VOLATILE SULFUR COMPOUNDS IN COASTAL ACID SULFATE SOILS, NORTHERN N.S.W
Andrew Stephen Kinsela
A thesis submitted in fulfilment of the requirements for the degree of
Doctor of Philosophy
School of Biological, Earth & Environmental Sciences THE UNIVERSITY OF NEW SOUTH WALES, AUSTRALIA
2007
DECLARATION
ORIGINALITY STATEMENT
‘I hereby declare that this submission is my own work and to the best of my knowledge it contains no materials previously published or written by another person, or substantial proportions of material which have been accepted for the award of any other degree or diploma at UNSW or any other educational institution, except where due acknowledgement is made in the thesis. Any contribution made to the research by others, with whom I have worked at UNSW or elsewhere, is explicitly acknowledged in the thesis. I also declare that the intellectual content of this thesis is the product of my own work, except to the extent that assistance from others in the project's design and conception or in style, presentation and linguistic expression is acknowledged.’
Signed …………………………………………………
Date ……………………………………………………
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ACKNOWLEDGEMENTS
There are numerous people who have assisted me throughout the course of my thesis. I therefore want to take this opportunity to thank a few of those who contributed appreciably, both directly and indirectly.
First of all, I would like to express my heartfelt gratitude to my supervisor, Associate Professor Mike Melville. Mike’s initial teachings as part of my undergraduate studies first sparked my interest in soils. Since then his continued enthusiasm on the subject has helped shape the way I approach my own work. His dedicated support on both an intellectual and personal level has provided much of the driving force behind the completion of my thesis.
The group of academics surrounding me throughout my candidature has made completing my project that little bit easier and a lot more enjoyable. Dr Ben Macdonald from the Australian National University (ANU) has been an excellent source of support, particularly regarding the friendship and assistance provided during fieldwork missions, but also with the numerous discussions about the direction of my thesis. Along with Dr Macdonald, the unsurpassed expertise in micrometeorological techniques provided by Dr Tom Denmead (CSIRO Land & Water) made large sections of my project possible. Initial meetings and advice within the early stages of my thesis with Professors Ian White (ANU) and David Waite (UNSW) largely shaped the direction my thesis, of which I am greatly appreciative.
Being part of a school that was not all that encouraging to the research undertaken by my supervisor I am also indebted to the support provided by my fellow acid sulfate soil doctoral researcher at UNSW, Jason Reynolds. Our frequent discussions on aspects of our projects and general philosophising were both a valuable asset to the advancement of my project as well as a welcome diversion from it.
I also wish to thank my fellow PhD students that I have worked with over the course of my project, again particularly for their assistance in the field and laboratory; Mira Dürr
Acknowledgments v
and Richard Reilly from ANU and Dr Jodie Smith from UNSW. Similarly, those students who were not necessarily involved with my project (either in the former School of Geography or within BEES) were also important in preserving my sanity.
The technical support provided, especially by Dorothy Yu, but also Irene Wainwright, Dr Gautam Chattopathy and Dr Richard Stuertz (all of UNSW) was also much appreciated; along with the administrative support provided almost single-handedly in BEES by Dr Louise Mazarolli.
Finally, I could not have completed my thesis without the love and support that I have received from my family; Mum, Dad and my sister Rebecca. I also particularly want to thank my girlfriend Ana Markovic, who has (foolishly?) stuck by me for the entire duration of my PhD, whilst undertaking the same painful experience in completing her own doctorate. The highs and lows that we have both experienced together in our respective fields will never be forgotten (no matter how much electro-shock therapy is endured).
Acknowledgments vi
PUBLICATIONS / CONFERENCES
Much of the work has been published in refereed journal articles and conference proceedings, or presented at national and international conferences, listed below;
REFEREED PUBLICATIONS Kinsela AS, Melville MD (2004) Mechanisms of acid sulfate soil oxidation and leaching under sugarcane cropping. Australian Journal of Soil Research 42, 569- 578.
Macdonald BCT, Smith J, Keene AF, Tunks M, Kinsela A, White I (2004) Impacts of runoff from sulfuric soils on sediment chemistry in an estuarine lake. Science of the Total Environment 329, 115-130.
Dürr M, Kinsela A, Macdonald BCT, White I (2004) Influence of land use on the emission of sulfur dioxide from acid sulfate soils. In 'Supersoil 2004: Program and Abstracts for the 3rd Australian New Zealand Soils Conference', University of Sydney. (Ed B Singh). (The Regional Institute Ltd)
Kinsela AS, Reynolds JK, Melville MD (2007) Agricultural acid sulfate soils: a potential source of volatile sulfur compounds? Environmental Chemistry 4, 18-25.
BOOK CHAPTERS Macdonald BCT, White I, Heath L, Smith J, Keene AF, Tunks M, Kinsela A (2006) Tracing the outputs from drained acid sulphate floodplains to minimize threats to coastal lakes. In ‘Environment and Livelihoods in Tropical Coastal Zones: Managing Agriculture- Fishery-Aquaculture Conflicts’ (Eds CT Hoanh, TP Tuong, JW Gowing, B Hardy). (CABI, Wallingford, U.K.)
PUBLICATIONS IN PROCESS Kinsela AS, Macdonald BCT, Melville MD, White I, Reynolds JK, Denmead OT (In prep.) Field measurements of sulfur gas emissions from an agricultural acid sulfate soil, Eastern Australia.
Macdonald BCT, Denmead OT, White I, Melville MD, Kinsela AS (In prep.) Harmful gas emissions from coastal soils?
Reynolds JK, Kinsela AS, Melville MD, White I, Macdonald BCT, Waite TD (In prep.) Iron, sulfur and manganese geochemistry in a coastal lowland sulfidic sediment.
Reynolds JK, Kinsela AS, Macdonald BCT, Melville DM (In prep.) Acidity and nitrate relationships during a flood event, Tweed River floodplain, NSW, Australia.
Publications / Conferences vii
CONFERENCE PROCEEDINGS Quirk R, Melville MD, Waite TD, Desmier R, Kinsela AS, Reynolds J, Smith J, Keene A, Macdonald BCT, White I, Donner E (2005) Acid sulfate soils: their drainage, oxidation, and best management. In ‘Proceedings of American Association for the Advancement of Science First International Conference’, New Orleans, U.S.A.
Denmead O, Macdonald B, White I, Reilly R, Kinsela A, Melville M, Griffith D, and Bryant G (2005) Acid sulfate soils: a new source of sulfur and greenhouse gases. In: Amstel, A. 4th International Symposium on non-CO2 greenhouse gases (NCGG-4), science, control, policy and implementation, 4-6 July 2005, Utrecht, Netherlands. Rotterdam, Netherlands: Millpress Science Publishers: 169-177.
WORKSHOP PROCEEDINGS Denmead OT, Macdonald BCT, White I, Reilly R, Kinsela A, Melville MD, Griffith DWT, Bryant G. (2006) Impacts of Acid Sulfate Soils on Air Quality. In ‘Proceedings of the Workshop on Agricultural Air Quality: State of the Science, Potomac, U.S.A., June 5-8’, (Eds Aneja VP et al.) pp. 575-582.
CONFERENCE PRESENTATIONS Kinsela AS, Melville MD, Macdonald BCT, White I, Denmead OT (2004) Sulfur gas emissions from coastal acid sulfate soils. Supersoil 2004: Proceedings of the 3rd Australian and New Zealand Soils Conference, University of Sydney, Australia, 5-9 December 2004.
Kinsela AS, Reynolds JK, Melville MD (2006) Acid sulfate soils: a source of volatile sulfur compounds? Soil Science Solving Problems Proceedings of the 4th Australian and New Zealand Soils Conference, University of Adelaide, Australia, 3-7 December 2006.
CONFERENCE ABSTRACTS/PRESENTATIONS (presented by others) Denmead OT, Macdonald BCT, Kinsela A, White I, Melville MD (2004) Emission of sulfur gases from acid sulfate soils. 26th Conference on Agricultural and Forest Meteorology, Vancouver, Canada.
Donner E, Keene A, Kinsela A, Smith J, Desmier R, Reynolds J, Macdonald BCT, White I, Tunks M, Quirk R,. Melville MD (2004) Is artificial drainage the cause of acid sulfate soil profile oxidation? Supersoil 2004: Proceedings of the 3rd Australian New Zealand Soils Conference, University of Sydney, Australia, 5-9 December.
Kinsela AS, Melville MD, Macdonald BCT, White I, Denmead OT (2005) Sulfur gas emissions from coastal acid sulfate soils. 1st International Conference on Environmental Science and Technology, New Orleans, U.S.A.
Publications / Conferences viii
Macdonald BCT, White I, Åström ME, Keene AF, Reynolds J, Kinsela A, Österholm P (2005) Acid and metal discharges from acid sulfate soils, eastern Australia and Finland. 1st International Conference on Environmental Science and Technology, New Orleans, U.S.A.
Denmead OT, Macdonald BCT, White I, Reilly R, Kinsela A, Melville MD, Griffith DWT, Bryant G (2005) Acid sulfate soils: a new source of sulfur- and greenhouse-gases. Fourth International Symposium on Non-CO2 Greenhouse Gases (NCGG-4): Science, Control, Policy and Implementation (Ed. A. van Amstel), 4-6 July 2005, Utrecht, The Netherlands.
Publications / Conferences ix
ABSTRACT
The cycling of biogenic volatile sulfur compounds (VSCs) within marine and terrestrial ecosystems has been shown to play an integral role in atmospheric chemistry; by influencing global climate change through the creation of cloud condensation nuclei and controlling acid-base chemistry; as well as influencing sediment chemistry including the interactions with trace metals, particularly regarding iron sulfide formation. Despite this, the examination of VSCs within Australian coastal acid sulfate soils (ASS) is an unexplored area of research. As ASS in Australia occupy an area in excess of 9 M ha, there is a clear need for a greater understanding of the cycling of these compounds within such systems.
This thesis looks at the emissions and concentrations of several VSCs within ASS on the east coast of Australia. This includes an initial study of the general porewater and sedimentary characteristics of two locations; Blacks Drain, a tributary of the Tweed River which is intensively cultivated by sugarcane, and Cudgen Lake Nature Reserve, a relatively undisturbed ASS, both of which contain > 1.5 % pyrite in the sediment profile.
Initial measurements of sulfur dioxide (SO2) were made using passive diffusion samplers, which showed that a range of different ASS were emitting the compound. This was followed by two separate detailed field-based studies looking at the concentrations and fluxes of both SO2 and hydrogen sulfide (H2S) at the Blacks Drain site using flux-gradient micrometeorological techniques in conjunction with ultraviolet fluorescent analysers. The results indicated that this agricultural ASS was a substantial -2 -1 source of atmospheric H2S (0.036 - 0.056 g S m yr ), and SO2 (0.095 - 0.31 g S m-2 yr-1) with flux values equating to many other salt- and freshwater marshes, swamps and mudflats. The concentration and flux data also suggested that the ASS could be a continual source of H2S which is photo-oxidised during the daytime to SO2. Measurements of both compounds showed separate, inverse correlations to temperature and moisture meteorological parameters, indicating possible contributing and / or causal factors to their release.
Abstract xi
In an attempt to identify the concentrations of these and other VSCs within the soil profile at Blacks Drain and Cudgen, laboratory-based gas chromatography was used on
ASS samples in combination with solid-phase microextraction. Although SO2 and H2S were not discovered within the headspace samples, two other VSCs important in atmospheric sulfur cycling and trace metal geochemistry were quantified; dimethylsulfide (DMS) and ethanethiol (ESH), with in-situ concentrations of the two compounds exceeding 300 μg/L and 4 μg/L respectively.
The measurements of H2S, DMS and ESH are the first quantifications within Australian ASS, and these preliminary measurements indicate that they may be important for refining regional or local atmospheric sulfur budgets, as well as interpreting previous
SO2 emissions from ASS. Ultimately the work from this thesis enhances our understanding of the biogeochemical cycling of sulfur in ASS, and demonstrates that sulfur is in a complex equilibrium with the other elemental cycles (particularly C and N), of which, the formation (and release) of volatile compounds is an integral part.
