterstreicht die Bedeutung der Partikelzirkulation und der Sedimentationsprozesse für die Wasserqualität und den Nährstoffumsatz in sauren Gewässern.

Welche Bedeutung hat die Studie?

Die Studie liefert wichtige Erkenntnisse für die Sanierung saurer Bergbauregionen. Sie zeigt die Wirksamkeit verschiedener Sanierungsmethoden und deren Einfluss auf die Sedimentation und den Nährstoffumsatz. Die Ergebnisse können zur Entwicklung optimierter Sanierungsstrategien für ähnliche Gewässer beitragen und die langfristige Entwicklung der Wasserqualität vorhersagen.

Wie wurden die Daten analysiert?

Die Daten wurden statistisch aufbereitet, um Messfehler zu berücksichtigen. Es wurde der Variationskoeffizient verwendet, um die Zuverlässigkeit der Messungen zu beurteilen. Bei großen Abweichungen zwischen parallelen Messungen wurden die Daten entsprechend korrigiert.

Welche Literatur wurde verwendet?

Die Arbeit verweist auf eine umfangreiche Liste von wissenschaftlichen Publikationen zu den Themen saure Bergbauregionen, Sedimentation, Nährstoffkreisläufe, Sanierungsmethoden und Wasserchemie. Die Referenzliste bietet detaillierte Informationen zu den verwendeten Quellen.

Table of Contents

Abstract

Table of Contents

List of Tables

Abbreviation and Annotations

Statement of Objectives

Chapter 1 Literature Overview
1.1. Acidic Mining Lakes in Lower Lusatia
1.2. Chemical Model of Acidification as a Result of Disulfide Oxidation
1.3. Ranges of pH in Acid Mining Lakes in Lusatia
1.4. Ways of Controlling Acidity
1.5. Primary Flux and Nutrients Cycling

Chapter 2 Materials and Methods
2.1. Mining Lake Grünewalde - Site description
2.2. Enclosure Experiment
2.2.1. Experimental Set-Up
2.2.2. Enclosure Treatment Hypothesis
2.3. Sedimentation Traps
2.4. Analyses
2.4.1. Temperature, pH, Dissolved Oxygen and Redox Potential
2.4.2. Total Phosphorus (TP) and Total Iron (TFe) Determination
2.4.3. Nitrogen and Carbon Determination
2.4.4. Rates of Sedimentation
2.5. Statistical Methods of Data Analysis

Chapter 3 Results and Discussion
3.1. Fluxes of Particular Matter
3.1.1. Sedimentation in the Lake and Controlling Enclosures (EC)
3.1.2. Sedimentation in the Neutralised and Fertilised Enclosures (ENeP)
3.1.3. Fehler! Textmarke nicht definiert.Sedimentation in the Neutralised Enclosure (ENe)
3.1.4. Sedimentation in the Fertilized Enclosure (EP)
3.1.5. Sedimentation in the Organic Matter Amended Enclosure (EPo)
3.1.6. Sedimentation in the Neutralised and Organic Matter Amended Enclosures (ENePo)
3.2. Chemical Changes in the Aqueous Phase and Nutrients Turnover in the Time Course of the Enclosure Experiment
3.2.1. Vertical Distribution of pH and Dissolved Oxygen in the Lake
3.2.2. pH Dynamics in the Course of the Experiment
3.2.3. Changes in the Carbon Concentrations
3.2.4. Changes in the Nitrogen Concentrations
3.2.5. Changes in the Total Phosphorus Content
3.2.6. Changes in the Total Iron Content
3.3. Nutrient Fluxes to the Sediments
3.4. Sedimentation and Water Quality

Chapter 4 Conclusions

Chapter 5 Summary

References

Appendices

List of Tables

Table 1. Reactions of acidification of lakes by oxidation of pyrite and marcasite (after Stumm & Morgan, 1981, Evangelou, 1995)

Table 2. Classification of AML based on acidity and pH (after Nixdorf et al., 1997, Nixdorf et al., 2005)

Table 3. (Bio)Chemical processes and reaction increasing alkalinity (adopted from Wend-Potthoff & Neu, 1998, Uhlmann et al., 2004)

Table 4. Morphometrical characteristics of Lake Grünewalde, Lusatia, Germany summarized from Grüneberg & Kleeberg (2005) (mean values of 2000 - 2003, n = 64) and Friese et al. (1998) (mean values for 1997 - 1998)

Table 5. Chemical characteristics of Lake Grünewalde, Lusatia, Germany (measurements of the Institute for Freshwater Ecology, Berlin (IGB) on 14th August 2002, n = 2)

Table 6. Enclosure treatments

Table 7. Error estimation in the spreadsheets

Abbreviation and Annotations

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Abstract

The objective of the study is charactering of the primary flux in an enclosure experiment of chemical and/or biological neutralisation in one of the numerous acidified lakes in the Lusatian mining district, (Germany) and tracing the consequences for the water chemistry, reps. water quality.

Sedimentation as one of the most important physical process in the water controls the energy flow through aquatic ecosystems thus normally transporting nutrients to the lake bottom. The rate of seston deposition has great importance for understanding the circulation of organic and mineral matter within a lake and can be regarded as a measure of the nutrient recycling there.

Acidity removal amendments (soda ash, organic mater and phosphate) added to the enclosures in different combinations resulted in components specific seston fluxes and in changes in the biological and chemical parameters of the water column. It was found that the addition of neutraliser and organic amendments (ENePo) resulted in a flux of authochthonous particulate matter of 6 - 4g/(m²d) and enhanced at most and in a long term (in the course of the experiment) the nutrients (P, C, N) cycling and the water quality (as increasing from the bottom pH gradient was observed).

The addition of organic matter to the acidic water column (EPo) resulted in similarly enhanced gross sedimentation but with a small time lag compared to the same treatment of the neutralized water as the assimilation of the nutrients was hindered by the low pH (retaining P in the sediment) and the initially low primary production. The addition of phosphates (ENePo, EP) did not result in a long term and self-sustained enhancement of the primary flux as after P being assimilated and transported to the sediment it remained immobilized there, thus the P elemental flux as well as the flux of particular matter dropped down to the usual for the lake values

Statement of Objectives

This work is based on an enclosure experiment where the addition of nutrients and/or organic matter leads to an enhanced primary production which in turns is related to an increased fluxes of particulate matter toward the sediment. Furthermore, the increase in supply of organic carbon stimulates microbial activity at the sediment water interface leading to alkalinity generation.

Sedimentation and nutrients turnover are interconnected as the recycling of elements in the water column is achieved downward through sedimentation and vice versa, the assimilation, retention, release, complexing or precipitation of those elements defines the allochtonous primary flux to the sediment. The results of this study focus on the sedimentation processes that have taken place in the component specific enclosures and in the lake, itself, and on the relationship sedimentation rates – nutrients turnover, in particular the elements phosphorus, carbon and nitrogen.

Thus, this study strives to answer the following questions:

- To what extent is the sedimentation in the enclosures enhanced, depending on the different treatments (nutrient and organic matter amendments), in comparison to the lake?
- Are there component - specific differences (TOC, TP, TFe, TN)?
- What are the consequences of an enhanced particulate matter flux for the water column chemistry?

Chapter 1 Literature Overview

1.1. Acidic Mining Lakes in Lower Lusatia

Germany has a number of lignite mining districts but the most important in terms of production are: the Lower Rhine district (Rheinland), the central German district in Saxony, Saxony-Anhalt (Mitteldeutschland), and the Lusatian district in the south-eastern part of Brandenburg (Lausitz)^[1] (Fig. 1).