Abstract xii
TABLE OF CONTENTS
DECLARATION...... iii
ACKNOWLEDGEMENTS ...... v
PUBLICATIONS / CONFERENCES...... vii
ABSTRACT...... xi
LIST OF TABLES...... xvii
LIST OF FIGURES ...... xix
LIST OF ABBREVIATIONS ...... xxiii
Chapter One: INTRODUCTION...... 1
1.1 PREFACE...... 1 1.2 CONCEPTUAL OVERVIEW ...... 1 1.3 THESIS OBJECTIVES ...... 3 1.4 THESIS OUTLINE ...... 3
Chapter Two: ACID SULFATE SOILS & SULFUR GASES BACKGROUND...... 5
2.1 ACID SULFATE SOIL BIOGEOCHEMICAL THEORY...... 5 2.1.1 Introduction...... 5 2.1.2 Iron Sulfide Formation...... 6 2.1.3 Iron Sulfide Oxidation...... 8 2.1.4 Oxidation Impacts ...... 10 2.1.5 Natural vs. Anthropogenic Oxidation of Acid Sulfate Soil Landscapes ...... 12 2.2 ACID SULFATE SOILS AND SULFUR GASES...... 17 2.2.1 Previous Research into Sulfur Gas Emissions from Acid Sulfate Soils...... 17 2.3 SULFUR GASES...... 18 2.3.1 Sulfur Dioxide ...... 19 2.3.2 Hydrogen Sulfide...... 20 2.3.3 Dimethylsulfide...... 22 2.3.4 Thiols...... 25 2.3.5 Other Volatile Sulfur Compounds ...... 27
Chapter Three: STUDY SITES BLACKS DRAIN & CUDGEN LAKE...... 29
3.1 INTRODUCTION...... 29 3.2 BLACKS DRAIN SITE...... 31 3.2.1 Background ...... 31 3.2.2 Methods...... 32
Table of Contents xiii
3.2.3 Results and Discussion ...... 36 3.3 CUDGEN LAKE NATURE RESERVE SITE...... 42 3.3.1 Background...... 42 3.3.2 Methods...... 43 3.3.3 Results and Discussion ...... 43 3.4 SUMMATION ...... 51
Chapter Four: PASSIVE GAS ANALYSIS - SULFUR DIOXIDE FIELD MEASUREMENTS...... 53
4.1 INTRODUCTION ...... 53 4.1.1 Passive Diffusion Samplers...... 53 4.2 MATERIALS AND METHODS...... 57 4.2.1 Study Sites...... 57 4.2.2 Methods: Passive Diffusion Samplers...... 61 4.3 RESULTS...... 62 4.3.1 December 2002...... 62 4.3.2 May 2003 ...... 63 4.3.3 Errors Associated with Passive Diffusion Samplers...... 67 4.4 DISCUSSIONS & PRELIMINARY CONCLUSIONS...... 69
Chapter Five: ADVANCED FIELD MEASUREMENTS HYDROGEN SULFIDE & SULFUR DIOXIDE...... 75
5.1 INTRODUCTION ...... 75 5.1.1 Micrometeorological Techniques ...... 76 5.1.2 Comparison with Alternative Techniques: Chambers ...... 78 5.2 MATERIALS AND METHODS...... 81 5.2.1 Study Sites...... 81 5.2.2 Gas Analysers ...... 82 5.2.3 Methods: Gas Sampling...... 83 5.2.4 Calculations...... 85 5.3 RESULTS 1 – NOV/DEC 2003...... 86 5.3.1 Soil Description ...... 86
5.3.2 SO2 and H2S Concentrations...... 87
5.3.3 SO2 and H2S Fluxes ...... 91 5.3.4 Micrometeorological Interactions ...... 96 5.4 RESULTS 2 – OCT/NOV 2005 ...... 101 5.4.1 Soil Description ...... 101
5.4.2 SO2 and H2S Concentrations...... 102
5.4.3 SO2 and H2S Fluxes ...... 102 5.4.4 Micrometeorological Interactions ...... 106
Table of Contents xiv
5.4.5 Statistical Analysis ...... 113 5.5 DISCUSSIONS & PRELIMINARY CONCLUSIONS ...... 118
5.5.1 H2S Measurements: Formation and Consumption Processes ...... 118
5.5.2 H2S Release Mechanisms ...... 122
5.5.3 SO2 Measurements ...... 129 5.5.4 Climatic Variable Interactions...... 132 5.5.5 Implications for Measurements ...... 135 5.5.6 Preliminary Conclusions...... 138
Chapter Six: LABORATORY-BASED SULFUR GAS MEASUREMENTS GAS CHROMATOGRAPHY...... 141
6.1 INTRODUCTION...... 141 6.1.1 Gas Chromatography...... 142 6.1.2 Solid-phase Microextraction ...... 143 6.2 MATERIALS ...... 147 6.2.1 SPME...... 147 6.2.2 Chromatography ...... 148 6.2.3 Sampling Containers ...... 150 6.2.4 Sampling Location...... 150 6.2.5 Extraction Procedure ...... 151 6.3 RESULTS ...... 152 6.3.1 Calibration ...... 152 6.3.2 Blacks Drain Samples ...... 153 6.3.3 Cudgen Lake Samples ...... 155 6.4 DISCUSSION & PRELIMINARY CONCLUSIONS...... 158 6.4.1 Dimethylsulfide...... 158 6.4.2 Ethanethiol ...... 164 6.4.3 Emissions of Volatile Sulfur Compounds to the Atmosphere ...... 168 6.4.4 Volatile Sulfur Compounds & ASS...... 171 6.4.5 Methodological Issues / Difficulties ...... 175 6.4.6 Combined Interpretations and Conclusions...... 179
Chapter Seven: DISCUSSION OF COLLECTIVE RESULTS AND CONCLUDING REMARKS...... 181
7.1 CONCLUDING REMARKS...... 181 7.1.1 Introduction...... 181 7.1.2 Summary...... 181 7.2 IMPLICATIONS FOR COASTAL LOWLAND ACID SULFATE SOILS ...... 186 7.2.1 Are Concentrations of Volatile Sulfur Compounds an Issue in Acid Sulfate Soils? ...... 186 7.2.2 Vegetation / Animal Impacts ...... 188
Table of Contents xv
7.2.3 Human Health Impacts ...... 191 7.2.4 Atmospheric / Climatic Impacts...... 194 7.2.5 Can the emissions from ASS be managed / decreased?...... 195 7.3 FURTHER RESEARCH...... 197
REFERENCES ...... 199
LIST OF APPENDICES...... 239
Table of Contents xvi
LIST OF TABLES
Table 2.1. Global atmospheric emission estimates for SO2...... 20
Table 2.2. Global atmospheric emission estimates for H2S...... 21 Table 2.3. Global atmospheric emission estimates for DMS...... 22 Table 3.1. Comparison of degrees of pyritisation between Cudgen Lake and other locations...... 46 Table 3.2. Measurements of porewater concentrations at Cudgen Lake which exceed levels set out by the ANZECC Guidelines, and the depths in the profile at which they were taken...... 50 Table 4.1. Descriptions of the landuse for the five sample locations – December 2002...... 58 Table 4.2. Descriptions of the landuse for the five sample locations – May 2003...... 60 Table 4.3. Average errors intentionally introduced to the passive diffusion sampling method...... 69
Table 5.1. Diurnal flux averages for SO2 and H2S for the 2003 sample period...... 96
Table 5.2. Correlation coefficients for SO2 and H2S concentration using Spearman’s Correlation...... 97
Table 5.3. Correlation coefficients for SO2 and H2S flux using Spearman’s Correlation...... 98
Table 5.4. Correlation coefficients for SO2 and H2S flux, restricted by ideal wind direction (80-160°), using Spearman’s Correlation...... 100
Table 5.5. Diurnal flux averages for SO2 and H2S for the 2005 sample period...... 105
Table 5.6. Correlation coefficients for SO2 / H2S concentration and flux using Spearman’s Correlation...... 113
Table 5.7. Variations in average biogenic H2S emissions from various wetlands/coastal marshes...... 137
Table 5.8. Variations in average biogenic SO2 emissions from available data...... 138 Table 6.1. Some examples of non-environmental SPME applications...... 143 Table 6.2. Some examples of environmental-based SPME applications...... 144 Table 6.3. Temperature parameters used in the GC / FPD method...... 149 Table 6.4. DMS porewater concentrations of this study compared to other published research...... 160 Table 7.1. VSC odour thresholds, short-term and long-term exposure effects...... 191
List of Tables xvii
LIST OF FIGURES
Figure 2.1. Conversion of DMSP in anoxic marine sediments...... 25 Figure 3.1. Locations of the two sampling sites on the NSW north coast; Blacks Drain and Cudgen Lake Nature Reserve...... 31 Figure 3.2. Schematic diagram of in-situ porewater samplers or ‘peepers’...... 35 Figure 3.3. Soil pH and redox values for a profile at Blacks Drain from November 2005...... 37 Figure 3.4. Chromium reducible sulfur (CRS) as a percentage (a.), reactive iron (b.), and degree of pyritisation (DOP) as a percentage (c.) for a soil profile at Blacks Drain from November 2005..... 38 Figure 3.5. Porewater total iron and manganese concentrations (a.), ferrous iron concentrations (b.), and sulfate concentrations (c.) for a peeper profile at Blacks Drain from May 2003...... 40 -2 Figure 3.6. pH - p relationship of the Blacks Drain samples overlain by S-O2-H2O (25°C, 10 M) thermodynamic stability diagram...... 41 Figure 3.7. Sampling locations within Cudgen Lake Nature Reserve...... 43 Figure 3.8. pH and redox values (from peepers) for the December 2002 sample period (a.). Soil and porewater pH values for the July 2004 sample period (b.)...... 44 Figure 3.9. AVS (a.), CRS (b.) and DOP (c.) from July 2004 soil profiles...... 45 Figure 3.10. pH values in addition to Fe and Mn (a.) and S (b.) concentrations from July 2004...... 47 Figure 3.11. Ni, Sn & Zn (a.), and Cr & U (b.) concentrations from July 2005...... 49 Figure 4.1. Schematic representation of a passive diffusion sampler...... 55 Figure 4.2. Location of the passive diffusion samplers during the December 2002 sampling period...... 59 Figure 4.3. Location of the passive diffusion samplers during the May 2003 sampling period...... 61
Figure 4.4. SO2 concentrations across different landuses, as measured by passive diffusion samplers, during December 2002...... 63
Figure 4.5. SO2 concentrations across different landuses, as measured by passive diffusion samplers, during May 2003...... 64 Figure 4.6. Average soil pH and redox values for the different landuse site during the May 2003 sampling period...... 65
Figure 4.7. SO2 concentrations at two different heights above the fallow soil surface (0.5 & 1.5 m) over a two-week period in May 2003...... 66 Figure 4.8. Atmospheric pressure, relative humidity, and rainfall for the time period before and after
sampling, along with the SO2 flux for the May 2003 sampling period...... 68 Figure 5.1. Diagrammatical simplification of micrometeorological theory...... 78 Figure 5.2. Location of the active gas analysis sites within the Blacks Drain area...... 81 Figure 5.3a. Setup of the active gas analysers and micrometeorological equipment during the 2003 sampling period...... 84 Figure 5.3b. Setup of the active gas analysers and micrometeorological equipment during the 2005 sampling period...... 85 Figure 5.4. Soil pH and boundary layers at the Nov/Dec 2003 gas analysis location within the Blacks Drain study site...... 87
List of Figures xix
Figure 5.5a. SO2 concentrations across the entire 2003 sampling period at 0.5 m above the ground surface...... 88
Figure 5.5b. H2S concentrations across the entire 2003 sampling period at 0.5 m above the ground surface...... 89
Figure 5.6. SO2 and H2S concentrations across a 32 hr period during the 2003 sampling period...... 90
Figure 5.7. Correlation between concentrations of SO2 & H2S during the 32 hr period in 2003...... 91
Figure 5.8a. SO2 flux across the entire 2003 sampling period...... 92
Figure 5.8b. H2S flux across the entire 2003 sampling period...... 93
Figure 5.9. Mean daily flux values for both SO2 and H2S across the entire Nov/Dec 2003 sampling period...... 94
Figure 5.10. SO2 and H2S fluxes across a 32 hour period during the Nov/Dec 2003 sampling period...... 95
Figure 5.11. PCA of SO2 (left) and H2S (right) concentrations and measured climatic variables during the 2003 sampling period...... 98
Figure 5.12. PCA of SO2 (left) and H2S (right) fluxes and measured climatic variables during the 2003 sampling period...... 99
Figure 5.13. PCA of SO2 (left) and H2S (right) fluxes and measured climatic variables, based on wind direction (80-160°) during the 2003 sampling period...... 100 Figure 5.14. Redox, pH and soil boundary layers for a profile taken at the study site during the 2005 gas sampling period...... 101
Figure 5.15. SO2 and H2S concentrations across the entire 2005 sampling period...... 103
Figure 5.16. SO2 and H2S fluxes across the entire 2005 sampling period...... 104
Figure 5.17. Averaged daily fluxes of SO2 and H2S throughout the 2005 sampling period...... 105
Figure 5.18. Comparison between SO2 flux and atmospheric pressure during the 2005 sampling period...... 107
Figure 5.19. Comparison between H2S flux and atmospheric pressure during the 2005 sampling period...... 108
Figure 5.20. Comparison between SO2 / H2S flux and rainfall during the 2005 sampling period...... 110
Figure 5.21. Comparison between SO2 flux and air temperature during the 2005 sampling period...... 111
Figure 5.22. Comparison between H2S flux and air temperature during the 2005 sampling period...... 112
Figure 5.23. PCA of SO2 (left) and H2S (right) concentrations and measured climatic variables during the 2005 sampling period...... 114
Figure 5.24. AHC of SO2 (top) and H2S (bottom) concentrations and measured climatic variables during the 2005 sampling period...... 115
Figure 5.25. AHC of SO2 (top) and H2S (bottom) fluxes and measured climatic variables during the 2005 sampling period...... 116
Figure 5.26. AHC of SO2 (top) and H2S (bottom) fluxes, flowing over the ideal fetch, and measured climatic variables during the 2005 sampling period...... 117
Figure 5.27. Formation and possible export of H2S from the decomposition of organic matter...... 120 Figure 5.28. Extractable DNA as a measure of soil biomass at the Blacks Drain and Cudgen study sites...... 126
List of Figures xx
Figure 5.29. Reaction pathways involving the photo-oxidation of H2S by the free radicals OH, NO2 and
O3, ultimately forming SO2...... 130 Figure 6.1. Schematic representation of the Shimadzu FPD 17A...... 142 Figure 6.2. Schematic representation of the SPME device, and a magnified representation of carboxen- PDMS fibre...... 146 Figure 6.3. Schematic representation of the fundamental GC setup employed for this study...... 148 Figure 6.4. Schematic and photograph demonstrating the use of SPME with the septa jars...... 152 Figure 6.5. Calibration curves for DMS (top) and ESH (bottom), and correlation coefficients...... 153 Figure 6.6. DMS and ESH concentrations down a soil profile taken at Blacks Drain during May 2006...... 154 Figure 6.7. Sediment sample chromatograph from within Blacks Drain taken during May 2006, and the calculated DMS concentration...... 155 Figure 6.8. DMS concentrations down a soil profile taken at Cudgen Lake Nature Reserve during May 2006...... 156 Figure 6.9. Sample chromatographs from the Cudgen Lake site surface sample...... 156 Figure 6.10. Chromatograph of a surface sample taken from the Cudgen Lake site with the addition of
NaN3...... 157 Figure 6.11. Sediment sample chromatograph from Cudgen Lake edge taken during May 2006, and calculated DMS concentration...... 158 Figure 7.1. Schematic summary of the processes important in the formation and release of VSCs from agricultural ASS...... 184 Figure 7.2. Schematic summary of the processes important in the formation and release of VSCs from undisturbed ASS...... 185
List of Figures xxi
LIST OF ABBREVIATIONS
3-MPA 3-Mercaptopropionate
ABL Atmospheric Boundary Layer Agglomerative Hierarchical IC Ion Chromatography AHC Clustering International Pollution Control Australian & New Zealand IPCC Commission ANZECC Environment & Conservation Council AVS Acid Volatile Sulfur LOD Limit of Detection
CCN Cloud Condensation Nuclei COS Carbonyl Sulfide MMPA 3-Methiolpropionate CRS Chromium Reducible Sulfur MSH Methanethiol
DMS Dimethylsulfide PDMS Polydimethylsiloxane DMDS Dimethyldisulfide PCA Principal Components Analysis DMSP Dimethylsulfoniopropionate DMSO Dimethylsulfoxide Degree of Pyritisation DOP Simultaneously Extracted DOS Degree of Sulfidisation SEM Metals SO4 Sulfate 2- SO3 Sulfite Sulfur Dioxide ESH Ethanethiol SO2 SPME Solid-Phase Microextraction SRB Sulfate-Reducing Bacteria FPD Flame Photometric Detector
WHO World Health Organisation GC Gas Chromatography
VSCs Volatile Sulfur Compounds VOSCs Volatile Organic Sulfur HCl Hydrochloric Acid Compounds High Performance Liquid HPLC Chromatography H2S Hydrogen Sulfide
List of Abbreviations xxiii
Chapter One: INTRODUCTION
1.1 PREFACE
This project was initiated as part of a collaborative research effort involving several research institutions, looking at the different element cycles (carbon, nitrogen, iron and sulfur) within acid sulfate soils (ASS). However, the focus of this thesis relates to aspects of the sulfur cycle in ASS.
Funding was provided by the Australian Research Council for ‘Interactions between sulfur, nitrogen and iron cycles in the sustainable management and use of acid sulfate soils’ (No LP0219475), as well as by our industry partners; the New South Wales Sugar Milling Co-operative, New South Wales Canefarmers Association, and Tweed Shire Council.
1.2 CONCEPTUAL OVERVIEW
Coastal ASS, although being part of the natural, global, biogeochemical sulfur cycle, have been described by some as “… the nastiest soils in the world” (Dent and Pons 1995). This is primarily because of the extreme levels of acidity generated upon their oxidation and associated environmental, agronomic and engineering consequences.
Many different aspects of ASS have been researched within Australia and internationally, including identifying the mechanisms of their formation (Dent 1986; Pons 1973; Rickard 1997; Rickard and Luther 1997; Rickard 1975; van Breemen 1973; 1976; 1982; 1988b) and oxidation (Dent 1986; Evangelou 1995; Nordstrom 1982); as well as the inter-conversion of reduced sulfides (Burton et al. 2006c; Luther III et al. 1986b; Luther III et al. 1982; Morse and Rickard 2004); quantifying acidity and metal
Chapter One: Introduction 1
discharge (Astrom 2001a; b; Astrom and Astrom 1997; Astrom and Bjorklund 1995; Lin et al. 1998a; Sundstrom et al. 2002); examining processes of oxidation-product discharge (Astrom and Spiro 2000; Kinsela and Melville 2004; Wilson et al. 1999); as well as looking at the resulting ecological (Macdonald et al. 2004b; Sammut et al. 1993; Sammut et al. 1995) and human impacts (Dent 1986). This list of publications is neither exhaustive in the issues it covers, nor does it represent the quantity of research within each area. Whilst the above areas address most of the fundamental aspects related to ASS chemistry, the issue of volatile compound formation and cycling is an area that has received little to no attention.
Preliminary research into ASS led van Breemen to suggest that a large fraction of the potential acidity present in pyritic sediments could leave the profile in an unneutralised gaseous form, as SO2 (van Breemen 1976). Despite this, there is little information on volatile sulfur compounds (including emissions estimates), within acid sulfate landscapes in Australia. Indeed, at the commencement of this thesis research, no publications had been produced regarding gaseous emissions from ASS in Australia.
Measurements of sulfur dioxide (SO2) from a coastal ASS were made during 1999 and 2000 (Macdonald et al. 2004a), which formed the basis of further research into this topic.
More generally, scientific research into sulfur gases has multiplied rapidly over the last three decades, primarily because of their close association with air pollution, precipitation chemistry and linkage to broad-scale climatic changes. Whilst a large proportion of the early research into sulfur gases focused on anthropogenic inputs, there has been a realisation that the identification of biogenic sources of these gases are of equal importance in the understanding of the global biogeochemical sulfur cycle. Indeed, understanding the sources of biogenic gases is of greater importance in countries such as Australia, where, owing to the relatively small population and urban isolation, anthropogenic inputs are substantially less than observed in North America and Europe (Ayers et al. 1995). Much of this initial work has focussed on detailing biogenic sulfur emissions from marine / oceanic sources owing to its dominance of the overall atmospheric sulfur budget. The identification that highly productive coastal systems, such as tidal marshes, wetlands and estuaries were also significant S gas emitters, has also meant a number of studies have been published in these areas.
Chapter One: Introduction 2
1.3 THESIS OBJECTIVES
As only very preliminary work in the area of gas emissions from ASS has been undertaken previously, the primary objective of this thesis is to challenge the long-held belief that agricultural soils are sinks for sulfur gases, and subsequently to investigate possible volatile sulfur compounds (VSCs) emitted from coastal ASS under sugarcane cropping. The compounds of major interest include the aforementioned SO2, and hydrogen sulfide (H2S), a compound integral in the cycling of sulfur within ASS. Other low-molecular weight VSCs are investigated in lesser detail using emergent techniques.
Specific Aims a.) To characterise and evaluate the ASS study sites on the New South Wales north coast for future gas emission sampling; b.) To verify the original measurements of Macdonald et al. (2004a), as well as
to assess differences in SO2 emissions across different landuses using passive diffusion samplers;
c.) To ascertain real-time concentrations and fluxes of SO2 and H2S using advanced micrometeorological techniques, and to establish diurnal fluctuations and relate measurements to meteorological parameters; d.) To establish a method for the measurement of low-molecular weight volatile sulfur compounds from ASS using gas chromatography.
1.4 THESIS OUTLINE
This thesis has a somewhat unconventional format as the experimental chapters (3 through to 6) are largely independent, containing individual background, methods, results and discussion sections. Before that, the initial background chapter (Chapter 2) outlines the existing knowledge regarding the formation and emissions of volatile sulfur compounds, as well as giving a general geochemical overview of ASS. The four experimental chapters that follow correspond to the specific aims set out above. Chapter 3 contains a characterisation of the two study sites used in this thesis; Blacks Drain (an
Chapter One: Introduction 3
agricultural ASS utilised for intensive sugarcane cultivation) and Cudgen Lake Nature Reserve (a relatively undisturbed ASS). The next chapter (Chapter 4) then focuses on the emissions of SO2 using passive diffusion samplers, as well as evaluating the technique for future gas sampling programmes. This is followed by a detailed analysis of SO2 and H2S concentrations and fluxes determined by micrometeorological methods (Chapter 5). This chapter also examines the potential relationships between the two gases and the climatic factors influencing emissions. Chapter 6 looks at the use of gas chromatography for the measurement of low molecular weight VSCs, comparing novel measurements from Blacks Drain and Cudgen Lake Nature Reserve. The thesis is concluded with a cumulative summary of all the results, followed by a discussion on the implications of the measurements and looking at possible future research directions based on the current results.
Chapter One: Introduction 4
Chapter Two: ACID SULFATE SOILS & SULFUR GASES BACKGROUND
2.1 ACID SULFATE SOIL BIOGEOCHEMICAL THEORY
2.1.1 Introduction
The accumulation of iron sulfides is a naturally occurring process forming part of the global sulfur biogeochemical cycle over geological time (Pons 1973). Although ASS have been known to exist since the 17th century (Dent and Pons 1995), research only reached international recognition, including Australia, in the mid-1970’s; see (Dost 1973; Walker 1972). ASS are now recognised as having adverse impacts on a variety of different landscapes across the globe, including areas of; Europe - Finland, Sweden, the Netherlands; North and South America - Brazil, U.S.A.; West Africa - Senegal, Guinea Bissau, Gambia, Sierra Leone; Asia - China, the Philippines, Vietnam, Indonesia, Malaysia, Sri Lanka, Thailand; and in Australia (Chen et al. 2006; Dent and van Mensvoort 1993; Dost 1973; 1988; Dost and van Breemen 1982; Konsten et al. 1994; Lin and Melville 1994; Lin et al. 1995b; Palko and Yli-Halla 1993). A comprehensive global history of ASS can be found in Dent and Pons (1995).
ASS have been loosely defined as any sediment that contains quantities of iron sulfides, that when oxidised to sulfuric acid, exceed the acid neutralising capacity of the carbonates in the soil (Melville et al. 1993; Pons et al. 1982). As such, they are characterised by a highly acidic (pH < 3.5) oxidised horizon, overlying a neutral to alkaline (pH > 6.5) reduced horizon which is rich in iron sulfides. In an undisturbed state these soils are relatively harmless, but when disturbed and subsequently oxidised via watertable fluctuations, either naturally from climatic / sea level changes and the isostatic readjustment of the lithosphere, or from artificial drainage for agriculture, large
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 5
quantities of sulfuric acid and other sulfur compounds are released. Because of their extremely variable nature, their classification has proven troublesome. According to the Australian Soil Classification (Isbell 1996), they are regarded generally as Hydrosols; or as Sulfaquepts and Sulfaquents within Soil Taxonomy classification (Soil Survey Staff 2006). As a vast quantity of literature has been devoted to the examination of iron sulfide formation and oxidation, what follows is only a very brief summary of the key points on ASS biogeochemistry.
2.1.2 Iron Sulfide Formation
The formation of iron sulfides is an integral part of the global biogeochemical sulfur cycle, generally requiring sufficient ferric iron from sediments, sulfate from seawater; organic matter, and a predominantly anaerobic environment (Berner 1970; Melville et al. 1993; Pons et al. 1982; van Breemen 1973). As a result, millions of hectares of coastal floodplains and wetlands across the globe are underlain by sulfidic sediments (Dent 1986). Conditions for the accumulation of pyrite primarily occur in wave- protected, low-energy environments, such as tidal marshes and backswamps (Melville et al. 1991). These conditions were typical of the eastern Australian coastline throughout the majority of the Holocene (Thom and Chappell 1975; Thom and Roy 1985), with the large tidal ranges, small catchments and low outflows providing for the development of substantial quantities of pyritic sediments (White et al. 1997). This is confirmed in the case of the Tweed River catchment where Holocene-age pyrite concentrations are approximately 3 %, and as such are the sediments of most concern (van Oploo 2000; White et al. 1997).
The formation of pyrite is a complex process involving several bacterial catalysts in addition to purely chemical processes. The overall reaction stoichiometry of pyrite formation takes place according to Equation 2.1;
2- - Fe2O3(s) + 4SO4 + 8CH2O + ½O2 bbb 2FeS2(s) + 8HCO3 + 4H2O [2.1]
To summarise, the process of pyrite formation involves the reduction of sulfate to sulfide, which is microbially catalysed by heterotrophic sulfate-reducing bacteria (SRB)
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 6
(primarily Desulfovibrio spp.). This is followed by its partial oxidation (potentially by Fe3+) to elemental sulfur or polysulfides. These are then able to react with various iron containing minerals to form iron sulfide (Berner 1970; van Breemen 1988b), Equation 2.2.
2FeOOH + 3H2S bbb FeS + FeS2 + 4H2O [2.2]
This process is typical only under acidic conditions (pH < 4). At higher pH (characteristic of most marine depositional environments), the reaction produces pyrite at a much slower rate forming primarily amorphous FeS (Berner 1967).
The final phase of pyrite formation is the transformation of iron monosulfides to iron disulfide (pyrite) (Berner 1970). At neutral pH, pyrite can form through the reaction of iron monosulfide with elemental sulfur (Berner 1970), as shown in Equation 2.3, which at low temperatures is actually S8 (Rickard 1975);
0 FeS + S bbb FeS2 [2.3]
This equation represents a generalised reaction for the formation of pyrite, showing the rate controlling molecular step. More specifically, the formation of pyrite from iron monosulfide can occur via two primary pathways. Firstly, the polysulfide pathway, which can be represented in many ways but was originally recognised by (Berner 1970), and is shown in Equation 2.4. This equation is a little misleading though, as the process has been shown to occur from the loss of ferrous iron rather than from the addition of zero-valent sulfur (Wilkin and Barnes 1996).
2- 2- FeS + Sx bbb FeS2 + S(x-1) [2.4]
And secondly, the formation of pyrite can occur through the direct interaction between iron monosulfide and hydrogen sulfide (Equation 2.5), as suggested by Wachtershauser (1988), and demonstrated by Rickard (1997), and Rickard and Luther (1997);
FeS + H2S bbb FeS2 + H2 [2.5]
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 7
This more recently demonstrated reaction has been shown to progress more rapidly than the polysulfide pathway, with the H2S acting as the necessary oxidising agent. Within natural systems, the net pathway by which iron monosulfides are formed, strongly depends on the pH and total dissolved sulfide concentrations (Rickard 1995).
Under certain conditions, iron monosulfides can persist in the environment, accumulating primarily in estuarine drainage channels and deeper man-made drains in ASS backswamps (Bush et al. 2004). Iron monosulfides have been shown to be particularly important in these environments, as they are the major solid-phase, controlling concentrations of the trace metals; As, Cd, Cu, Ni, Pb & Zn, as well as Fe and Mn, through processes of adsorption and co-precipitation (Billon et al. 2001b; Huerta-Diaz et al. 1998; Morse and Arakaki 1993; Simpson et al. 1998), and concentrations can accumulate to over 18% FeS (Bush et al. 2004).
Whilst this is merely a brief summary of the key components of iron sulfide formation, further details can be obtained in the references listed above, as well as from the following reviews (Berner 1984; Goldhaber and Kaplan 1974).
2.1.3 Iron Sulfide Oxidation
Iron sulfides are stable under reducing conditions such as when completely dry or when permanently below the watertable. However, exposure to air can result in their oxidation. The exposure of sediments to air can result by natural means such as landscape uplift, lowering of sea levels, and general climatic fluctuations; or alternatively through human-induced mechanisms such as excavation or drainage; see Dent (1986), Kinsela and Melville (2004), Lin et al. (1995b), and Section 2.1.5 for a detailed discussion.
The overall stoichiometry of the oxidation of pyrite with oxygen can be written simply as;
15 5 FeS2 + /4O2 + /2H2O bbb FeOOH + 2H2SO4 [2.6]
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 8
In actual fact though, the oxidation of pyrite, as is the case with its formation, is a complex process involving many redox sensitive chemical and microbial reactions (Nordstrom 1982). When such sediments are exposed to oxygen, the initial products formed are Fe(II), SO4 and acidity, as shown in Equation 2.7. This process occurs very slowly.
7 2+ 2- + FeS2 + /2O2 + H2O bbb Fe + 2SO4 + 2H [2.7]
Ferrous iron can be transported upwards in the soil profile by diffusion or capillary rise (Lin et al. 1998a; Patrick and DeLaune 1972), after which it can be exported from the profile through drainage and leaching processes (van Breemen 1988a). These movements of Fe(II) ultimately result in its further oxidation, as shown in Equation 2.8. This reaction is bacterially catalysed by the ubiquitous chemoautotrophs Acidithiobacillus ferroxidans and Acidithiobacillus thiooxidans at pH < 4, increasing reaction rates by 5 to 6 orders of magnitude (Singer and Stumm 1970).
2+ + 3+ Fe + ¼O2 + H bbb Fe + ½H2O [2.8]
Whilst the pH remains neutral to alkaline, the Fe(III) precipitates as iron hydroxide. Once the pH drops below 3-4, the ferric iron remains soluble and it can subsequently anaerobically oxidise pyrite producing a substantially greater quantity of acidity and at a much faster rate than O2-driven oxidation (Equation 2.9);
3+ 2+ 2- + FeS2 + 14Fe + 8H2O bbb 15Fe + 2SO4 + 16H [2.9]
The resulting products of this reaction, Fe(II) and H+, further drive the oxidation process by creating more ferric iron, as per Equation 2.8. To further generalise, the oxidation of pyrite is predominantly initiated by oxygen, and after the subsequent decrease in pH, ferric iron acts as the dominant oxidant. As Fe(II) can be reduced from pyrite (Equation
2.9) faster than Fe(III) can be regenerated by its oxidation by O2, Equation 2.8 is commonly referred to as the rate determining step in pyrite oxidation (Singer and Stumm 1970).
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 9
It should be noted that the initial oxidation of pyrite may be brought about by the oxidation of iron monosulfides, which are more reactive than pyrite (Bloomfield 1972; Bush and Sullivan 1997; van Breemen 1988b). Although a complicated process (see Rickard and Morse (2005) for an in-depth review), the overall reaction proceeds according to the following equation (Saulnier and Mucci 2000);
9 2- + 2FeS + /2O2 + 5H2O bbb 2Fe(OH)3 + 2SO4 + 4H [2.10]
Again, this is only a brief summary of the key components of iron sulfide oxidation, with further details available from the references listed above, as well as from the following reviews (Evangelou 1995; Nordstrom 1982).
2.1.4 Oxidation Impacts
The acidity generated is of particular environmental concern, whether through the oxidation of pyrite itself or from the subsequent hydrolysis of the acidic metal ions (Al3+ and Fe2+) derived from aluminosilicate dissolution; e.g. Equation 2.11 from Sposito (1989).