The central and Lusatian districts have been exploited since the 17th century with a production peak of 330 million tons in 1989 (Schrenk & Glässer 1998). In the last fifteen years the extraction of lignite was greatly reduced leading to a brown coal production of about 59 million tons in the Lusatian district in 2004 compared to the production of 195.1 million tons in 1989^[2]. The areas of open cast mines were reclaimed after closure of the mining activity and in the district of lower Lusatia a total of 36 km² has become lake areas (Henning 1994, cited in Schultze & Geller, 1996) as a consequence of the infilling of the pits from open cast lignite mining with ground or river water. More than 400 artificial mining lakes (Nixdorf et al., 1997, Lessmann et al., 2000) of quite different size, shape, morphometry, mixing regime and hydrochemistry exist nowadays. They include some of the largest and most acidic lakes in Germany (Nixdorf et al., 1997).

Mining activities cause severe landscape disturbance worldwide and surface mining is inevitably associated with several environmental problems such as disturbances in the natural water balance, mass transfer of billion tons of soil and landscape disturbances (Friese et al. 1998). Moreover, open-cast mining operations replace the former dynamic equilibrium of the landscape, eventually resulting in the development of new ecosystems (Hüttl & Gerwin, 2005).

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Fig. 1. Major mining districts in Germany (adopted from Schrenk & Glässer 1998)

The water composition of the newly formed lakes occupying the former open pits is a consequence of the chemical reactions and the biogeochemical interaction that have taken place in the mixing and removing of the overburden and the increased groundwater level. These lakes are characterized with high concentrations of sulphate and iron (e.g. Friese et al., 1998, Marchand & Silverstein, 2002), products of the oxidation of sulphide - containing minerals within the open pits. Primary production and/or microbial processes are limited mainly by the inorganic carbon (DIC) and the total phosphorus (TP) usually does not exceed 5μg/l (Nixdorf et al., 1997) due to e.g. co-precipitation with Fe(III) oxyhydroxides (Lessmann et al., 2000). Based on the phosphorus concentrations the lakes are oligotrophic or mesotrophic. Prediction of the trophic state from the lake morphometry suggests that the lakes eventually would develop to oligotrophic or mesotrophic (Nixdorf et al., 1997).

These lakes being very young have shallow sediment cover, muddy and fluffy, between 15 and 19 cm, influenced by chemical precipitation (Fyson et al., 1998) and reaching water contents of up to 90%. The organic contents in the sediment is reported to be remarkably significant but still limiting the microbial turnover rates (LOI = 52.16% [illustration not visible in this excerpt] 2.73%, n = 14, App. II).

1.2. Chemical Model of Acidification as a Result of Disulfide Oxidation

One of the main problems of the mining lakes is the extreme acidification resulting from oxidation of pyrite and marcasite, the two major forms of [illustration not visible in this excerpt], iron - disulfide, the most abundant sulphuric mineral in the earth’s crust (Marchand & Silverstein, 2002). Both materials have the same chemical composition, but differ crystallographycally - the pyrite is isometric ([illustration not visible in this excerpt]-160.2 kJ/mol), but the marcasite is orthometric and less stable ([[illustration not visible in this excerpt]]-158.4 kJ/mol) (Evangelou, 1995). Pyrite is also the form that is normally encountered in recent environments. Pyrite is formed in a reducing environment with a continuous supply of sulphates and iron in the presence of easily decomposable organic matter. Carbon-to-sulphur ratio, availability of iron, and oxidation potential are the major factors that determine the rate of pyrite formation (Evangelou, 1995).

In order to obtain coal the pyrite containing overburden of the lignite area has to be removed. During hauling, due to changes in the environmental conditions in the overburden as a result of contact with oxygen contained in the air, these sulphide species are oxidized to sulphate, iron and hydrogen ions. Iron bacteria catalyse the reactions of pyrite oxidation (Stumm & Morgan, 1981, Prein & Mull, 1998) and the amount of weathering products depends on the availability of oxygen.

The rate of oxidation depends on a number of factors such as pH, O2, specific surface and morphology (grain size, crystal structure) of pyrite, presence or absence of bacteria, hydrological conditions of the environment (Evangelou, 1995; Evangelou, 1998).

The reactions of pyrite oxidation after exposure to air and water are shown in Table 1.

Direct oxidation of pyrite follows Eq. (1) in Table 1 and is the strongest (2 mol of [illustration not visible in this excerpt] for 1 mol [illustration not visible in this excerpt], Stumm & Morgan, 1981) naturally occurring acidification reaction known in the nature. This slow oxidation leads to the development of acidic conditions under which the ferrous iron Fe (II) formed is relatively stable in the presence of oxygen (Singh et al., 1997).

Table 1. Reactions of acidification of lakes by oxidation of pyrite and marcasite (after Stumm & Morgan, 1981, Evangelou, 1995)

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Subsequently, acidophilic lithoautotrophic bacteria, such as Acidithiobacillus ferrooxidans^[3], catalyze the oxidation of ferrous Fe (II) to ferric iron Fe (III) (Küsel 2002), shown in Eq. (2), which is soluble under acidic conditions below pH 2.5. Upon initiation of pyrite oxidation, the ferric iron can be reduced by the pyrite itself as in Eq. (4). Therefore, pyrite continues to oxidize as long as ferric iron is regenerated and it is quite accepted that this is the rate limiting reaction in the oxidation of pyrite (Evangelou, 1995, Singh et al., 1997, Pronk & Johnson, 1992). The ferrous iron formed is oxidized again microbiologically. Thus, pyrite oxidation increases its rate. Moreover, if freshwater is available and mixed with the acid mining drainage, three more protons are released, following Eq. (3) after the formation of solid hydroxide.

To sum it up, ferric Fe(III) iron is considered the major pyrite oxidant in the acid pH region (Stumm & Morgan, 1981), while oxygen is expected to be the direct oxidant at circumneutral pH. However, Evangelou (1998) supports the opinion that ferric Fe(III) iron could be the major direct pyrite oxidant at neutral pH. In any case, since oxidation of ferrous to ferric iron is very slow at the pH of 3, it turns out that the reaction is catalyzed by microorganisms. At low pH, an acidophilic chemoautotrophic bacterium T. ferrooxidans, able to oxidize iron, catalyzes and accelerates the oxidation of Fe2+ by a factor larger than 106. Thus the process of reduction of the Fe3+ to Fe2+ by pyrite and oxidation of Fe2+ to Fe3+ by atmospheric O2, catalyzed by T. ferroxidans, constitutes an effective continuous oxidation cycle (Küsel, 2003).

For this reason T. ferroxidans is considered to be primarily responsible for the rapid oxidation of pyrite in mine waste at low pH. Based on the above, pyrite oxidation can be significantly reduced by complexing/precipitating Fe3+, inhibiting Fe3+ production and/or depriving the system of oxygen (O2) (Evangelou, 1995).

1.3. Ranges of pH in Acid Mining Lakes in Lusatia

Despite their similar formation processes and relative proximity a number of chemical differences in the surface water as well as in the sediment exist in the acidic mining lakes of Lusatia.

Schultze & Geller (1996) suggest grouping the acidic mining lakes in the Lusatian district of Germany into three groups according to their hydrochemistry and particularly, pH values, Fe and Al concentrations. The most acidic waters with pH 2.7 to 3.4 have very high Fe concentrations, low Al concentrations and high ionic strength (e.g. ML 78, Grünewalde Lake, Koschen Lake, Lake Waldsee, etc.). The second group of lakes has intermediate pH-values but still higher concentrations of Al than Fe (e.g. Rosa Lake), and the third group has neutral pH values with low concentrations of both Al and Fe (e.g. Senftenberger See, Blauer See, etc.).

The pH of circumnatural lakes is usually stabilized by the CO2-bicarbonate-carbonate buffering system, which dominates the ionic composition of natural waters independently of their ionic strengths. Increased anionic contents exhausts the buffering capacity of the carbonate system at pH 6, resulting in a rapid decrease in the pH value with further addition of acid (Geller et al., 1998). In acid rain-acidified softwater lakes and rivers, the pH range is again stabilized between pH 4.5 and 5.5 by the ionic and hydroxides Al species dissolved in the acidic rainwater. A third stabilization maximum between pH 2 and 4 is observed in the acidic mining lakes in Germany (Geller et al., 1998). As solubility of the metals increases with decrease of the pH (Stumm & Morgan, 1981), in those strongly acidic lakes, concentrations of Fe could be up to 1 kg Fe m-³ (Geller et al., 1998). The different species of ferric hydroxides and the ional Fe(III)-forms show a buffering capacity that stabilizes the pH values between 2 and 4, which is comparable with that of Al (pH 4.5 - 5.5) and of the carbonate system (pH 6 - 7). The Fe(OH)x-buffering system is based on the different species of iron oxyhydroxides and correlated with high concentrations of iron. The Fe(OH)x system was known to be present in acidified soils where efficacy of the Fe buffer is assigned to the range of 2.4 - 3.0 (Ulrich 1981 cited in Geller et al., 1998).

Furthermore, Geller et al. (1998) divide the Lusatian lakes according to their sediment pH into 3 groups:

- acidic lakes with acidic sediments - ML-F, ML-Skado, ML-Koschen, and the investigated ML 117
- acidic lakes with natural sediments, which sounds quite unusual but the sediment alkalinity is supposed to be result of organic matter degradation - ML-­78, ML-Halbendorf, etc.
- natural lakes with natural sediments - ML-Senftenberg, ML-4, ML-B

Nixdorf et al. (1997) suggest alternative classification (Table 2) based on the hydrogen concentrations and the acidity KB4.3 of the lake water.

Table 2. Classification of AML based on acidity and pH (after Nixdorf et al., 1997, Nixdorf et al., 2005)

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In conclusion, in the recent years a number of investigations on the mining lakes in Lusatia have been conducted, defining their status quo, testing the buffering capacity of the waters or suggesting restoration techniques.

1.4. Ways of Controlling Acidity

Acidified freshwater systems of former open-cast mines can stay stable for several decades (Yokom et al. 1997, cited in Kleeberg & Grüneberg, 2005). Under such acidic conditions aquaculture is excluded and swimming or other sports are universally recommended at pH > 6 (Totsche et al., 2003). The general question of whether to do something against the acidity or not is a matter of water policy that takes into account the demands of the society, different options for lake utilization, quality demands, etc. (Klapper et al., 1998). However, there is political pressure to develop the vast Lusatian region for recreational purpose (Totsche et al., 2003). Therefore, “intended” removal of the acidity in the lake ecosystems of the Lusatian region might be necessary in certain cases.

No sustainable method for the treatment of whole lakes acidified by AMD is available in practice up to date (Wendt-Potthoff et al., 2002). Three types of alternative acidity removal treatments are most popular (Uhlmann et al., 2004, Totsche & Steinberg, 2004):

i. flooding with (euthrophicated or circumneutral) water
ii. conventional chemical neutralization
iii. microbiological alkalinity production

The extremely high acidity of the lake water can be diluted and partially neutralized after flooding with neutral river water. However, water insufficeincy in some cases requires alternatives. Acidity can be chemically neutralized with addition of soda or lime. Finally, the microbiological treatment acts as the reverse of the acidification processes where reduction of Fe(III) and sulphate results in a net gain of alkalinity (e.g. Peine & Peiffer, 1996, Wend-Potthoff & Neu, 1998). These processes of desulphurication and pyrite formation are carried out by bacteria in anoxic conditions such as those found usually in sediments. Furthermore, as most sulphate reducing bacteria are heterotrophs and as already mentioned acidic mining lakes are generally characterized with low DOC contents, organic carbon concentration needs to be increased. It can be enhanced in two ways: (i) direct addition of carbon source and/or (ii) stimulation of the primary production, e.g. by controlled eutrophication/ saprobisation.

Peine & Preiffer (1996), Friese et al. (1998) e.g., have shown that acid mining lakes neutralize over time. Some possible in-lake neutralization mechanisms are still to be elaborated in details. Nevertheless, there are a number of in-lake chemical reactions that are known to remove hydrogen ions from freshwater ecosystems such as the buffering capacity of the water body as well as ion exchange reaction with colloidal materials (Wend-Potthoff & Neu, 1998). In addition, there are several biological processes as well. The significant biological reactions (Table 3) are oxygen reduction, nitrate reduction, manganese and iron reduction, sulphate reduction and methanogenesis.

Microbial processes as a potential approach to in situ acidic lake remediation include a number of reduction processes where an important point is a suitable electron donor to support heterotrophic activities and permanent elimination of the metal sulphides to prevent further generation of acidity (Wend-Potthoff & Neu, 1998)

Table 3 . (Bio)Chemical processes and reaction increasing alkalinity (adopted from Wend-Potthoff & Neu, 1998, Uhlmann et al., 2004)

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1.5. Primary Flux and Nutrients Cycling

Sedimentation is one of the most important physical process in the water (Kozerski, 1994) and a process controlling the energy flow through aquatic ecosystems (Cálvez & Niell, 1994). Inorganic and organic nutrients are normally transported to the lake bottom from the source (euphotic or coastal zones) through sedimentation and the settling materials make up seston. The seston, being comprised of living organisms, biological debris (tripton), organic macromolecules, clays, minerals and different oxides, has complex structure and chemical composition. These particles in lakes are inorganic minerals or organic products of decomposition of the terrestrial biomass coming form the watershed (allochthonous matter) or/and predominantly organic in nature particles produced by algae (authochthonous matter) (Tartari & Biasci, 1997). However, generally most of the suspended particles are created inside the lake. It follows, the rate of particulate matter deposition can be regarded as a measure of the nutrient recycling in a lake.

Furthermore, the sedimented particular matter has its influence on primary production in lakes and through the primary production on most of the biological and chemical activities. The chemical influence, itself, is threefold, according to Golterman (1984) and mainly concerns the two elements, phosphorus and nitrogen, in the following ways:

i. Part of the mineralization processes take place in the sediments and refractory material can accumulate;
ii. Nutrients are directly adsorbed onto sediment particles;
iii. The mineralization process provides reducing capacity;

Thus, the the general principle of nutrient cycling between the lake sediments and the water seem to be well known and comparatively simple to outline. Primary produced allochtonous and autochtonous organic mater may, after sedimentation, undergo complete decomposition due to microbial mineralization. After being incorporated into comparatively insoluble particular material in different decomposition stages, nutrients are by sediment metabolism transformed into simple soluble and also gaseous and/or volatile forms such as phosphate, ammonium, nitrate and carbon dioxide. In these forms the nutrients can be assimilated within the sediments, transferred between different pools, or transported upwards by diffusion into the overlying water.

The general description of nutrient cycling is well known, however, quantifying those processes is uneasy task. The sedimenting material may be separated into humic and non - humic substances. The humic compounds derived from biomass material decomposed by microbial activities are relatively resistant to further degradation. Non - humic material e.g. carbohydrates, proteins, amino acids, fats, pigments and other low molecular compounds, is more labile and relatively easily decomposed by organic by microorganisms (Forsberg, 1989). Different quantities of these material become stored in the sediments, where it is further mineralized, or stored in decomposition resistant forms. Dissolved low molecular weight organics typically occurring in concentrations of 1-5 μg/l are generally very rapidly metabolised while macromolecular material can be quite resistant to microbial decompo­sition. Moreover, humic substances which are increasingly more difficult to decompose with time tend to accumulate in the sediments (Forsberg, 1989). As summarized by Wetzel (2001) the highest bacterial abundance are to be found in the upper 2 cm of the sediment while at the depth of 10 cm the number of bacteria is reduced to less than one of the population maximum.

To sum it up, knowledge of temporal and spacial patterns of seston fluxes and interactions at the water-sediment interface is crucial for understanding the processes in the lakes as the release of nutrients from the sediments affects the production of authochtonous matter and the trophic state of lakes.