3+ + Al + 3H2O bbb Al(OH)3(s) + 3H [2.11]
For example, Sammut et al. (1996) estimated acid discharges of the order of 950 t of sulfuric acid (H2SO4) during a single flooding event on the Richmond River, NSW, Australia. Accumulations of the minerals resulting from pyrite oxidation include a wide range of iron oxides (goethite, schwertmannite, haematite) and sulfates (jarosite), which are primarily the result of the changes in pH and redox activity that is induced (Fanning et al. 2002). These metastable minerals form both within the soil profile, and at a distance, coating banks of drainage lines, creating a temporary buffering of the waters to very acidic conditions. Their dissolution associated with changes to redox and pH re- releases the acidity and metal ions back into solution, as per Equation 2.12 below for jarosite (van Breemen 1982). The importance of schwertmannite has recently been identified in controlling the activity of Fe3+ in ASS (Sullivan and Bush 2004), primarily because of its preferential formation at low pH (Bigham et al. 1996). The leaching of
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 10
metals through cation exchange and aluminosilicate dissolution is an associated problem of acidic discharge, and it has been extensively researched in acid sulfate environments both in Australia (Lin et al. 1998a) and globally (Astrom 2001b; c; Astrom and Bjorklund 1995).
2- + + KFe3(SO4)2(OH)6 bbb 3FeOOH + 2SO4 + 3H + K [2.12]
The impacts of these oxidation products are also far-reaching, with the acidity and dissolved metal concentrations (particularly aluminium) being responsible for large fish kills, general ecosystem degradation (Callinan et al. 1993; Callinan et al. 2005; Easton 1989; Sammut et al. 1993; Sammut et al. 1995; Willett et al. 1993), including the promotion of algal blooms (e.g. Lyngbya majuscula) from enhanced iron concentrations (Watkinson et al. 2005), as well as deleterious impacts on the anthropogenic environment including impacts on aquaculture (Dent 1986; Gosavi et al. 2004).
In addition to discrete S-phases, many metals including As, Cd, Cu, Mo, Ni, and Zn can be adsorbed to, as well as be co-precipitated with Mn or Fe sulfides (Billon et al. 2001b; Cooper and Morse 1996; 1999; Helz et al. 2004; Huerta-Diaz and Morse 1992; Huerta- Diaz et al. 1998; Moore et al. 1988; Morse and Luther III 1999; Sohlenius and Oborn 2004; Vorlicek et al. 2004)
The oxidation of iron monosulfides from their disturbance has been also shown to have wide ranging impacts, especially as their oxidation proceeds very rapidly releasing large quantities of acidity, associated metals, as well as deoxygenating the water column 0 (Bush et al. 2004). It is thought that the oxidation of S8 , rather than FeS itself, produces the majority of acidity during oxidation events (Burton et al. 2006a), as shown by its formation from FeS (Fe2+ & S2-) Equations 2.13 & 2.14, and oxidation Equation 2.15 below, from (Burton et al. 2006a);
2+ + Fe + ¼O2 + 1½H2O bbb 2H + FeOOH(s) [2.13] 2- + 0 S + ½O2 + 2H bbb H2O + YS8 (s) [2.14] 0 2- + YS8 + 1½O2 + H2O bbb SO4 + 2H [2.15]
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 11
The reduction of oxidation products is also an issue of concern in ASS. This is because the reduction is often only a temporary store of potentially detrimental compounds leached from ASS, frequently at a distance from their point of origin (Morse 1994b).
Further information on the biogeochemical cycling within ASS can be sourced in the following thesis literature reviews (Keene 2001; Smith 2005; van Oploo 2000; Wilson 1995).
2.1.5 Natural vs. Anthropogenic Oxidation of Acid Sulfate Soil Landscapes
It needs to be recognised that the processes threatening the exposure of the pyritic sediments, and therefore environmental impacts, differ widely on a global through to even a local scale. For example, whilst the isostatic adjustment of the land surface in conjunction with intensive drainage is of vital importance along the Finnish coastline (Astrom and Bjorklund 1997; Astrom and Spiro 2000; Osterholm and Astrom 2002; Palko and Yli-Halla 1993), the issue of deep drainage for aquaculture is a more relevant means of exposure within the developing coastal location in the southeast Asian region, see Minh et al. (2002) for example. This particular section looks at the often contentious contributions of natural versus anthropogenic oxidation of ASS within an Australian context, particularly at the sampled study sites.
Most of the Tweed River floodplain is used for broad-acre sugarcane production with well established man-made drainage systems. It is commonly believed that drainage systems with 1-way flap gates have caused the observed oxidation of pyritic sediments and the associated lowering of the watertable relative to the sulfidic sediments (Walker 1972). However, Lin et al. (1995a) demonstrated that oxidation of ASS landscapes occurred even without drainage systems.
The Australian ASS literature appears to generally propose that European drainage of backswamp ASS landscapes has been the predominant cause of the observed oxidation and acidification problems, perhaps greatly influenced by the earlier literature based on drainage-induced acidification in Dutch polders that were previously shallow coastal
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 12
seas, e.g. Pons (1973). This is also unambiguously the case in Australia where bund walls and flap-gates were constructed in permanently tidal-saturated landscapes, such as in the mangrove swamps on the southern shore of Trutes Bay (a left-bank tidal embayment of the lower Tweed River), or on the right-bank shore of East Trinity Inlet, Queensland; see Cook et al. (2000). However, in estuarine floodplain backswamps, the belief of a European drainage-caused landscape oxidation appears to be founded on the general observation that when artificial drainage systems were first installed, increased acidic discharge and their environmental impacts occurred. However, such observed acidic discharge would also occur if the drainage system only provided the new conduit for enhanced acidity export from an already naturally oxidised landscape. We have now proposed the importance of this latter, natural pedogenesis, acidity cause, based on many field observations (Donner 2001; Kinsela and Melville 2004; Lin et al. 1995a; Smith et al. 2003). Unfortunately, it is now difficult to test natural pedogenesis satisfactorily because most of these floodplain backswamps on east coast Australia have already had some degree of drainage, and no sufficiently detailed pre-drainage analysis of ASS and the landscape hydrology has been undertaken.
It is logical that such drainage systems markedly reduced the duration of backswamp inundation (White et al. 1993), but it has not been established that the changed degree and duration of backswamp inundation altered watertable elevations and ASS profile saturation sufficiently to cause the observed, almost ubiquitous, >1 m depth of ASS profile oxidation. The lowering of the watertable should be seen as preferential oxidation down drained, generally vertical macropores. To some degree this does occur; however, most commonly at McLeod’s Creek and at other sites where such drainage systems exist, a uniformly horizontal oxidation front occurs within the saturated sediment and below any observable watertable elevation (van Oploo 2000; Wilson et al. 1999). This oxidation depth of generally > 1 m exists in the fine-grained sulfuric/sulfidic sediment with a very small pore size that gives an effective capillary fringe and saturated sediment near to the ASS mineral profile’s surface. To drain the pores in this sediment would require a watertable elevation much lower (>> 1 m) than any observed in the past 15 years of study at McLeod’s Creek. This period recently coincided with the most severe and prolonged drought recorded in European settlement of the Tweed area. Such an apparently uniform and deep oxidation front is also observed where no artificial drainage has occurred; see Lin et al. (1995a). The existence
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 13
of this oxidation front below the watertable appears most likely due to downward - diffusion through the saturated soil of some soluble oxidants, such as NO3 . This is currently under research at McLeod’s Creek so that a clearer picture of ASS landscape pedogenesis may be obtained.
Some authors (Blunden and Indraratna 2000; Johnston et al. 2004) have presented results where at least some of their study site materials are sandier and/or apparently have a greater degree of lateral hydraulic conductivity than that described for McLeod’s Creek. Nevertheless, these represent exceptions to the nature of the materials deposited in the geomorphic mud basins that generally formed behind east coast sand barriers.
The geomorphological variations in sediment characteristics and stratigraphy are not well recognised in ASS research, so that the oxidised sediment at the top of the ASS profile is presumed to have been the same as that below the oxidation front, particularly with respect to the original sulfide mineral content. This is unlikely to be the case because, whereas the deeper estuarine sediment was deposited in brackish tidal water conditions (rich in dissolved sulfate), the material closer to the surface will have had an increasing fluvial influence. Walker (1970) provided an excellent early view on geomorphic/pedogenic development in the estuarine/fluvial evolution of the lower Macleay River floodplain. Dalrymple et al. (1992) and Roy et al. (2001) also provide models to explain the infilling of east coast estuarine embayments with their sand- barrier-induced, inner mud basins where sulfidic sediments accumulated initially, later to be overtopped by fluviatile sediment. The degree of estuary embayment infilling and emergence of the mud basin surface above a shallow tidal brackish lake will vary, depending on sediment input rate and the relative sea level in the estuary embayment. There are some estuary embayments that are ‘mature’ and completely infilled (e.g. most of the Clarence embayment) and other ‘immature’ embayments where significant parts have not been infilled (e.g. Cudgen Lake). Hashimoto’s (2005) thesis provides a recent detailed picture of the contrasting Tweed, Richmond and Clarence Rivers’ floodplain evolution.
The history of the last global post-glacial sea level rise from about -125 m at 20 ka to its present position at about 6.5 ka is now well established; see Lambeck and Chappell (2001). Thom and Roy (1985) proposed their models of east coast Australian
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 14
geomorphic evolution based on a constant sea level since about 6.5 ka. This is likely to be true of the sea level as controlled by the balanced inputs/outputs of a constant water volume into a globally constant ocean basin volume. However, recent research on east coast Australia shows that sea level reached a maximum at about 6.5 ka but this was up to 1 - 3 m above today’s relative sea level. Isostatic readjustment of the lithosphere along the Australian continental margin caused coastal uplift, but the degree of this uplift varied even at locations only 100 km apart, and increased with distance inland from the coast (Lambeck and Nakada 1990). Some work suggests this uplift rate (and effective sea level fall) has been constant until the present; see Lambeck and Chappell (2001), Lambeck and Nakada (1990), Nakada and Lambeck (1989). However, Baker and Haworth (1997), Baker et al. (2001), Flood and Frankel (1989) used fossil intertidal tube worm deposits stranded on coastal cliffs to infer a more prolonged, constant high relative sea level until about 2 ka, then a fall to present relative sea level. Whatever the timing of a relative sea level fall by up to 1 - 3 m, the magnitude and duration of this drainage base level decline is sufficient to account for much of the observed backswamp ASS oxidation, certainly where contact with tidal influence has been maintained.
The weather patterns of today seem viewed by some as representative of the climate throughout the Holocene epoch (< 10 ka). That this is unlikely to be true can simply be seen from variations in rainfall and associated flooding during the second half of European settlement. Pittock (1975) identified a 10 - 20 % increase in mean annual rainfall over much of NSW and Queensland during the 1940s to 1970, compared with the first half of the 20th Century. W. D. Erskine and his colleagues extended this study to include records from the 19th Century and later 20th Century and proposed multi- decadal periods (about 50 years) of drought-dominated rainfall regime (about 1890– 1946) and flood-dominated rainfall regime (< 1890 and 1946+), mostly with increased summer rainfall. Bell and Erskine (1981), Erskine (1986), Erskine and Warner (1988) showed this latter increase in rainfall in the Nepean and Hunter Valleys gave an upward shift by 50 - 100 % in the flood frequency curve. Smith and Greenway (1983) also showed an increase in flood height on the Tweed, Richmond, and Clarence Rivers after the mid 1940s. We do not have evidence of the existence of the weather patterns before European settlement, but the global ‘Medieval Warm Period’ (about 1100 - 1300 AD) had mean temperatures several degrees celcius warmer than today. Elevated
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 15
temperatures will be expressed in increased evapotranspiration, which has been shown as a major control on ASS floodplain watertable elevation, e.g. White et al. (1997).
The south-eastern Australian climate of the Holocene, since present sea level was attained (about 6.5 ka), has been deduced by many researchers from lake levels and salinities, from vegetation signatures, and from aeolian dust deposit records. In general, from about 8 to 5 ka, the climate was wetter than present (Anker et al. 2001; Bowler 1981; Chivas et al. 1993; Chivas et al. 1985; Dodson 1986; Magee et al. 1995). From about 5 to 2 ka, conditions were drier than present with significant changes in Tasmanian pollen spectra (Anker et al. 2001) and changes in inland lakes and sediments, for example, Stanley and De Deckker (2002). Bowler (1981) and Chivas et al. (1985) showed that from about 2000 BP to 300 or 400 BP, Lake Keilambete in Victoria returned to perennial lake conditions. However, this return of high lake levels was not shown by Dodson (1986) for Breadalbane or by Coventry and Walker (1977) at nearby Lake George, NSW. It is unclear exactly what natural vegetation changes have occurred on coastal floodplains over the past 6.5 ka in response to changes that occurred in climate and associated evapotranspiration-driven drainage; it is likely that they have been profound. The shift to drier conditions over much of the last few thousand years compared with conditions initially experienced by landscapes accumulating sulfidic sediments favours natural landscape drainage. This is particularly the case when base level is also lowered.
It seems to us (Kinsela and Melville 2004) that natural processes of landscape / hydrology evolution and pedogenesis can account for the almost ubiquitous existence and degree of much of east-coast Australia’s backswamp ASS acidification. Artificial drainage systems may not have caused the acidity formation but they do provide the conduit for its enhanced export. Therefore, if any such landscapes have not been drained, any proposal to initiate their drainage should be avoided. For those backswamps already drained, the management of the acidity export is essential and must focus on mechanisms to maximise acidity retention in the landscape, and treatment of any acidity being exported in drain systems.
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 16
2.2 ACID SULFATE SOILS AND SULFUR GASES
Soils have long been thought of as an important natural sink for many gases including ammonia, nitrogen dioxide as well as sulfur gases (Babich and Stotzky 1978; Bremner and Steele 1978). Bremner and Banwart (1976) have shown that both moist and air- dried soil sorb SO2 and H2S. It is especially the case within moist or waterlogged soils
(Freney and Williams 1983), such as ASS, with regards to SO2 owing to its high solubility in water (9.4 g / 100 g water) compared to other gases (e.g. 0.112 g for NO2 and 0.33 g for H2S) (figures from Aylward and Fidlay (1994)). Organic sulfur gases are also readily sorbed by soils, but most soils have a high capacity to adsorb significantly more SO2 and H2S compared to DMS (Bremner and Banwart 1976; Smith et al. 1973).
In an attempt to balance acid-base budgets for ASS profiles, van Breemen suggested that a large fraction of the potential acidity present in pyritic sediments could potentially leave the sediment in an un-neutralised form as SO2 (van Breemen 1976; 1982; 1993).
The emission of SO2 from ASS was first quantified by Macdonald et al. (2004a) using a combination of static chambers and micrometeorological techniques. The results of this study indicated that emissions of SO2 were a previously unaccounted-for source of global sulfur (S), estimated to be of the order of 3 Tg/year, equivalent to 3% of known anthropogenic emissions of S (Macdonald et al. 2004a).
2.2.1 Previous Research into Sulfur Gas Emissions from Acid Sulfate Soils
Sulfur is an important redox element under natural conditions, being responsible for numerous biogeochemical processes such as sulfate reduction, pyrite formation, metal transport and capture, as well as atmospheric sulfur emissions (Luther III et al. 1986b). Very little research into sulfur gas emissions from sugarcane growing on ASS has been conducted. Research on this subject has primarily been focused around emissions from coastal marsh and wetlands (Cooper et al. 1989); marine / open oceans (Andreae 1990); upland soils (Staubes et al. 1989); rice paddies (Kanda and Minami 1992; Nouchi et al. 1997); maize and wheat (Kanda et al. 1995).
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 17
The most recent publication on this topic is that by Macdonald et al. (2004a). Based on work by Barnett and Davis (1983), Saltzman et al. (1983) and Barnett (1985),
Macdonald and colleagues, speculated that emissions of SO2 from ASS were possibly 2- due to the evaporation of soil porewaters containing sulfite (SO3 ) in a manner similar to that shown in Equations 2.16 - 2.18. Small concentrations of sulfite were measured in the surface soils from where SO2 emissions were measured.