Chapter 2 Materials and Methods

2.1. Mining Lake Grünewalde - Site description

This study is based on a field work carried out on Lake Grünewalde (also known as ML 117 or Plessa 117 or ML 117 Lauchhammer or Grünewalder Lauch) in the time period August 2002 - October 2002. Lake Grünewalde is an oligotrophic dimictic lake with anoxic hypolimnium in some years, situated in sub-district Lauchhammer, about 100 km south of Berlin (Germany), with a relatively large for an acid mining lake surface area (0.94 km²), a mean depth of 6 – 7 m and a maximum depth of 14.4 m. An aerial photo of the lake and its surroundings is shown in App. I.

The lake exists for more than 40 years and together with 8 more relatively smaller interconnected lakes was formed after filling of an old dump with groundwater. Grünewalde water has pH 3 [illustration not visible in this excerpt] 0.1 and its sediment has pH 4.2; about 0.75 mg TP/g DW is the average concentration in the epilimnion, and 16 mg/l iron (TFe) (s. Table 4 and Table 5).The lake is very transparent with Secchi depth of approximately 6 m during summer (BTUC measurements in the time period August - October 2002). Information about the phosphorus concentrations and the phosphorus flux in the lake four months before the experiment was undertaken are shown at Fig. 2.

Limnological, morphological and chemical parameters of the lake are summarized in Table 4 and Table 5.

Table 4. Morphometrical characteristics of Lake Grünewalde, Lusatia, Germany summarized from Grüneberg & Kleeberg (2005) (mean values of 2000 - 2003, n = 64) and Friese et al. (1998) (mean values for 1997 - 1998)

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Table 5. Chemical characteristics of Lake Grünewalde, Lusatia, Germany (measurements of the Institute for Freshwater Ecology, Berlin (IGB) on 14th August 2002, n = 2)

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The number of species in the extremely acid lakes is very low in general; however, the rate of photosynthesis and bacterial production there is as high as in non-acidic lakes and the biological productivity is not limited by the pH, but by the low nutrient concentrations (Lessmann et al., 1999, Peine & Peiffer, 1996), mainly min. production of autochthonous carbon, esp. unavailability of inorganic carbon and very low phosphate supply. The phytoplankton may develop a high biomass in the hypolimnion or near the sediment with dominant taxa belonging to Chrysophyceae (Ochromonas, Chromulina) , Chlorophyta (Chlamydomonas) and Dinophyta (Gymnodinium, Peridinium umbonatum) (Nixdorf et al., 1998, Lessmann & Nixdorf, 2000).

Fig. 2. Epi- and hypolimnion Total Phosphorus (TP), Available Phosphorus(SRP) and Phosphorus Flux during Spring/ Summer 2002 in Lake Grünewalde sampled with a floating sedi­menta­tion trap at the depth of 8 m and 12 m (by courtesy of B. Grüneberg)

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2.2. Enclosure Experiment

2.2.1. Experimental Set-Up

A pair of 12 enclosures was set up at a distance of 15 - 20 m from each other on two different dates for technical conveniences, respectively on 14.08.2002 and 21.08.2002.

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Fig. 3. Enclosure set-up

Solid nylon cylinders supported by metal rings, 1 m in diameter, were installed from the bottom to the water surface and buoyed up by a rectangular tube. The enclosures (Fig. 3) were situated at a spot of 6 m depth which is the mean depth of the lake itself.

2.2.2. Enclosure Treatment Hypothesis

Six different treatments have been applied to the 12 enclosures, and most of the single treatments were duplicated or triplicated, in order to obtain base for statistics or in case an enclosure breaks down and starts leaching. The additives were chosen taking into account the particular chemical and biochemical parameters of the lake water, as well as the application costs of the treatments.

The different treatments and the objective of the additives applied there are summarized in Table 6 with respect to each enclosure.

The addition of phosphorus and a neutraliser (ENeP) aims at a rapid generation of autochthonous material. After neutralization, Fe- and Al- concentrations in the water column were expected to drop down as a result of precipitation or complexing and DIC concentrations to increase. A small increase in the phosphorus concentrations should enhance the primary production, and result in a plankton biomass, which after dying and sinking to the bottom will be transformed into autochthonous organic matter necessary for sulphate and iron reduction and will partially inhibit further pyrite oxidation.

A milder biotechnological approach was applied to four other enclosures (ENePo, EPo) where in total three kg of potatoes per enclosure were added according to previously conducted mesocosm experiment (Fyson et al., 1998) at the rate of 0.17g raw material/l (which equals 0.03g DW/l or 2.8 g DW/m2).

Potatoes provide decomposable supply of carbon, 65% DW starch as well as 25 mg/g DW phosphorus (Scherz & Senser 1994, cited in Fyson et al., 1998). The addition of potatoes is expected to result in pH increase in the near bottom layers and in increased concentrations of dissolved Fe, P and C in the system (Fyson et al., 1998). A pH gradient which developed in the mesocosm experiment is expected to develop here as well. Therefore, pH first in hypolimnion and afterwards in epilimnion is expected to increase resulting in slight chemical stratification.

Table 6. Enclosure treatments

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Fig. 4. Addition of potatoes as a source of organic carbon and nutrient

One enclosure (ENe) has been only neutralized (addition of Na2CO3) to the pH of app. pH 7. The water in the enclosure is expected to become most probably again acidic over time as a result of diffusion. Nevertheless, even if the change in the pH is not permanent or have unstable behaviour, enhanced sedimentation rates are expected as a result of increased local pH and TIC. However, it might be not enough for significant changes in the primary production, as phosphorus could remain limiting factor.

Only phosphorus has been added to one of the enclosure (EP) to develop the effect of controlled eutrophication. The primary production is expected to increase as well as the settling fluxes of TOC. In the course of time the phosphorus concentrations are expected to decrease but still to remain slightly higher compared to that of the lake.

Three enclosures (EC) do not contain additives but are intended to be a reference point for the processes taking place into the lake itself with the slight difference that water column in those enclosures is expected to be more stable and maybe to develop a chemical stratification of some kind.

A pair of sedimentation traps is submerged to the bottom of the lake in close proximity to the enclosures. They record the processes and gross sedimentation taking place in the lake during that certain course of time.

2.3. Sedimentation Traps

The primary flux, e.g. sedimentation, is an essential loss process of particular matter and energy in the aquatic systems (Kozerski, 1994). Sedimentation traps (Fig. 5) were used to measure the total sedimentation and the respective nutrient fluxes of the nutrients of interest. Two transparent Plexiglas cylinders 0.3 m in height and 0.11 m in diameter (aspect ratio = 2.7), were submerged to the bottom of each enclosure. The sedimentation traps were hitched up to the surface with the help of a buoyed rope and carefully submerged to the bottom to be emptied weekly during the period from 27/08/ to 18/09/ (4 samplings) and afterwards biweekly until 28/10/ (3 more samplings). Upon retrieval, the water of the top 23 cm of the traps was siphoned off over 1 hole and the amount of settled material has been determined form the bottom 7 cm water volume (Fig. 5, h).

Despite existing uncertainties sediment collecting has its history and traps are considered a good tool to measure the settling fluxes (Bloesch & Burns, 1980). The instrumental errors according to the literature are within reasonable limits (differences between parallel cylindrical traps are mostly within [illustration not visible in this excerpt] 10%). A properly proportioned cylinder is said to collect 95 - 100% of the real sedimentation rate (Bloesch & Burns, 1980) as choosing a cylindrical trap with appropriate ratio can avoid overtrapping as well as undertrapping of particulate material.

Based on theoretical considerations, invitro and in situ experiments and practical experience Bloesch and Burns (1980) recommend the appropriate trap for a lake study should be a simple cylinder made of transparent plastic with a diameter varying between 5 and 20 cm and having an aspect ratio (height : diameter) of about 5 for calm water bodies.

The settling flux of particles into sediment traps depends almost entirely on their concentration and their settling velocities. Turbulence does not affect the settling velocity of particles significantly, but can affect the flux if concentration gradients are present. This necessitates that the cylinders have aspect ratios which increase with the degree of turbulence of the waters in which they are immersed (Bloesch & Burns, 1980; Kozerski, 1994). Therefore, an aspect ratio of 2.7, which according to the sedimentation traps literature is low, is considered sufficient in the conducted enclosure experiment as no horizontal currents are expected to affect the isolated enclosures (D = 1 m, H = 6 m) and the water column in the tubes is assumed to be calm (Fig. 6).

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Fig. 4 . Sedimentation trap design

Fig. 5 . Enclosure design with hypothetical placement of the pair sedimentation traps (not drawn in scale)

Advisably traps are used in duplicates and the standard deviation with replicates catches usually within 10% and rarely exceeds 20% (Bloesch & Burns, 1980). Unfortunately, it has not been observed to be the case in this experiment, where the standard deviation of parallel traps is in the range of 5 to 252%. Possible reasons could be that the enclosures were small in diameter (d = 1 m) and in case one of the sedimentation traps was placed too close to the nylon wall, there has been the effect of shadow on the trap caused by the nylon. In such cases there has been under-trapping. Another possible reason could be that the trap tilted over due to slight water movements or bottom roughness and it was not any more vertically positioned during the exposure time (Fig. 7).

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Fig. 6 . Inclination of sediment vessel (after Blømqvist & Håkanson, 1981)

In conclusion, a pair sedimentation traps has been submerged to the lake bottom in each enclosure to measure the settling nutrient fluxes. The sedimentation traps were designed according to the established requirements. However, an experimental error, resulting in large deviations of the measurement of two parallel cylinders, seems to have taken place, e.g. because of the small diameter size of the enclosures. The data from the sedimentation traps have been statistically processed as described in 2.4. Data Analysis (p.67, this chapter).

2.4. Analyses

2.4.1. Temperature, pH, Dissolved Oxygen and Redox Potential

Depth profiles of temperature, oxygen, pH, electrode potential and conductivity were measured in situ at 0.5 m depth intervals with a H20® multiprobe (HYDROLAB) for each enclosure. The deepest point of the measurement was just above the water-sediment interface. After comparing the measured values, no stratification with respect to the pH was observed in the different enclosures, but only insignificant changes with depth. Therefore, pH and temperature are not calculated as mean or median value for the entire enclosure, but only the measurement at the depth of 6 m (water-sediment interface) is used (if not stated otherwise). Dissolved oxygen and redox profiles show the recorded changes in space and time. Detailed information for the pH, oxygen and redox profiles of all enclosures can be found in App. IV.

2.4.2. Total Phosphorus (TP) and Total Iron (TFe) Determination

Approximately 750 ml aliquot (hypolomnion water and suspended particulate matter) was taken from each sedimentation trap for analysis (TP and TFe) and transported in 1.5 - liter polyethylene bottles kept in cooling boxes and filtered on the same day (4 to 6 hours after the field sampling) through weighted pre-combusted (3h, 550°C) 47 mm Whatmann (GF/F) filter. The filters were dried in an oven (105°C [illustration not visible in this excerpt] 0.5°C) overnight and reweighed to determine the total amount of the particulate matter in each trap ([g] DW). For determination the amount of organic settling particulate matter the filters were combusted at 550°C [illustration not visible in this excerpt] 50°C for 3 hours and weighed again. The aliquots of every two sedimentation traps which had been submerged into the same enclosure were mixed together and left aside to settle down. The dry sediment for the determination of TP and TFe was obtained after carefully disposing the clear supernatant and drying the sediment slurry at 60°C for 12 h. The concentrations of soluble reactive phosphorus (SRP) were determined photometrically with molybdenum-blue method and ascorbic acid as reducing agent at the wavelengths 880 nm (Murphy & Riley 1962) after digestion of the dried and ground sediment with 0.5M HCl and potassium peroxodisulfate (K2S2O8). The concentrations of TFe were determined photometrically as well at the wavelength of 511 nm after digestion with 0.5M HCl.

The TP and TFe extraction schemes are outlined in App. VI.

2.4.3. Nitrogen and Carbon Determination

Approximately 100 ml of each well stirred water sample with settled particulate matter was transferred into a small plastic bottle and frozen for further analyses of TN, TC, TOC, DIC performed IR-spectrometrically by DIMATEC-TOC 100 (Department of Wastewater and Sewage Engineering, BTUC). The value for each single sedimentation trap was recorded, but the mean value for every two traps in each enclosure is used in the results. Data is statistically processed as described in the section Statistical Methods of Data Analysis (p. 65 ).

The organic settling particulate matter was recorded as loss on ignition (LOI [%]). Detailed information is available in App. II.

The sedimentation rate for each single sedimentation trap was recorded, and the mean value for each enclosure is used in the results. This gives the opportunity to have two measurements in the same conditions and to calculate the standard deviation as well as the relative error between each two measurements.

2.4.4. Rates of Sedimentation

The sedimentation rate was determined as mentioned elsewhere (e.g. Kalff, 2002).

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2.5. Statistical Methods of Data Analysis

A set of samples was taken during sampling. Each enclosure was equipped with two sedimentation traps for the purpose of obtaining more reliable data. The rate of sedimentation in an enclosure was estimated as mean value of the sedimentation rates measured within the respective sedimentation traps: [illustration not visible in this excerpt] with estimated standard deviation [illustration not visible in this excerpt] (e.g. Peiffer & Pecher, 1997).

Due to the various sampling errors (partially mentioned in Chapter II, 2.2. Sedimentation Traps), the sedimentation rates measured from a pair of traps from the same enclosure could happen to be significantly differed from each other.

Sampling errors were estimated with the use of the coefficient of variation (%) [illustration not visible in this excerpt] (Håkanson & Jansson, 1983), where [illustration not visible in this excerpt] is the standard error. In the case [illustration not visible in this excerpt] the mean value was rejected and the values from the two sedimentation traps were compared with the values from enclosures with similar treatment. On this consideration as summarised in Table 7, if the relative error between two traps from the same enclosure exceeds 30%, only the more reasonable of them was considered and the other one neglected.

Table 7. Error estimation in the spreadsheets

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Chapter 3 Results and Discussion

3.1. Fluxes of Particular Matter

3.1.1. Sedimentation in the Lake and Controlling Enclosures (EC)

Sedimentation in Lake Grünewalde vary seasonally and with depth. The suspended particulate matter in the lake epilimnion and hypolimnion settling down towards the sediment has been recorded to be respectively 2 and 3 g/(m2d) in the summer (beginning of August 2002) and lower in spring of about 0.7 to 1 g/(m2d) (e.g. in April 2002) (ref. to Fig. 2).

The sedimentation rate of seston in the lake at the depth of 6 m[illustration not visible in this excerpt]0.3 m in near proximity to the enclosures was 1 g [illustration not visible in this excerpt] 0.2 g/(m2d) (Fig. 8) during the course of the experiment and this value is in a good agreement with the average expected sedimentation rate for lake Grünewalde at that time period (Fig. 2).

The settling matter was 1 g/(m2d) in the controlling enclosures, quite the same to that measured in the lake itself. The sedimentation rate in the controlling enclosures is slightly enhanced in comparison to the lake as a result of less water movements (vertical and horizontal) in the smaller enclosure volume.

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Fig. 7. Rates of seston flux measured in the lake (n = 2) and controlling enclosures
(n = 6) in the time course of the enclosure experiment

It is also worth mentioning that a few days before the last recorded sampling a storm had taken place over the lake and there were evidences of disturbed traps not to mention the fact that a few of them were not at all found. Despite the fact that the sedimentation rates from the 76th day (28.10.) were recorded and they are shown graphically, the question how reliable they are remains. Therefore, the sedimentation rates of this last sampling (28.10.) are not used in the results discussion, but only the different component concentrations in the seston (e.g. TP, TN, TC, TFe) were considered.