+ 2- - H + SO3 ↔↔↔ HSO3 [2.16] + - H + HSO3 ↔↔↔ SO2.H2O(l) [2.17]
SO2.H2O(l) ↔↔↔ SO2(g) + H2O [2.18]
2.3 SULFUR GASES
Sulfur cycling in the troposphere plays an important role in atmospheric transformations, both through its formation of aerosol particles and acid-base chemistry. Emissions to the atmosphere can be broadly segregated into two categories, anthropogenic and natural (or biogenic). Although estimates of anthropogenic emissions are relatively accurate, because of large spatial and temporal variations there still exists a large discrepancy in estimations of biogenic emissions (Andreae and Jaeschke 1992). Even though there are large variations in emission estimates from biogenic sources, its overall contribution to the atmospheric sulfur cycle are within a similar range to that from anthropogenic sources.
The biogeochemical transformations of organic and inorganic sulfur have large impacts on global chemistry, impacting on the biosphere (microbial turnover, food-chain developments), lithosphere (mineralisation processes) and atmosphere (volatilisation, acid-base alterations, global climate changes). It is therefore difficult to separate the sulfur processes from the other closely related biogeochemical cycles; carbon (C), nitrogen (N) and iron (Fe). In addition to this, the varying oxidation states of sulfur (-2 through to +6) means that it can undergo a multitude of biogeochemical transformations within the natural environment (Luther III et al. 1986b). Some of the sulfur species that may occur or be emitted from ASS will now be considered.
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 18
2.3.1 Sulfur Dioxide
The existence of anthropogenic sources of SO2 has been widely acknowledged for a considerable period of time. Studies on the emission of SO2 have centred on the combustion of fossil fuels and the smelting of metals (Cullis and Hirschler 1980), primarily because of awareness surrounding impacts of the created air pollution on human health (Kunzli et al. 2000; Pope et al. 1995; Samet et al. 2000). SO2 has been documented as a significant pollutant in densely populated areas as well as around major coal-burning power stations, oil processing, and the smelting of ferrous and non- ferrous ores (Badr and Probert 1994), primarily through its conversion to sulfuric acid.
Indeed, the International Pollution Control Commission (IPCC) identified SO2 as one of the top five atmospheric pollutants (IPCC 2001). The impact of SO2 emissions on the environment is two-fold, with it changing the acid-base chemistry of the atmosphere; e.g. Galloway (1995), as well as altering the radiation balance of the Earth by changing the shortwave reflective properties of clouds which alters planetary albedo (Charlson et al. 1990; Charlson et al. 1987; Charlson et al. 1992).
As anthropogenic sources dominate SO2 emissions (Table 2.1), although to a lesser extent within Australia (Ayers and Granek 1997), the measurements of natural sources (aside from volcanism) have been less extensively researched. Additionally, temporal and spatial variations in emissions are highly variable, making individual measurements hard to extrapolate to global fluxes. Nonetheless, it is still worthwhile looking at these cases to examine the relative contributions and relative importance of these sources.
A list of the major sources of SO2 to the atmosphere is shown in Table 2.1. The sinks of
SO2 include its diffusion to the terrestrial and oceanic surface (as SO2 and SO4), and deposition through precipitation washout (again as SO2 and the oxidised SO4) (Kellogg et al. 1972). The efficiency of these processes is dependent on diurnal and seasonal patterns influenced by photochemical activity, cloud cover, rainfall and surface vegetation properties in the main (Ayers and Granek 1997).
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 19
Table 2.1. Global atmospheric emission estimates for SO2.
Source Flux Estimate Reference (Tg S yr-1)
Anthropogenic Sources Fossil fuels & industry 70 (Berresheim et al. 1995)
Biomass burning 2.8 (Berresheim et al. 1995)
Natural Sources Volcanoes (active) 7-8 (Berresheim and Jaeschke 1983)
Volcanoes (average) 0.5 (Berresheim and Jaeschke 1983)
TOTAL SOURCE 80.3-81.3
2.3.2 Hydrogen Sulfide
Prior to the late 1970s, H2S was considered the primary sulfur species being emitted to the atmosphere (Natusch and Slatt 1978; Rodhe 1972). This calculation, however, was based on estimations rather that any direct measurements, as a simple remedy to balance the global atmospheric sulfur budget (Watts 2000). The discovery that dimethylsulfide
(DMS) was being emitted from the worlds oceans at levels far exceeding any H2S measurements (Graedel 1979), meant that H2S has played a decreasing role in the global atmospheric sulfur budget, and is now confined to the discussion of local impacts (Shooter 1999). Nevertheless, reduced inorganic sulfur compounds are still key components in the production of reduced sulfur gases (Visscher 1996), and questions still remain about the role that H2S plays in the global sulfur cycle (Shooter 1999).
H2S is the dominant end-product of dissimilatory sulfate reduction, and although marine sediments are a major producer of H2S, very little escapes to the atmosphere because of its high rates of in-situ oxidation (Berresheim et al. 1995). Other significant contributions to the global H2S atmospheric budget are made by anoxic soils, wetlands (marine and freshwater), vegetation, as well as anthropogenic sources (Aneja and Cooper 1989; Bates et al. 1992). Table 2.2 lists the estimated annual global fluxes of
H2S to the atmosphere. Overall, the values shown are somewhat flexible, particularly regarding the wetland values, which vary in excess of 5 orders of magnitude (Bates et al. 1992). The primary source of biogenic H2S emissions to the atmosphere are
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 20
volcanoes and other geothermal activities (Berresheim et al. 1995), with the other sources playing a limited role within the global sulfur cycle. The primary sink for atmospheric H2S is its reaction with the free radical oxidants (Jaeschke et al. 1980); primarily hydroxide (OH), and to a lesser extent, the nitrogen free-radical (NO2), and ozone (O3) (Cox and Sandalls 1974). The reactions subsequently form SO2, and eventually SO4, leading to the mechanisms for deposition and wash-out described previously.
Table 2.2. Global atmospheric emission estimates for H2S.
Source Flux Estimate Reference(s) (Tg S yr-1)
Anthropogenic Sources
Biomass burning < 0.01 (Berresheim et al. 1995)
Natural Sources
Volcanoes / geothermal 1-2 (Pham et al. 1996)
Oceans < 0.3 (Yvon et al. 1993)
Wetlands 0.8-2.0 (Andreae and Jaeschke 1992; Aneja 1990)
(Andreae and Jaeschke 1992; Berresheim Vegetation (and soils) 0.5-4.3 et al. 1995)
TOTAL SOURCE 2.3-8.6
Despite these studies, questions still remain as to the role H2S plays in the global sulfur cycle, as well as its functioning in acid sulfate environments where it has yet to be researched. One of the first comprehensive assessments of biogenic reduced sulfur gas evolution from a wide variety of soils comes from Adams et al. (1981b). Adams and his research group found that H2S emissions varied substantially across and within the soil orders measured. Measurements from saline marshes ranged from 0.02 to 601.6 gS m-2yr-1, while histosols (peat) and inceptisols both measured less than 0.158 gS m-2yr-1. For further detailed reviews on the contributions of different sources and sinks to the global atmospheric H2S budget, refer to the following reviews (Aneja 1990; Aneja et al. 1982; Aneja and Cooper 1989; Berresheim et al. 1995; Jaeschke et al. 1980; Jaeschke et al. 1978; Shooter 1999; Watts 2000).
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 21
2.3.3 Dimethylsulfide
As mentioned previously, since its recognition in the early 1970s (Lovelock et al. 1972; Rasmussen 1974), DMS has been shown to be, by far, the major contributor to the global oceanic sulfur budget (Andreae and Raemdonck 1983; Turner et al. 1989). This has resulted in a large number of studies focussing on its levels across coastal shelf and open ocean areas compared to terrestrial sites; see Cooper and Matrai (1989), Wakeham and Dacey (1989), and references within. Despite this, the total marine flux of DMS is extremely variable, depending on the method of measurement and the season (Ayers et al. 1997; Ayers and Gillett 2000). Early estimates of the global oceanic DMS flux were quite varied, primarily existing within the lower range of 11.9 - 105.4 Tg S yr-1 (Andreae 1990); for example 15 Tg S yr-1 (Erickson et al. 1990). However, more recent analyses, including the use of remotely-sensed biogeophysical data, would suggest that the value was between 15 and 35 Tg S yr-1 (Kettle and Andreae 2000; Simo and Dachs 2002). In addition to the lack of research, along with the extreme variability in coastal areas, there is still uncertainty on the contributions of DMS from terrestrial sources (Cerqueira and Pio 1999). Estimates of the global atmospheric emissions for DMS are shown in Table 2.3.
Table 2.3. Global atmospheric emission estimates for DMS.
Source Flux Estimate Reference(s) (Tg S yr-1)
Anthropogenic Sources
Fossil Fuels / Industry 2.2* (Berresheim et al. 1995)
Natural Sources
(Kettle and Andreae 2000; Simo and Dachs Oceans 15-35 2002)
Wetlands 0.003-0.68 (Berresheim et al. 1995)
Vegetation (and soils) 0.05-0.16 (Berresheim et al. 1995)
TOTAL SOURCE 15-38 * Denotes total reduced sulfur compounds
In natural systems, DMS is produced by several distinct mechanisms. Of these mechanisms, the degradation of the tertiary sulfonium compound 3-
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 22
dimethylsufoniopropionate (DMSP) is thought to be the most important in open oceans (Kiene 1990; 1992; Murrell et al. 1996), coastal (Spartina alterniflora) salt marshes (Cooper et al. 1989; Dacey et al. 1987; Kiene and Capone 1988; Kiene and Visscher 1987; Morrison and Hines 1990), estuaries (Jorgensen and Okholm-Hansen 1985; Kiene and Service 1991), coral reefs (Broadbent and Jones 2006), and eutrophic marine basins (Wakeham et al. 1987). DMSP in an osmolyte which is synthesised in numerous species of marine phytoplankton (Keller et al. 1989), macroalgae (White 1982), as well as within certain higher plants (Dacey et al. 1987).
DMSP has been shown to readily degrade in a number of different environments including; anoxic marine sediments (Kiene and Taylor 1988), as well as open oceans. The enzymatic cleavage of DMSP results in the formation of DMS and acrylate (Cooper and Matrai 1989), as shown in Equation 2.19. The enzymes capable of cleaving DMSP are present in a diverse range of bacteria (Taylor and Visscher 1996) as well as in some phytoplankton (Stefels 2000).
CH3-S-CH2-CH2-COOH bbb CH3-S-CH3 + CH2=CH-COOH [2.19] (DMSP) (DMS) (Acrylate)
However, research has shown that DMS is only a minor product of DMSP metabolism within marine environments (Kiene 1996c; Simo et al. 2000), with an alternative pathway favoured. This alternative demethylation/demethiolation pathway of DMSP results in the formation of 3-methiolpropionate (MMPA) and methanethiol (MSH). Because of its preference over enzymatic cleavage, this process effectively controls the formation of DMS from DMSP within marine environments (Kiene et al. 2000). Part of this reaction involves the reversible redox conversion between DMS and dimethylsulfoxide (DMSO) (Taylor 1993). Indeed, it has also been recently demonstrated that many groups of marine bacteria reduce DMSO to DMS (Gonzalez et al. 1999; Zinder and Brock 1978b), a process which can also proceed chemically in the presence of sulfide (Zinder and Brock 1978a). However, only the bacterially-driven reduction has been observed in sediments (Kiene and Capone 1988).
An alternative process resulting in the production of DMS, proceeds via the transmethylation of thiols in anaerobic marine and freshwater sediments (Finster et al.
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 23
1990). This bacterial process occurs via the transfer of a methyl group which can originate either from the biomethylation of H2S (Equation 2.20 & 2.21) or by the degradation of organic sulfur precursors (Bak and Finster 1993; Kadota and Ishida 1972; Kiene and Capone 1988; Stets et al. 2004; Zinder and Brock 1978c). Studies have shown that marine-dominated sediments produce predominantly MSH, whereas freshwater samples produce primarily DMS (Bak and Finster 1993). This variation is thought to be the result of differences in the available sulfur pools in both types of sediments. In marine sediments, methyl groups are preferentially bound by excess H2S, whereas in freshwater samples where H2S is limiting, MSH can react with the methyl group to form DMS (Bak and Finster 1993).
R-OCH3 + H2S bbb R-OH + CH3-SH [2.20]
R-OH + CH3-SH bbb R-OH + CH3-S-CH3 [2.21]
(where R = aromatic residue)
Although the biochemistry and enzymatic pathways are not completely understood, the general processes of DMS formation from DMSP degradation within anoxic marine environments is summarised in Figure 2.1.
It is also possible for DMS to form (indirectly) from the degradation of methionine, although this yields primarily MSH (Kiene and Capone 1988; Kiene and Visscher 1987; Segal and Starkey 1969).
The primary atmospheric sink for DMS is the reaction with the hydroxide radical, with reaction kinetics influenced by heterogeneous reactions, particularly temperature dependence (Campolongo et al. 1999), as well as halogen chemistry (von Glasow and Crutzen 2004). As such, the reaction pathways are not as straight-forward as is the case with H2S oxidation.
Further information on biogenic DMS sources and processes can be found in the following review articles (Bentley and Chasteen 2004; Kadota and Ishida 1972; Kiene 1996a; Kiene 1996b; Taylor 1993; Yoch 2002).
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 24
Dimethylsulfoniopropionate (DMSP)
Cleavage Demethylation
Acrylate + 3-Methiolpropionate Dimethylsulfide (DMS) (MMPA)
Demethiolation
Demethylation Methanethiol Demethylation (MSH)
Methanogenic Archaea, Sulfate-reducing Bacteria, Denitrifying Bacterium
Methane + 3-Mercaptopropionate Hydrogen Sulfide (3-MPA)
Figure 2.1. Conversion of DMSP in anoxic marine sediments. Modified from van der Maarel and Hansen (1997).
2.3.4 Thiols
Thiols are intermediates in the microbial cycling of sulfur, playing essential biogeochemical roles in maintaining macromolecular structures, binding metals to active sites in enzymes, and acting as coenzymes (Mopper and Taylor 1986; Vairavamurthy and Mopper 1987). Also, because of their volatility, thiols play an important role in geochemical processes by forming strong complexes with some transition metal ions such as iron and copper, whilst mobilising others e.g. arsenic (Boulegue et al. 1982; Mopper and Taylor 1986). They have also been shown to play an integral part in the organic Hg speciation within interstitial waters (Zhang et al. 2004), with Harris et al. (2003) finding that methyl-Hg in fish is controlled by more toxic thiol- Hg complexes. This further highlights their importance in the binding of heavy metals and thus controlling their toxicity.
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 25
Many thiols occur within natural systems, with MSH and MMPA being dominant in anoxic coastal sediment porewaters (Kiene and Capone 1988; Kiene et al. 1990; Kiene and Taylor 1988; Luther III et al. 1986b; Mopper and Taylor 1986). Other thiols occur in lower concentrations; ethanethiol (ESH), monothioglycerol, 2-mercaptoethanol and 2-mercaptopyruvate, often in association with the solid-phase (Mopper and Taylor 1986).