3.1.2. Sedimentation in the Neutralised and Fertilised Enclosures (ENeP)

The addition of soda ash to the acidic water column in the enclosures on the 14th August resulted in a rapid neutralization and precipitation of newly-formed carbonates. Seven days after the addition of the neutralizer (21.08.) the settling matter, reddish and rather fluffy, was recorded to be extraordinary high in the range of 350 - 450 g/(m2d) caused by settling down of the precipitates (data shown graphically in App. V).

The combination of neutralizer and fertilizer (ENeP) increases the nutrient turnover in the water column and respectively the sedimenting flux in the range of about 2 to 3 times compared to the mean value for the lake (1 g ± 0.2 g/(m2d)) as shown at Fig. 9. The direct addition of phosphorus to the water column caused an algal mass response of 350 µg Chorophyll per litre and Secchi depth of 0.5m (Totsche et al., 2003, Fyson & Gelbrecht, 2004). That enormous increase in the algal biomass (Fig. 10) was dominated by green alga Scenedesmus sp. not found in the lake before (Totsche et al., 2003) but the species was observed only in the first weeks of the experiment after which its population collapsed. Not surprisingly the highest sedimentation rate of 3.1 ± 0.3 g/(m²d), n = 4 was observed app. 30 - 40 days after the beginning of the experiment as a result mainly of sinking of the dead phytoplankton biomass. After the collapse of the algal population the gross sedimentation decreased to 2.5 ± 0.08 g/(m2d) on the 63rd day.

The sedimentation in the neutralised and fertilised enclosures (Fig. 9, solid line) remained relatively enhanced for over two months during the experiment although the last measurement shows a tendency of decrease. Its further development would be quite interesting but the time frame here does not allow following it.

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Fig. 8. Rates of seston flux measured in the neutralised and fertilised enclosures (n = 6) during the course of the enclosure experiment

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Fig. 9 . Sample of sedimentation trap aliquot (app. 700 ml) from a neutralised fertilised enclosure taken on the 28th day of the experiment Fehler! Textmarke nicht definiert.

3.1.3. Fehler! Textmarke nicht definiert.Sedimentation in the Neutralised Enclosure (ENe)

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Fig. 10 . Rates of seston flux in the neutralised enclosure (n = 2) in the time course of experiment

The sedimentation rate in the enclosure where only neutralizer (Na2CO3) was added (ENe) remained close to that of the lake. It increased in the first 25 days up to the maximum value of 3.3 g/(m2d) after which started constantly decreasing but still remained slightly higher than the average for the lake (1 g/(m2d)) most probably as a result of more stable water column (Fig. 11). The max. value of 3.3 ± 0.4 g/(m2d) suggests being a consequence of the elevated pH (5.5), TOC (7 mg/l) and TIC (1.5 mg/l) in the enclosure at that time (ref. To Fig. 18)and the drop down of pH and C concentrations in the enclosures resulted in immediate decrease of the settling particular matter.

3.1.4. Sedimentation in the Fertilized Enclosure (EP)

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Fig. 11. Rates of seston flux in the fertilized enclosure (n = 2) in the time course of the experiment

The sedimenting flux in the phosphorus amended enclosure was highest (5.5 ± 0.4 g/(m2d)) 14 days after the addition of phosphates, after which gradually decreased to 2.0 - 2.5 g/(m2d) but remained slightly higher than the average for the lake.

This initially enhanced sedimentation rate followed by a decrease can be explained with the sensitivity of the element phosphorus. Direct addition of phosphate to the water column resulted in immediate algal bloom which, however, did not remain stable for longer time due to low pH and limiting TN and DIC concentrations. Thus, sedimentation rates rapidly decreased after the crack of the population but remained higher than the average for the lake.

3.1.5. Sedimentation in the Organic Matter Amended Enclosure (EPo)

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Fig. 12. Rates of seston flux in the organic matter amended enclosures (n = 4)
in the time course of the experiment

Enhanced sedimentation was also observed after the addition of organic matter (EPo). Although the sedimentation rate was quite similar to that of the lake at the beginning of the experiment significant increase at the rate of 3.5 ± 0.8 g/(m2d) was recorded on the 28th day (10.09.) of the experiment. The sedimenting flux remained in the range of 2.9 - 3.5 g/(m2d) until the end of the experiment.

The enhanced sedimentation rates in the organic matter amended enclosures after the first month of the experiment sequels the decomposition of the organic amendments and the elevated DIC concentrations of the water column which reached the value of 2.5 ± 0.5 mg/l for that period (Fig. 18).

3.1.6. Sedimentation in the Neutralised and Organic Matter Amended Enclosures (ENePo)

Quite similar to the above discussed observations were made in the enclosures where additionally to the organic matter, a neutralizer (Na2CO3) was added (ENePo). There the sedimentation rate seems to have large deviations from its track. However the average sedimentation is estimated in the range 4 - 6.0 g/(m2d), (excluding the settling flux measured on the 28th Oct which was measured after the storm on the lake). The maximum rate 6.6 g/(m2d) [illustration not visible in this excerpt] 0.15 g/(m2d) was measured 22 days after potatoes had been added; thereafter the primary flux decreased to 3.0 [illustration not visible in this excerpt] 1.0 g/(m2d) but remained relatively high, of about 4.0 g/(m2d) at average.

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Fig. 13 . Rates of seston flux in the neutralized and organic matter amended enclosures (n = 4) in the course of the experiment

The combination of neutraliser and potatoes enhanced at most the sedimentation processes in the water column. After the degradation of the organic matter near to the anoxic conditions occurred at the water-sediment interface (DO 32% at 0 m, s. App. IV) and microbial neutralization took place where the iron - and sulphate - reducing bacteria partially removed acidity as well as sulphate. Fig. 15 shows an aliquot sample from this enclosure 28 days after the beginning of the experiment where the traces of black particles (FeS2) result from sulphate reduction.

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Fig. 14 . Sample of sedimentation trap aliquot (app. 700 ml) from a neutralised organic matter- amended enclosure taken on the 28th of the experiment

In conclusion, most enhanced sedimentation processes were observed in conditions of direct phosphate addition at the rate of 0.5 mg/l; however, sedimentation was relatively enhanced as long as and very shortly after phosphate was supplied to the water column as a result of algal bloom. Not considering this initial peak (from a single measurement) due to the direct addition of phosphate the rates of sedimentation can be estimated as follows:

Seston flux (ENePo) > Seston flux (EPo) > Seston flux (ENeP) > Seston flux (EP) > Seston flux (ENe) > Seston flux (EC) [illustration not visible in this excerpt] Seston flux (Lake)

3.2. Chemical Changes in the Aqueous Phase and Nutrients Turnover in the Time Course of the Enclosure Experiment

3.2.1. Vertical Distribution of pH and Dissolved Oxygen in the Lake

Physical-chemical factors such as temperature, pH, redox conditions (Eh) and oxygen content can regulate the biologically mediated nutrient cycling, while at the same time the biological processes can change the physical-chemical environment. The water column of the dimictic Lake Grünewalde was entirely mixed and no stratification was observed up to the depth of 6m for the time period August - October 2002. The oxygen concentration was above 90% for the whole course of time and even mostly above 95% with the exception of the first week of Oct when it declined slightly to 90% (Fig. 16).

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Fig. 15 . Temperature [t° C] and oxygen distribution [% DO] profiles in Lake Grünewalde for the time period mid August - mid Oct 2002

The water column was supersaturated with dissolved oxygen at the depth of 6 m on the 28 August (appr. 101 - 103% O2, s. App. IV for details) but this is believed not to be caused by primary production but rather by hydrostatic pressure in the water column. Furthermore, the pH and redox potential of the lake remained relatively constant during the whole time period and did not change with depth; those parameters were in the range of respectively 3.0 ± 0.1 and 550 - 750 mV (data not shown graphically, but available in App. V).