Thiols can be formed via a number of distinct pathways. In natural sediments they appear to be principally derived from the microbial conversion of sulfur-containing amino acids such as methionine, see Equation 2.22 (Mopper and Taylor 1986; Zinder and Brock 1978c). Alternatively, they are able to form from the successive microbial demethylations of DMSP, without necessarily forming DMS (Equations 2.23, 2.24 & 2.25) (Yoch 2002); shown also in Figure 2.1. Demethylating bacteria have been identified under both aerobic (Visscher and Taylor 1994) and anaerobic conditions (van der Maarel et al. 1996).
Biotic Processes - CH3S-CH2-CH2-CH(NH2)-COOH bbb CH3SH + CH3-S-CH2-CH2-CO-COO [2.22] (Methionine) (MSH) (2-Keto-4-methiolbutyrate)
- CH3-S-CH2-CH2-COOH bbb CH3-S-CH2-CH2-COO + X-CH3 [2.23] (DMSP) (MMPA)
- - CH3-S-CH2-CH2-COO bbb H-S-CH2-CH2-COO + X-CH3 [2.24] (MMPA) (3-MPA)
- - CH3-S-CH2-CH2-COO bbb CH3SH + CH2=CH2-COO [2.25] (MMPA) (MSH) (Potentially Acrylate)
Note: X-CH3 represents an unidentified molecule with a terminal methyl group.
It is also possible for thiols to form via an abiotic Michael addition reaction between sulfide (or polysulfide) and unsaturated organic compounds such as acrylate, which can
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 26
be formed from the cleavage of DMSP (Mopper and Taylor 1986; Vairavamurthy and Mopper 1989). This reaction, which is shown in Equation 2.26, illustrates a pathway for the incorporation of sulfur into sedimentary organic matter during early diagenesis, and may be responsible for the formation of other thiols in sediments (Francois 1987; Vairavamurthy and Mopper 1987).
Abiotic Process - - CH2=CH-COOH + HS bbb HSCH2CH2COO [2.26] (Acrylate) (3-MPA)
Limited research oin the cycling of alkylsulfides (other than methylated compounds) has been undertaken (Sipma et al. 2004), especially with regards to acid sulfate environments. Clearly, thiols and DMS are inherently inter-linked within the geochemical (Vairavamurthy and Mopper 1987) and biological sulfur cycles.
2.3.5 Other Volatile Sulfur Compounds
The focus of this thesis is primarily on the four previously mentioned compounds. However, two other volatile compounds that should be mentioned because of their occurrence in natural environments include; carbonyl sulfide (COS), carbon disulfide
(CS2).
COS is produced by the photochemical degradation of organic sulfur compounds; such as DMSP, methionine and S-containing amino acids (Ferek and Andreae 1984). COS also affects the earth’s radiation balance by ultimately oxidising to sulfate aerosols within the stratosphere. Its atmospheric residence time is much longer than other volatile sulfur gases, surviving for a year or greater. COS has been measured in estuarine porewaters and within the water column (Bodenbender et al. 1999; Cutter and Radford-Knoery 1993).
CS2 has been shown to be emitted from coastal S. alterniflora marshes along with DMS (Bodenbender et al. 1999; Cooper et al. 1987a; Steudler and Peterson 1984). Numerous studies have also shown that CS2 is released from both unamended and treated soils under anaerobic conditions (Adams et al. 1979; Banwart and Bremner 1975; 1976a; b;
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 27
Minami and Fukushi 1981a; b), including waterlogged rice paddies (Nouchi et al. 1997; Yang et al. 1996).
Further information on the contributions of the two compounds to the global sulfur cycle as well as measured sources and impacts can be found in the following papers, and references within (Aneja et al. 1979; Watts 2000).
Chapter Two: Acid Sulfate Soils and Sulfur Gases - Background 28
Chapter Three: STUDY SITES BLACKS DRAIN & CUDGEN LAKE
3.1 INTRODUCTION
The ASS on the New South Wales north coast, Australia, are located in what were originally tidally influenced floodplain areas, adjacent to ecologically sensitive land and aquatic ecosystems. The large tidal range, small catchment size, and low outflows characteristic of eastern Australian embayments have resulted in the extensive deposition of typically fine-grained Holocene sediments (White et al. 1997). This is particularly relevant within the Tweed River catchment, where the lack of fluvial inputs resulted in shallow water depths persisting in the Tweed central basin for much of its infilling history (c. 7 to 3 ka) (Hashimoto 2005).
The Tweed region experiences a humid and subtropical climate with a pronounced wet season extending from December to March, and an average annual rainfall of > 1400 mm (Wilson et al. 1999). Much of the Tweed Valley floodplain was originally low-lying swampland dominated by grasslands, Casuarina and Tea-tree tidal swamps (Nielsen 1993). Because of the fertile soils and plentiful water resources, the eastern Australian coastal floodplains have experienced long periods of European agriculture, especially cattle grazing (King 1948), as cited in White et al. (2006). Since the 1960s though, the floodplain has been highly modified for sugarcane production, with the majority of backswamps being drained to increase land area and enhance the removal of surface runoff (White et al. 1993). Whilst receiving much of the attention for the general degradation of local waterways, the sugarcane industry was not solely to blame, with much of the drainage work being encouraged by various government agencies (White et al. 2006).
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 29
The New South Wales northern rivers region along with south eastern Queensland, are coming under increasing pressure from the traditional agricultural practices, such as sugarcane and dairy farming, in competition with a growing demand from urban expansion. The population in the Gold Coast – Tweed area is growing at one of the fastest rates within Australia, increasing by over 70,000 people from 2000 to 2005, or at an annual rate of 2.5 % (2004/5 figure) (ABS 2006). As a result, there remains little in the way of undisturbed areas for the examination of the underlying sediments and the processes of pedogenic transformation, as well as determining the effects of the current and projected growth. In fact, the Tweed floodplain has almost entirely been cleared for the cultivation of sugarcane and pastoral use (Pressey and Griffith 1992). With this in mind, the sediments within the Cudgen Lake Nature Reserve were chosen for the examination of VSCs in an undisturbed ASS system, as the reserve is part of the largest remnant of native vegetation along the northern NSW coastline (NPWS 1998). Therefore, in order to show the ASS characteristics under contrasting landuses, one study site was from a sugar cane area, the other from within Cudgen Lake Nature Reserve.
The general locations of the two sample sites (Blacks Drain and Cudgen Lake Nature Reserve) are shown below in Figure 3.1. Numerous research projects have been undertaken at various tributaries along the Tweed River (Smith 2005; van Oploo 2000; Wilson 1995), and should be consulted for further detail. This chapter examines the two locations with regards to the solid-phase sulfur chemistry and more general porewater chemistry, in an attempt to understand the basic processes operating within each system.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 30
Sampling Location
Pacific Ocean
Tweed River
Cudgen Lake
Blacks Drain
Figure 3.1. Locations of the two sampling sites on the NSW north coast; Blacks Drain and Cudgen Lake Nature Reserve. Top image courtesy of Geoscience Australia. Bottom image courtesy of Google Earth Beta v 4.0.
3.2 BLACKS DRAIN SITE
3.2.1 Background
The Blacks Drain sub-catchment (520 ha) is a right-hand tributary of the Tweed River, located in the north eastern corner of New South Wales (28°34′S, 153°38′E). The Tweed River catchment (< 1,100 km2) is relatively small when compared to others in the region, originating as a network of numerous sub-catchments draining the interior
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 31
caldera of the Tertiary Tweed volcano (Hashimoto 2005). The Tweed catchment’s valley-fill consists primarily of Holocene-age estuarine-deltaic deposits, excluding the Pleistocene and Holocene back barrier sediments along the coastline (Hashimoto 2005). A complete description of the morphostratigraphic evolution of the Tweed River can be found in Hashimoto (2005). Various field measurements were undertaken at Blacks Drain during June and December 2003, July 2004, October 2005 and May 2006, across several fallowed and growing sugarcane blocks.
The upper reaches of Blacks Drain primarily consist of pasture land for dairy cattle, with the lower reaches dominated by intensive sugarcane cultivation. The location for most of the sampling undertaken for measurements of the VSCs were under these intensively cultivated sugarcane paddocks, or alternatively, fallowed sugarcane fields.
3.2.2 Methods
Many of the methods contained within this section can be found in more detail in the references provided. Therefore, only brief descriptions are given below.
General methods All instantaneous measurements of pH and redox were made using intermediate junction electrodes (Ionode IJ44; Tennyson, Australia) connected to calibrated, handheld water quality meters (TPS WP-81 or 90FL-MV; Springwood, Australia).
Soil analysis Chromium reducible sulfur Chromium reducible sulfur (CRS) extractions were performed on field-frozen soil cores in a method similar to that outlined in Sullivan et al. (2000). Briefly, the method, which is derived from work by Canfield et al. (1986), involves the reduction of the total reduced inorganic sedimentary sulfur to H2S, which can be trapped and then measured by iodometric titration. The CRS extraction is a measure of the pyrite, elemental sulfur (S0) and acid volatile sulfur (AVS) content of a sample. As S0 was not able to be measured, it was assumed that the CRS (minus AVS) represented the pyritic fraction of
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 32
the soil, notwithstanding its limitations; particularly, the lack of mineral specificity (Rickard and Morse 2005) including the possible interference of sulfate being reduced by the technique (Mylon et al. 2002).
Reactive iron Reactive iron measurements were made on the same frozen soil cores to calculate the degree of pyritisation (DOP) of the sediment (shown in Equation 3.1). Reactive iron is a term used to describe the fraction of iron that readily reacts with sulfide (Canfield 1989). The term DOP was originally proposed by Berner (1970) to determine the limiting conditions (Fe or S) of the profile. Numerous methodologies for calculating the reactive iron content have been outlined in Raiswell et al. (1994). Because of the efficacy in extracting FeS (Cooper and Morse 1998b), a 24 hr extraction using 1 M HCl was used in accordance with that outlined by Billon et al. (2002).
FeCRS DOP (%) = x 100 [3.1] FeCRS + FeHCl
The DOP can strongly underestimate the degree of sulfidisation in sediments that have significant amounts of iron monosulfides (Boesen and Postma 1988). Therefore, the term degree of sulfidisation (DOS) was constructed from the DOP equation to provide a better indication of iron-limitation under such conditions (shown in Equation 3.2 below).
FeAVS + FeCRS DOS (%) = x 100 [3.2] Fe + Fe CRS HCl
Acid volatile sulfur Measurements of AVS were made as per the method outlined in Fyfe et al. (in press), a field-technique which also traps H2S using an overnight, cold hydrochloric acid (HCl) extraction. This diffusion-based method is an expansion on techniques previously well established (Brouwer and Murphy 1994; Hsieh and Shieh 1997; Hsieh and Yang 1989; Leonard et al. 1996; van Griethuysen et al. 2002). To summarise, the method dissolves iron monosulfides in 9 M HCl in an enclosed nitrogen-purged apparatus. The evolved
H2S precipitates as ZnS in an alkaline 3 % zinc acetate solution. After the extraction, the
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 33
ZnS is resuspended with acid and iodometrically titrated using a starch indicator. It should be explained that the term AVS primarily includes; amorphous iron monosulfides (FeS), mackinawite (~ FeS) and greigite (~ Fe3S4). Although it is recognised that the methodology and indeed the terminology AVS is rapidly becoming redundant, it was used as a basic qualitative indicator of transitory reduced sulfides. Refer to Morse and Rickard (2004), Rickard and Morse (2005) for a thorough discussion on the issue.
Porewater sampling and analysis Sample collection. In-situ dialysis membrane samplers, commonly referred to as ‘peepers’, were employed for the collection of all porewater samples. A schematic representation of the samplers is shown in Figure 3.2. Preparation of the peepers involved attaching inert polysulfone dialysis membrane (Gelman; 0.45 μm) to the thoroughly cleaned (1 month) polypropylene outer-casings, which contained de- oxygenated water prepared by the vigorous bubbling with high-purity nitrogen (N2) gas.
Once constructed, the peepers were continuously purged with high-purity N2 gas for no less than 72 hrs prior to inserting them into the sediment to ensure the complete removal of all but trace amounts of dissolved oxygen (Carignan et al. 1994). The peepers were inserted into the sediment for 10 days to allow for equilibration with the interstitial water (Hesslein 1976). After equilibration, the peepers were extracted from the sediment and the outer facings washed with de-ionised water. Further information on peeper construction, preparation, and application in ASS are detailed in van Oploo (2000).
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 34
Figure 3.2. Schematic diagram of in-situ porewater samplers or ‘peepers’.
Elemental analyses. Porewater samples for major cations and trace metals were transferred to pre-acidified (1N HNO3) polyethylene vials. Trace metal analyses were performed by ICP-MS (ELAN 6100; Perkin Elmer, Waltham, U.S.A), and major/base cations by ICP-OES (Optima 3000DV; Perkin Elmer) usually within 1 to 2 weeks from sample extraction. The trace metals analysed included (Al, Ba, Be, Ce Co, Ce, Mn, Sn, Ti, and Zn). Major/base cations included (Al, Ca, Fe, K, Na, Mg, Mn, S, Si, and Sr). Ferrous and ferric iron were measured by the phenanthroline method (3500-Fe B) set out in the APHA (1998). The colourimetry of the samples were analysed on either a portable spectrophotometer (HACH; Loveland, U.S.A), or by a laboratory-based spectrophotometer (UNICAM 8635; Cambridge, U.K), usually within 7 days of - - - - sampling. Porewaters were also analysed for sulfate and other anions (Br , Cl , I2 , NO2 ,
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 35
- and NO3 ) by ion chromatography using a Dionex AS14a column (Sunnyvale, U.S.A) and a sodium carbonate (8 x 10-3) / bicarbonate (1 x 10-3) eluent.
The acidity and alkalinity of the porewater samples were determined by titration in accordance with the APHA methods 2310B and 2320B (APHA 1998) within 6 hours of extracting the peepers.
3.2.3 Results and Discussion
It should be noted that only portions of the above mentioned experimental methodology will be shown and discussed within this section, primarily due to the relevance of the material towards the overall thesis.
Soil Characteristics The pH and redox measurements, shown in Figure 3.3, illustrate a typical agricultural ASS profile from the Blacks Drain study site. It is characterised by a dark brown / black (10YR2/1) organic topsoil with large amounts of fine roots and cane debris. Below this layer was a thin (< 50 mm) peat layer which was more acidic than the topsoil above it. This peat layer was observed to be ‘burnt-off’ within some of the upper, pasture- dominated reaches of the Blacks Drain catchment, a practice common in certain agricultural ASS areas (White et al. 1997). It had resulted in barren, scalded areas devoid of grasses or other vegetation. Beneath this peat layer came the oxidised ASS, a predominately clay matrix interspersed with varying amounts (5 - 30 %) of orange and red mottling, presumed to be iron oxides / hydroxides (oxidation products of pyrite). Also of note within this layer was the continuing presence of roots and root channels. The oxidation products were clearly concentrated around these existing and previously- used root channels as well as around shear planes. Yellow mottling (possibly jarosite) was also observed (< 10 %) within the oxidised ASS layer below the watertable. As can be seen in Figure 3.3, the pH reached a minimum at the bottom of this layer (~ 2.88). Beneath the oxidised layer was the transition zone, an area which exhibited characteristics of both the layers directly above and below it. The shift from acidic / oxidising to neutral / reducing conditions occurs within the transition zone, the bottom of which can be described as the unoxidised ASS. This material extended to the bottom
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 36
of the sampled profile and beyond, exhibiting a neutral pH and a progressively more reducing state. The unoxidised ASS was a characteristic blue-grey (Gley Chart 1-3/N) clay. Interestingly, both shear planes and root material were still present at these depths (> 1.4 m).
Figure 3.3. Soil pH and redox values for a profile at Blacks Drain from November 2005. Note; dashed lines represent layer changes within the profile, and SHE refers to standard hydrogen electrode.
The CRS (~ pyrite) content of the soil (Figure 3.4 a) was minimal within the first metre of the profile, after which it rapidly increased, eventually to a maximum of just over 1.8 % S (wet weight). Below 1.6 m and continuing into the gel the CRS content appeared to stabilise at about 1.5 % S. It should be noted that due to the strongly oxidising conditions of the profile, AVS would not be the preferential iron phase, and subsequently AVS measurements were below detection limits of the field-based technique used, and are subsequently not illustrated.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 37
Iron limitations within the sediment are assumed to occur at high DOP values (Raiswell and Berner 1985). However, this is not conclusive as the actual reactivity of the iron towards sulfide is not accounted for; see Canfield et al. (1992) for further discussion on this. The reactive iron and DOP values for the sampled profile at Blacks Drain are shown in Figure 3.4; b & c. Reactive iron was clearly not limiting pyrite formation within the upper horizons of the Blacks Drain site samples, with its formation likely constrained by sulfate reduction rates. Although the DOP shows the possibility of reactive iron limiting the formation of pyrite in the lower sections of the profile (> 1 m), the lack of reactive organic matter or sufficient oxidants within the lower part of the profile are more likely candidates. The increase in reactive Fe below 1.5 m may be due to the dissolution of Fe-rich clay minerals which can be leached by the 1 M HCl (Neumann et al. 2005), although it is more likely a result of the volcanic nature of the soils surrounding the Tweed valley. The DOP values for the clay-gel were on average 10 - 20 % higher than measured by Lin et al. (1998b) at McLeods Creek, another tributary (down stream) of the Tweed River. This data indicates a potential excess of sulfide within the lower horizons, enabling its accumulation and diffusive movement to other parts of the profile. The constant DOP levels at depths > 1.5 m suggests a cessation of pyrite formation (Berner 1970).
a. b. c.
Figure 3.4. Chromium reducible sulfur (CRS) as a percentage (a.), reactive iron (b.), and degree of pyritisation (DOP) as a percentage (c.) for a soil profile at Blacks Drain from November 2005. Note; dashed lines represent layer changes within the profile, and error bars equate to the standard error of the mean.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 38
Porewater Characteristics Selected components of the soil porewaters are shown in Figure 3.5. It should be noted that the illustrated porewater parameters (obtained using peepers) were taken at a different time (May 2003), and at slightly different locations (three cane paddocks northwards) to the soil samples. Therefore, interpretations and comparisons between the sediment and porewater chemistry will be merely qualitative.
The porewater profiles showed that the peaks of Fe and Mn occur within the transition zone, with the Mn maxima overlying the Fe maxima in observance with their reaction kinetics and thermodynamics (Burdige 1993; Stumm and Morgan 1996). The notable absence of Fe (and to a lesser extent Mn) within the oxidised zone was most likely due to their precipitation as insoluble iron oxides / hydroxides or removal as a consequence of upward leaching processes. Figure 3.5 b showed that the majority of interstitial iron was present as ferrous iron, with only trace amounts of Fe(III) being measured in the topsoil samples. The accumulation of Fe(II) within this area was most likely due to the periodic/cyclical reduction of iron oxides, after Fe(II) has diffused upwards from the anaerobic sediments below (see Huerta-Diaz et al. (1998) for a detailed discussion). The peak in Fe (and Mn) was therefore due to both their removal from the layers above and below the transition zone, via leaching and incorporation into sedimentary sulfides respectively.
The sulfate profile (Figure 3.5 c), as measured by ion chromatography, showed an increasing concentration in both the isopropyl alcohol (layered over sample) and 10% formaldehyde samples. These additives were originally used in an attempt to prevent the (chemical and biological) oxidation of the reduced inorganic sulfur species in the 2- 2- 2- porewaters; SO3 , thiosulfate (S2O3 ), and tetrathionate (S4O6 ). Although this proved unsuccessful in their measurement, it certainly had an effect on the sulfate concentrations, with the 10% formaldehyde samples being on average 45% greater than the isopropyl-layered ones. This information indicates that there is an additional source of sulfur being transformed within the porewaters between sampling and analysis, most 2- 2- 2- likely an oxidative process. As SO3 , S2O3 , and S4O6 were not detected, the most plausible answer would be that the sulfate was derived from the oxidation of other reduced inorganic sulfur compounds (S2- or S0), or from reduced organic S species. A
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 39
more detailed analysis using internal standards and controls clearly needs to be carried out with this method in order to quantify the magnitude of the oxidative processes and establish correct techniques for the future handling of porewaters containing reduced sulfides.
a. b. c.
Figure 3.5. Porewater total iron and manganese concentrations (a.), ferrous iron concentrations (b.), and sulfate concentrations (c.) for a peeper profile at Blacks Drain from May 2003. Note; dashed lines represent layer changes within the profile.
The fact that porewater sulfur (measured as sulfate) is not depleted within the lower layers of the profile (Figure 3.5 c), would agree with the DOP values in indicating that iron or organic matter, rather than sulfur , is limiting within this system. The values of interstitial S > 300 mg/L would confer a large sulfate reduction potential and possibly an excess of sulfide within the unoxidised ASS. Unfortunately, this assertion is not confirmed by the thermodynamic sulfur stability diagram for the samples (Figure 3.6).
Although the samples at depth show a distribution closer to H2S thermodynamic 2- stability, the dominant aqueous species remains SO4 . It should be noted that this does not preclude the formation of H2S, as clearly sulfate reduction is occurring at those depths. This idea is supported by the measurements of CRS, and the concomitant reductions in porewater trace metals (Co, Ni and Zn - data not shown) associated with
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 40
reduced sulfides, principally pyrite (Morse and Luther III 1999). The rudimentary parameters used for this diagram, particularly including the large variability in measurements derived from redox spear probes (Sposito 1989), would suggest that the values are somewhat flexible. One particular point to be drawn from Figure 3.6 and Equation 3.3 is that the speciation of the reduced sulfide, even at their highest pH values, will be as H2S.
+ - H2S = H + HS pK = 7.0 [3.3]
Upper / oxidising profile samples
Lower / reducing profile samples
-2 Figure 3.6. pH - p relationship of the Blacks Drain samples overlain by S-O2-H2O (25°C, 10 M) thermodynamic stability diagram. Adapted from Langmuir (1997) and Stumm and Morgan (1996).
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 41
3.3 CUDGEN LAKE NATURE RESERVE SITE
3.3.1 Background
Cudgen Lake (~ 1.7 km2) itself is a shallow (< 2 m), semi-tidal back barrier lake located on the east coast of Australia (28°20′S, 153°29′E) (Macdonald et al. 2004b), and is the catchment immediately south of the Tweed River. The lake acts as a natural retention basin for the surrounding watershed during flooding and its backswamp areas are subjected to extended periods of water logging (Roy 1973). As a result the vegetation association of the study site consists of a Melaleuca quinquenervia (paper bark) swamp forest, combined with a dense ground stratum of sedges (Lepironia articulata), grasses and ferns (NPWS 1998; Pressey and Griffith 1992).
Based on models of Australian east coast embayments (Dalrymple et al. 1992; Roy et al. 2001), Cudgen Lake can be described as an ‘immature’ embayment, still in the process of infilling, especially with the decreasing tidal influence. Also as a result of this, large areas of the tributaries draining into Cudgen Lake experience prolonged periods of inundation, with the lake itself serving primarily as a retention basin during flooding in the upstream catchment (Roy 1973).
Two different locations along the western edge of Cudgen Lake were studied at different points in time (shown in Figure 3.7). As can be seen from the map, the sampling during July 2004 was undertaken within a backswamp retention area of Cudgen Lake that occasionally fills with water during extended periods of rainfall. The retention basin is covered by a 10 mm thick layer of iron oxides/hydroxides, giving the orange/brown colour shown on the satellite photograph. Sampling prior to this in December 2002 was performed within 10 m of the aforementioned location, in a northerly direction away from the centre of the retention basin. Unfortunately, no AHD measurements could be made on the locations, but the December 2002 site was noticeably elevated (> 100 mm) compared to the July 2004 site, and located beneath flourishing Melaleuca’s.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 42
Western edge Dec 2002 of Cudgen Lake
July 2004
Clothiers Creek
Figure 3.7. Sampling locations within Cudgen Lake Nature Reserve. Original image courtesy of Google Earth Beta v 4.0).
3.3.2 Methods
All of the methods employed at the Cudgen Lake Nature Reserve study site were performed in the same manner as described for the Blacks Drain site. Refer to section 3.2.2 for details.
3.3.3 Results and Discussion
Soil Characteristics The pH and redox profiles, shown in Figure 3.8, illustrate a different system to that measured within the agricultural ASS at Blacks Drain. As is shown in the July 2004 sampling period, there is a limited (< 20 mm) brown oxidised organic crust on the surface of the profile, which is clearly the most acidic part due to the oxidation of iron sulfides. Below this is an unoxidised, intensely-fibrous organic layer of relatively neutral pH. The rapid increase in pH within this short boundary is likely due to the formation of iron monosulfides locking up acidity (Smith and Melville 2004), as in
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 43
Equation 3.4 below. This layer overlies the unoxidised clay-gel which was found to be interspersed with sand lenses, a point not unexpected owing to it proximity to the coast and likely depositional history. As mentioned previously, this area is intermittently overlain with water in periods of excess rainfall, and the watertable lies within the top 50 mm of the sediment surface. This ensures that reducing conditions exist throughout the entire profile in all but the top oxidised crust.
2- + 4FeOOH(s) + 4SO4 + 9CH2O + 8H bbb 4FeS(s) + 9CO2 + 15H2O [3.4]
The remainder of the samples taken for this thesis came from the July 2004 location, and so the focus of the discussion will remain on this site. However, the December 2002 pH and redox values were included (Figure 3.8 a) to show the transition away from the low-lying backswamp area. Within 10 m of the 2004 sampling point there is a major shift in the chemical characteristics of the ASS, due most likely to the influence of the growing vegetation (Melaleuca’s) impacting on watertable dynamics, and the associated accumulation of fluviatile sediments.
a. b.
Figure 3.8. pH and redox values (from peepers) for the December 2002 sample period (a.). Soil and porewater pH values for the July 2004 sample period (b.). Note; dashed lines represent layer changes within the profile, and error bars equate to the standard error of the mean.
The reduced sulfur fractions AVS and CRS are shown in Figure 3.9. As can be seen from both of these graphs, there is only minimal contribution from AVS to the total
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 44
reduced sulfur in all except the top two samples. The profile is dominated by CRS with extremely large accumulations (> 20 % dry weight ~ > 2.7 % wet weight) occurring within the fibrous layer. Within the unoxidised gel, the CRS concentrations return to values (0.75 to 1.5 % wet weight) comparable with that measured at Blacks Drain. As is shown in the DOP values (Figure 3.9 c) the Cudgen site profile exhibits a greater DOP than at the Blacks Drain site. The high DOP is more characteristic of euxinic depositional environments (such as the Black Sea) with depletion of the reactive iron contents limiting pyrite formation (Calvert and Karlin 1991; Canfield et al. 1996; Wijsman et al. 2001). Table 3.1 compares the measured DOP values for the Cudgen samples against other studies across varying sediment types.
Reactive iron appears to be limiting in all but the upper oxide crust. Because of the relatively small values of AVS in the profile, in comparison to CRS, the DOS is almost identical to the DOP (data not shown), suggesting an abundance of S intermediates facilitating an efficient conversion to pyrite. Although it cannot be discounted that organic matter may be limiting pyrite formation within these sediments, the extensive presence of decaying organics throughout the profile and nearby concentrations of organic carbon > 10 % (Macdonald et al. 2004b), suggests otherwise.
a. b. c.
Figure 3.9. AVS (a.), CRS (b.) and DOP (c.) from July 2004 soil profiles. Note; all values are as percentages of wet weight soil, and dashed lines represent layer changes within the profile.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 45
Table 3.1. Comparison of degrees of pyritisation between Cudgen Lake and other locations.
Site Location Sediment Type DOP Range Reference
Clearwater Lake, Lake sediments < 65 % (Huerta-Diaz et al. 1998) Canada Fresnaye Bay, Estuarine sediment 24-37% (Billon et al. 2001a) France Saguenay Ford, Riverine sediment 0.1-11.0% (Gagnon et al. 1995) Canada
Black Sea Euxinic sediment 20-78% (Canfield et al. 1996)
Long Is. Sound, Estuarine sediment 39.4-44.3% (Canfield et al. 1992) U.S.A Cudgen Lake Estuarine ASS 10.0-94.9% This study site, Australia
Porewater Characteristics The Fe and Mn porewater concentrations obtained from peeper sampling are shown in Figure 3.10 a, along with dissolved sulfur (Figure 3.10 b). The results show that the iron concentrations are comparable to the maximums reached at the Blacks Drain site (~ 120 mg/L), with the peak occurring in the unoxidised organic layer alongside the peak in CRS. Mn, also in similar concentrations to that measured at the Blacks Drain site, exhibits a peak just below that from Fe, but continues to increase down the profile. These trends would suggest several distinct mechanisms occurring within these sediments.
Firstly, the absence of Fe within the top 0.1 m is likely due to its incorporation as iron (oxy)hydroxides, or alternatively due to upward leaching. Below this, within the more reducing sediments, iron is more soluble (as Fe2+) and thus detected in the porewaters. Beneath this, its more gradual decline is due to incorporation into the solid-sulfide phase (FeS / FeS2). This is supported by the coincident peaks of porewater Fe with those of AVS and CRS, indicating that iron sulfides are controlling the source of interstitial Fe. The dual peaks in interstitial Fe at ~ 0.2 and 0.5 m are most likely due to the burial of sediments over time as well as seasonal redox changes, reducing overlying iron oxyhydroxides. Therefore, the Fe peak shape implies a recycling of Fe(II) and Fe(III) between areas of accumulation in favourable redox locations, followed by
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 46
diffusion transport or burial to differing oxidation potentials causing a continuing redox transformation cycle.
a. b.
Figure 3.10. pH values in addition to Fe and Mn (a.) and S (b.) concentrations from July 2004. Note; dashed lines represent layer changes within the profile.
In a similar fashion to Fe, Mn concentrations also decline within the top 0.1 m. Again, this is possibly due to its oxidation by O2 and formation as manganese(IV) oxides (Shaw et al. 1990), or more likely removed due to leaching, as Mn (oxy)hydroxides are unlikely to form within acidic conditions (Huerta-Diaz et al. 1998). It is interesting to note though, at depths where Mn is more soluble (Mn2+), there is only a stabilisation of the interstitial concentration rather than a decrease, as was observed in the Fe concentrations. This would indicate a more limited incorporation of Mn into the solid- sulfide phase, either as distinct MnS or more likely by coprecipitation / adsorption processes (Arakaki and Morse 1993; Morse and Luther III 1999).