3.2.2. pH Dynamics in the Course of the Experiment

The changes of the pH with time in all 12 enclosures and the lake during the experiment are shown at Fig. 17. The enclosures neutralized with soda ash (Na2CO3) added on 14th August have as expected elevated pH values although in some of the enclosures the neutralization had short-term effect. Only 4 out of 6 neutralized enclosures had pH values above 5.5 a fortnight after having been neutralized. Another 2 neutralized enclosures (ENeP, ENePo) had still higher pH values, pH 3.5 - 4.0, than that of the lake, but already in the acidic range.

About 30 days after the neutralization only 2 out of 6 enclosures had pH values higher that 5.5 and 49 days after the beginning of the experiment only one enclosure (ENePo) had pH above 6 (Fig. 17A). Those results confirm Evangelou’s statement (1995) that “neutralizers are known to be efficient as long as they are applied” as in some cases a non-soluble complex forms on the active surface of the substance or alternatively due to constant recharge with acidic groundwater.

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Fig. 16 . pH dynamics in all enclosures and in the lake during in the time course of the experiment

3.2.3. Changes in the Carbon Concentrations

Organic material which undergoes degradation in the water column is incorporated in the sediments, where it can be more or less preserved or subjected to further biochemical degradation. By autotrophic, reductive assimilation green plants convert inorganic carbon to organic forms, which settle down as allochthonous and/or autochthonous material. Reduced organic carbon provides the main source for heterotrophic organisms in the sediments.

The total carbon (TC) concentration in Lake Grünewalde was between 2 and 3.5 mg/l before starting the enclosure experiment (ref. to Table 4). Seven days after the specific additives were applied (27.08.) to the different enclosures, the TC concentrations were slightly higher in the neutralized enclosures - 5.1 mg/l TC compared to those where no soda ash was added – 4.2 mg/l TC (Fig. 18A). That is evident to be a result of the increased local pH and the introduced inorganic carbon that has changed the buffering system from Fe3+ to carbonate/bicarbonate buffered water in the neutralized enclosures as the dissolved TFe decreased from 10 mg/l to < 0.5 mg/l, and Kb, 4.3 from 1.9 mmol/l to 0.5 mmol/l. This observation is supported by the sharp increase in the DIC where the maximum values were measured on the 27.08. and they were 2 to 4 times higher for the chemically neutralized enclosures, 2 - 3.8 mg/l DIC, in comparison to the untreated lake column, 0.9 - 1.0 mg/l (Fig. 18C).

TC max values of 55 [illustration not visible in this excerpt] 5 mg/l were observed 28 days (10.09.) after the additives were applied to ENeP (Fig. 18A), of which about 55% is organic carbon (TOC) and only 9% is dissolved inorganic carbon (DIC). This observation suggests that the effect of soda ash has diminished, but the initially high TIC and the addition of nutrients (P) has enhanced the primary production and the settling flux of autochthonous organic material (Fig. 18A, C).

One month after the addition of nutrients (18.09.) the TOC concentrations were observed to decline again suggesting that the organic decomposition products have been utilized and converted to DIC (Fyson et al., 1998, Fyson & Gelbrecht, 2004). This hypothesis is completely supported by the fact that DIC was 3 [illustration not visible in this excerpt] 0.5 mg/l in the neutralized and phosphate amended enclosures whereas it is of about 1.5 [illustration not visible in this excerpt] 0.5 mg/l in the controlling enclosures (Fig. 18B, C).

No relationship was found between the concentrations of TC, DOC, DIC or LOI and the gross sedimentation fluxes of settling particles.

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Fig. 17 . Dynamics of TC, DIC and DOC in the time course of the enclosure experiment

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Fig. 18. Dynamics of TC, DIC and DOC in the time course of the enclosure experiment (continued)

3.2.4. Changes in the Nitrogen Concentrations

When nutrients enter a too much enriched lake, they are likely to be taken-up by phytoplankton (except during winter). The algae will use these elements for a while but after cell lysis and death, large parts are released as o-phosphate and ammonia. Usually only small quantities of both are not mineralized and are finally incorporated into the sediments.

TN in lake Grünewalde had mean values of 2 [illustration not visible in this excerpt] 0.5 mg/l (Table 4) which is also confirmed by the measured concentrations in the controlling enclosure (Fig. 19). As already discussed the direct addition of phosphate into the neutralized water column resulted in short term bloom of blue-green algae able to reduce nitrogen. Thus, TN concentration in those enclosures increased four times (8 ± 2.5 mg/l) in the first 20 days in comparison to its initial state after nitrogen fixation in the presence of increased algal population. After the collapse of the algal population, the TN decreased but remained higher than the average for the lake until the end of the experiment.

The combination of potatoes and neutraliser (EPoNe) resulted initially in more rapid increase of the TN compared to the enclosures where only potatoes were applied (EPo); however, on the 35th day the TN concentrations in the acidic and neutralized water had very similar values (2.5 - 3 mg/l), implying that enhanced bacterial fixation contributed to the initial TN increase in the neutralized water but after the crack of the algal population TN concentrations were likely due to break down of proteins in the decomposing potatoes (Fyson et al., 1998).

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Fig. 18. Dynamics of total nitrogen contents in the time course of the enclosure experiment

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Fig. 19. Dynamics of total nitrogen contents in the time course of the enclosure experiment

Correlation (Fig. 20) was found between the TN concentrations and the sedimentation rates but only for the organic matter amended enclosures. No correlations were found between the TN and the sedimentation rates in the other component specific enclosures or with any other chemical parameter. The dependence found suggests that N could have been the limiting factor in those enclosures. Because of the time-limited experiment and the restricted number of sampling (n = 13, r2 = 0.61) no further evidences supporting this hypothesis are available.

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Fig. 19. Sedimentation rate dependence on TN in the organic matter treated enclosures,
n = 13, r² = 0.61

3.2.5. Changes in the Total Phosphorus Content

Phosphorus cycling between lake sediments and water is most often associated with the release or retention of phosphorus coupled to reduced or oxidized forms of iron found by Mortimer (1941, 1942) (cited in Forsberg, 1989). So far it has been demonstrated that the interactions are more complex (Friese et al., 1998). It was found out that recycling of phosphate from the sediment is favoured by low redox- potential, high pH and high microbiological activities and it takes place under anoxic as well as under oxic conditions (Boström, 1982, Boström et al., 1982).

The low SRP concentrations in the AML have been reported elsewhere and the TP in lake Grünewalde determined from its settling seston was found to be at average 1 ± 0.5 mg TP/g DW (ref. to Fig. 2). The gross P sedimentation in the lake and in the controlling enclosures confirms these measurements (A).

The addition of phosphate resulted in TP content of 18 mg TP/g DW, app. 17 times higher than that of the lake in the first 13 days (Fig. 21A). After another 7 days it, however, decreased to 2 mg TP/g DW, slightly above the average for the lake. Assuming that the enclosure was not damaged, this implies that the phosphate added to the water column at the rate of 0.5 mg/l was rapidly assimilated and settled down as (in)organic authochtonous/ allochtonous material or was fixed in Fe/Al complexes and remained immobilized in the sediments (Fig. 22). This suggestion is confirmed by enhanced sedimentation rate of 3.5 - 4 g DW/(m2d) at that time.

The addition of phosphate and soda ash resulted in maximum TP content of 8 [illustration not visible in this excerpt] 1 mg TP/g DW in the first 30 days of the experiment (Fig. 21A) and it coincides with the enhanced sedimentation rate in that enclosure for the same time period after which the TP decreased gradually but still remained higher (4 mg TP/g DW) than that of the lake (1 [illustration not visible in this excerpt] 0.5 mg TP/g DW).

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Fig. 21 . Major mechanisms of phosphorus deposition
(adopted from Boström, 1982)

Neutralized and acidic potato - treated enclosures had TP concentrations of 1.5 [illustration not visible in this excerpt] 0.5 mg/g DW 13 days after the beginning of the experiment. In the following 20 days the TP in the neutralized potatoes - treated enclosures increased twice to the value of 3 mg TP/g DW. The TP content in the enclosures where only potatoes were added did not change significantly and remained quite similar to that of the lake.

3.2.6. Changes in the Total Iron Content

To investigate the effect of the organic matter addition on the microbial activity, total iron concentrations in the potato amended neutralized and acidic enclosures (ENePo, EPo), and in the controlling enclosures (EC) were measured and the results are graphically shown at Fig. 23.

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Fig. 22. Total Fe in the potatoes amended neutralized and acidic and controlling enclosures^[4]

The addition of soda ash increased the pH in the water column to appr. 5.5 - 6 and resulted in the formation of FeCO3 and Fe(OH)3 that rapidly settled down in the form of reddish precipitate. Therefore, the steady increase of the TFe (Fig. 23, solid line) in the neutralized potato - amended enclosures is most probably due to iron reduction. This suggestion is supported by the establishment of pH gradient (Fig. 25) in the water column as well as by the redox potential profile of the enclosure (s. App. IV), where both imply reductive conditions favouring microbial activity.

TFe in the controlling enclosure remained relatively constant.

3.3.Nutrient Fluxes to the Sediments

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Fig. 23. Phosphorus flux during the course of the experiment

Phosphorous flux in the lake was estimated to be 0.5 mg TP/(m²d) and it was confirmed by the measurements in the controlling enclosures. The addition of neutralizer altered the TP flux enormously.

The phosphorus flux in the neutralized phosphate treated enclosures was relatively enhanced in the first 25 days after which decreased but remains in average 10 mg TP/(m²d), which makes up about 20 time higher downwards P flux than the average for the lake (Fig. 24).

The max P flux in the neutralized organic matter amended enclosures (ENePo) was 9 mg TP/(m²d). This value decreased slightly but remained stable with time in the range 6 - 8 mg TP/g DW.

The highest P flux was observed after direct phosphate addition (EP). In the first two weeks the P flux was extraordinary high almost 180 times higher than that of the lake. However it had very unstable behaviour and rapidly decreased to stay as high as the average for the lake.

3.4. Sedimentation and Water Quality

The addition of organic matter, neutralizer and/or fertilizer alone or in different combinations enhanced the the flux of particular matter to the sediments (Figs. 8-14) and thus modified the water chemistry(Fig. 17-19) and the sediment microbiology(Fig. 15). Results from 2 enclosures (ENePo) showed some heterogeneity but confirmed the establishment of (although not very steep) pH gradient (Fig. 25) with depth. Similar studies based on biotechnological modification of the sediment water interactions (Koschorreck et al. 2002, Potthoff 2002, Fyson et al. 1998) measured such particularly steep gradient, e.g. from 3.5 to 5.5 (Koschorreck et al. 2002), but in the upper 5 mm of the sediment. Thus the experiment shows that such increasing with time pH gradient can be also easily and for relatively short time (in appr. 80 days) be established in the sediment overlying water column.

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Fig. 24 . Detailed pH profile of the neutralized organic matter amended enclosure

Of course one should keep in mind that the water was initially neutralized to the pH of 5.5 (App. IV) which assisted the quick and significant increase in the pH form pH 5.5 to pH 6. Additionally, the underlying sediment was observed to have black particles (Fig. 15) probably caused by precipitated FeS. Such conditions result in permanent increase of the TP (Fig. 21B) and DIC (Fig. 18C) as well as in the primary production (Fig. 10; Fyson et al. 1998, Fyson & Gelbrecht 2004) which in terms of relatively enhanced sedimentation (in the range 4 - 6.0 g/(m2d) at Fig. 14) increases the overall organic matter in the water column and maintains self-sustaining biochemical cycling. Referring further to the question of the water quality in the process of lake maturation, Kleeberg & Grueneberg (2005) using the bentic Fe:P atomic ratio found out that increased C concentrations leading to SO42- reduction and iron sulphide formation alter the bentic Fe:P atomic ratio and thus a lake of similar kind might become liable to euthophication during its aging.

Chapter 4 Conclusions

The flux of particular matter to the sediments in the acidic mining lake Grünewalde is generally quite low in the range of 1 mg DW/(m2d) and it was mostly influenced by the addition of C source and P.

The considerable presence of P in the water column simulated rapid eutrophication and thus deposition or authochthonous predominantly organic matter to the sediment due to enhanced primary production.

However, considerable differences in the temporal pattern of seston flux were observed in the neutralized and in the acidic water column after the addition of phosphate. The initially increased to the pH 6 enclosure water developed flux of particulate matter that remained relatively stable at the range of 3.1 mg DW/(m2d). The pH also remained in the neutral range as long as the settling flux was high and it decreased with the decrease of the gross sedimentation. On the contrary, the sedimenting flux was relatively enhanced only in the first 30 days after the addition of phosphate into the acidic water column. Most interesting behaviour was observed in the case of the P elemental flux, where P was very rapidly deposited to the sediment and retained there.

Additionally, the elemental phosphorus flux to the sediment was found out to be relatively constant and even slightly increasing at the end of the experiment in the case of organic matter addition to the neutralised water column (ENePo). The additional carbon to the already neutral water resulted in iron sulphide formation and in decreased Fe and Al concentrations due to complexing/ precipitation, which in terms enhanced the water quality in this enclosures.

This study revealed the importance of seston particles circulation in the water column and the ways of formation of bottom fluxed during experimental treatment for lake remediation.

Chapter 5 Summary

The work reports on the sedimentation processes and refers to the nutrients turnover estimated in terms of the gross sedimentation in an acidic water body during experimental treatment for lake remediation. Chemical and biological neutralisation have been applied in situ to increase the alkalinity in the acidic lake.

Sedimentation traps meeting the established technical specifications have been used to measure the daily primary fluxes after the addition of component specific amendments to the enclosures. Methodological and logistic difficulties with trap measurements (resuspension, mineralisation, fouling, over- /underestimation, etc.) have been reported and were also encountered in this case. In order to have most accurate and realistic model of the lake processes, raw data has been statistically processes as described elsewhere.

The chemical and biological parameters measured/ analysed in order to trace the processes taking place as a result of different amendments were pH of the water column, dissolved oxygen and redox potential, carbon and nitrogen concentrations, phosphorus and iron content.

The study confirms that gross sedimentation is a highly sensitive process that responds to the slightest chemical or biological change of the water column and can be of great importance in studying most environmental problems in lakes, such evaluating the distribution of chemical species, studying biochemical cycles, determining water quality. It was found that most important for enhanced primary flux in the acidic water bodies is the addition of carbon source. Only with that type of treatment increasing with time sedimenting flux was recorded.

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Appendices

Appendix I

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^[1] Statistik der Kohlenwirtshaft e.V.

^[2] Statistik der Kohlenwirtshaft e.V. http://www.kohlenstatistik.de/home.htm last accessed 15.08.2005

^[3] numerous iron-oxidizing bacteria and organisms have been isolated or detected in acidified environments but A. ferroxidans is often used as a model organism for low pH iron oxidation

^[4] The pH values on Fig. 23 refer to the ENePo (neutralised potato-amended enclosure) (solid line)

• College: Brandenburg Technical University Cottbus
• Grade: 1,7
• Author: Iglika Gentcheva (Author)
• Publication Year: 2006
• Pages: 75
• Catalog Number: V111578
• ISBN (eBook): 9783640096282
• ISBN (Book): 9783640120628
• Language: English
• Tags: Sedimentation, Processes, Nutrients, Turnover, Acidic, Water, Bodies, Experimental, Treatment, Lake, Remediation