Porewater sulfur shows some variation in the upper portion of the profile with the decreases likely due to its incorporation into the solid-phase, a point confirmed by the simultaneous lowering of the Fe values and concurrent measurements of CRS and AVS. Beneath this, the S concentration then steadily increases down the profile. It cannot be assumed that all of this S within the unoxidised layers will be present as sulfide, as
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 47
thermodynamics would suggest that sulfate persists within these sub-oxic depths, as well as being enhanced via downward diffusion. The concentrations of S are one-third lower than that measured at the Blacks Drain site, suggesting that trace metals in the Cudgen backswamp sediments are less likely to be bound as pure Me-S minerals, with metal-ligand formation of greater importance.
Some of the trace metals measured within the Cudgen sediments during 2004 are illustrated in Figure 3.11, showing their variability within the sediment porewater. Ni, Sn and Zn have accumulated within the porewaters of the brown oxidised surface samples, most likely due to lower pH and their association with Fe (oxy)hydroxides (Douglas and Adeney 2000; Petersen et al. 1995; Saulnier and Mucci 2000), or alternatively through the oxidation of organic matter (Gaillard et al. 1986).
Figure 3.11 a shows that below this layer, the concentrations of these trace metals (Ni, Zn; as well as Ag, Pb & Cu; not shown) decrease rapidly, where the AVS values are at their maximum. This would suggest that the trace metals are immobilised from the porewaters by co-precipitation and adsorption with FeS, or through the formation of discrete insoluble metal sulfides (MeS(s) shown in Equation 3.5), rather than regulation with oxyhydroxides (Huerta-Diaz et al. 1998; Simpson et al. 1998). There is substantial evidence that AVS (primarily FeS) is an important factor in the control of porewater heavy metals (Boulegue et al. 1982; Elderfield et al. 1981; Emerson et al. 1983), and therefore their toxicity in estuarine, marine and laboratory constructed sediments (Allen et al. 1993; Boothman et al. 2001; Cooper and Morse 1998a; Di Toro et al. 1992; Hansen et al. 1996; Helz et al. 2004; Morse 1994a; Saulnier and Mucci 2000). Di Toro et al. (1992) further identifies that only small concentrations of AVS are required (> 0.1 )mol/g) to sequester a significant quantity of metals. This factor, in combination with elevated DOP values, indicates that metal-sulfide formation is still an important sink of trace metals with the Cudgen Lake study site.
2+ 2+ Me + FeS(s) bbb MeS(s) + Fe [3.5]
Several of the trace metal profiles showed accumulations in the zone of sulfate reduction (i.e. < 0.2-0.4 m), with Cr and U shown in Figure 3.11 b. These measurements 2- 2- could potentially be the result of mobilisation by strong inorganic (e.g. CO3 , SO4 , or
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 48
HS-) or organic ligands (e.g. cysteine and glutathione) (Huerta-Diaz et al. 1998). Indeed, Huerta-Diaz et al. (1998) tentatively identified that thiols appeared to be particularly efficient in complexing Pb, but less so with regards to Ni, Cu and Mo. Alternatively their measurement could be the result of temporary oxygen penetration to these depths resulting in a non-steady state or transient peaks from oxidised iron sulfides (Shaw et al. 1990).
a. b.
Figure 3.11. Ni, Sn & Zn (a.), and Cr & U (b.) concentrations from July 2005. Note; dashed lines represent layer changes within the profile.
The levels of trace metals are of concern to current and future environmental health at this location (Macdonald et al. 2004b), as the formation of sulfides and oxyhydroxides are not a permanent sink for trace metals. Table 3.2 shows the trace metals that exceed the guidelines set out by the Australian and New Zealand Environment and Conservation Council (ANZECC) for the protection of 95% of species in marine waters. It should be noted that there are no trigger values for many of the heavy metals measured as part of this study (e.g. Be, Cs, La, Mo, Sn), potentially expanding this list. Perhaps more of a concern is that with the interstitial measurements which exceeded the trigger values, almost all (excluding Se) occurred within the top fraction of the sediment profile (see Table 3.2). The reductive dissolution of iron hydroxides (Equation 3.6) would be important within the Cudgen study site because of their propensity to release the trace metals accumulated through adsorption and co-precipitation processes (Petersen et al. 1995). Therefore, any disturbance which changes the pH or redox
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 49
equilibria, such as during an extended rainfall event, or alternatively the continued burial and reduction of these authigenic oxides could potentially release the metals to the overlying waters and the Cudgen Lake estuary.
+ - 2+ Fe(OH)3(s) + 3H + e bbb Fe + 3H2O [3.6]
Alternatively, depending on its concentration and reactivity, any oxidation of iron monosulfides (along with pyrite) and associated trace metals via could similarly lead to the liberation of trace metals to the water column (Calmano et al. 1994). Oxidation of reduced sulfides may come in the form of physical disturbances, seasonal water availability changes or as recently demonstrated by Choi et al. (2006), from plant- mediated oxygenation.
Table 3.2. Measurements of porewater concentrations at Cudgen Lake which exceed levels set out by the ANZECC Guidelines, and the depths in the profile at which they were taken.
Trigger values for 95% Concentration measured Depth Element species protection ()g/L) in porewater ()g/L) (mm)
Ag 1.4 24.90 10 Co 1 3.45 30 Cu 1.3 29.83 10 Pb 4.4 12.11 510 “ 4.4 9.16 10 Se 11a 55.60 770 Zn 15b 108.64 10
a No value for marine water. Value is for freshwaters b Figure may not protect species from chronic toxicity
Clearly, the early diagenesis of acid sulfate sediments has an important influence of the water quality within Cudgen Lake. Trace metals in the porewaters appear to be controlled by the solid phase sulfides, with various metal cations displacing Fe from amorphous FeS and forming stable metal sulfides (Cooper and Morse 1999). The processes of co-precipitation and adsorption also play an important role in the immobilisation of trace metals in such sediments (Elderfield et al. 1979). The oxidation of the pyritic fraction also needs to be considered in the trace metal dynamics at sites such as Cudgen, as the incorporation of such metals in authigenic FeS2 have been demonstrated to be an important diagenetic process (Morse 1994a; b; Vorlicek et al.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 50
2004). Reductive dissolution of oxyhydroxides also appear to have an effect on releasing trace metals within surface porewaters, allowing the downward diffusion to underlying horizons (Gaillard et al. 1986), as well as potentially exporting them to overlying waters and Cudgen Lake itself.
3.4 SUMMATION
The preliminary conclusions and interpretations are limited by the data taken, and need to be verified and augmented by a number of other measurements, particularly simultaneously extracted metals (SEM) and other extraction data. Similarly, spatial and seasonal variations are expected to occur with these trace metal measurements (Birch et al. 2001; Cooper and Morse 1998a; Grabowski et al. 2001), as well with the reduced sulfur measurements (Krairaponond et al. 1991; 1992; Panutrakul et al. 2001; van Griethuysen et al. 2003). Nonetheless, the porewater and sediment measurements still give a good basic indication of the diagenetic processes occurring in both systems.
The study sites present two contrasting acid sulfate environments that exist on the NSW northern coastline. The Cudgen site presents a location where the unoxidised material exists very close to the sediment-air surface, separated only by a thin highly oxidising surface crust. This represents the sort of situation that would arise on these estuary floodplain surfaces immediately after the marine sediment infilled surface emerged from below the flooded embayment after present sea level was established. Conversely, substantial alluvial inputs and differing geomorphically-induced hydrology have resulted in a greater degree of oxidation in the upper sections at the Blacks Drain site. The mobilisation of trace metals is more important within the Cudgen Lake backswamp samples when compared to those at Blacks Drain, as they are more likely to be associated with hydroxides and organic ligands. As the reducing sediments are close to the surface at the Cudgen site, small changes to water levels could result in oxidative events to the detriment of coastal communities.
Nonetheless similarities between the two sites exist. Comparisons of the underlying sediments of both locations show that the sulfidic sediments are similar, with both profiles showing evidence of the continued redox cycling, particularly of iron and
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 51
sulfur. The two catchments have experienced disparate geomorphic histories resulting in the above demonstrated chemical parameters.
Chapter Three: Study Sites - Blacks Drain & Cudgen Lake 52
Chapter Four: PASSIVE GAS ANALYSIS - SULFUR DIOXIDE FIELD MEASUREMENTS
4.1 INTRODUCTION
As mentioned in Chapter 2, in an attempt to balance acid-base budgets for ASS, van Breemen (1976; 1982; 1993) suggested that a large fraction of the acidity could potentially leave the ASS profile in an un-neutralised form as SO2. However, the emissions (concentration and flux) of SO2 from ASS were first quantified by Macdonald et al. (2004a) using a combination of static chambers and micrometeorological techniques at two fallow sugarcane blocks at McLeods Creek, a tributary of the Tweed River, about 5 km NW of the Cudgen site.
Therefore, an aim of this thesis work was to identify whether ASS areas other than sugarcane agriculture were also responsible for SO2 emissions, and to investigate potential seasonality effects between measurements made by Macdonald et al., and those made as part of this study during December and May. Simple passive diffusion samplers for determining SO2 in the atmosphere were used as part of this particular study, and their suitability was also examined for future research into SO2 emissions from ASS.
4.1.1 Passive Diffusion Samplers
Passive diffusion samplers were established as analytical instruments in the 1970s when
Palmes and Gunnison (1973) developed a tubular sampler for indoor, occupational SO2 exposure; see Cox (2003) for a comprehensive history on their evolution. Since then, passive diffusion samplers, also know as ferm tubes, have been employed for the
Chapter Four: Passive Gas Analysis - SO2 53
detection and quantification of SO2 within urban, industrial and more recently for environmental monitoring (Ayers et al. 1998; Carmichael et al. 2003; Ferm and Rodhe 1997; Ferm and Svanberg 1998; Macdonald et al. 2004a). They have also proven to be an effective method for measuring NO2, NH3 and many other trace gases (Ayers et al. 1998; Brown 2000; Ferm 1979; Ferm and Rodhe 1997; Ferm and Svanberg 1998; Gair and Penkett 1995; Gair et al. 1991; Gillett et al. 2000; Lan et al. 2004).
The basic theory behind the passive diffusion sampler is that an impregnated filter paper within a tube of known dimensions absorbs a desired gas, the concentration of which can be calculated by simple diffusion principles. Their distinct advantage over other techniques is that they are simple, relatively accurate and cost-effective, making them suitable for individual exposure monitoring as well as for broad scale atmospheric measurements (Cruz et al. 2005). The comparative characteristics of passive diffusion samplers compared to continuous monitoring devices are outlined in Krupa and Legge (2000).
Construction As can be seen in Figure 4.1, the samplers are a relatively uncomplicated apparatus, resembling a small snap-together polypropylene tube (∅ 25 mm). Within the polypropylene tube there are two filter papers as well as a stainless steel mesh. The stainless steel mesh sits at the diffusion inlet point of the sampler, acting as the primary buffer to particulates entering the tube, as well as ensuring the structural integrity of the Teflon and impregnated membrane and the tube itself. Similarly, the PTFE (Teflon) filter paper further minimises fine particulates, water and sulfate aerosols (Kasper-Giebl and Puxbaum 1999) from entering the tube. It also helps to ensure that the internal length of the tube is free from intrusion by turbulent eddies (Ferm 1991, as cited by Ayers et al. (1998)), vital to its theory of operation. The second filter paper acts as a trap for the desired gas, by converting it into a soluble form making it easier for analysis and thus quantification.
Chapter Four: Passive Gas Analysis - SO2 54
Snap-on polypropylene cap with hole for atmospheric diffusion
Stainless steel mesh
Teflon (1.0 )m) filter paper
Centre polypropylene ring piece
Paper filter coated with trapping solution
Polypropylene snap-one end-cap
Ø 21 mm
ASSEMBLED 10 mm DIMENSIONS
Ø 25 mm
Figure 4.1. Schematic representation of a passive diffusion sampler. Modified from Ferm (1991) as cited in Ayers et al. (1998).
Theory of Operation The passive diffusion samplers used for these experiments are an advancement of the technology employed by Palmes et al. (1976) in their personal sampler for nitrogen dioxide. The Palmes tubes were developed for high-dose indoor gas measurements. The major modification made by Ferm to the original Palmes tube was to shorten its length and increase its diameter to the dimensions shown in Figure 4.1. As these passive diffusion samplers are intended for outdoors, with much higher wind speeds, the
Chapter Four: Passive Gas Analysis - SO2 55
dimensional modifications were necessary to ensure that the measured gas was only transported by molecular and not turbulent diffusion (Ferm and Svanberg 1998).
Despite the modifications to the dimensions, the theory of operation behind all passive diffusion samplers remains the same. As their design relies on molecular diffusion through a static air layer, the samplers are capable of measuring gas samples without involving active air movement through or across their surface. As such, the sampling rate can be deciphered from the cross sectional area of the tube, perpendicular to the distance that the gas has to diffuse, which can be calculated by the integration of Fick’s first law of diffusion:
F = - D (dc/dt) [4.1]
Where F is the flux of the measured gas (g m-2 s-1), D is the diffusion coefficient of the gas (m2 s-1), and dc/dt the concentration gradient in the sampler. Furthermore, if the absorbent is sufficiently concentrated, and therefore an effectively infinite sink, the gas on the surface of the filter can be assumed to be zero, meaning the concentration gradient can be replaced by the outside concentration of the gas divided by the length of the sampler (Makkonen and Juntto 1997). To simplify then, the sampling rate of the tube is directly proportional to the cross-sectional area and inversely proportional to its length (Ferm and Rodhe 1997). When the effects of the stainless steel mesh and teflon filter paper are taken into account, the concentration of the gas outside the sampler can be calculated from:
C = X / (TD) (LR/AR + LF/AF + LM/AM + LLBL/AR) [4.2] where X is the amount of gas trapped on the filter (g), T the time of exposure (s), C the 2 -1 diffusion coefficient (m s ), LR the length of the sampler (m), AR the cross section of 2 the sampler (m ), LF the thickness of the teflon filter (m), AF the area of the holes on the 2 2 teflon filter (m ), LM steel mesh (m), AM the area of holes on the steel mesh (m ), and
LLBL the thickness of the laminar boundary layer (m).
Chapter Four: Passive Gas Analysis - SO2 56
In the case of SO2 measurements, NaOH is used as the trapping solution (Ayers et al.
1998). The conversion of gaseous SO2 to the measured SO4 takes place according to the stoichiometry of equation 4.3:
- + 2- 2SO2 + 2OH + O2 bbb 2H + 2SO4 [4.3]
The concentration of the dissolved sulfate, as measured using ion chromatography (IC), can then give the amount of SO2(g) as measured in the passive diffusion samplers (equation 4.4), which is a simplification of equation 4.2;
2- L × VE × [SO4 ] SO2(g) = [4.4] T × D
-1 2- Where; L is the total air resistance (41.2 m ), VE is the extraction volume (mL), [SO4 ] is the sulfate concentration (μmol/L), T is the sampling time (seconds), and D is the diffusion coefficient (1.32 x 10-5 m2 s-1). It should be noted that the diffusion coefficient for SO2 will vary from the effects of both temperature and pressure. The value of 1.32 x 10-5 m2 s-1 that was used for this analysis is the same as that determined by Macdonald et al. (2004a) in their passive diffusion experiments.
Even though the experimental work is not performed under ideal conditions, the figure for SO2(g) concentration can then be converted to ppbv using the ideal gas equation (Equation 4.5) to give an easily comparable concentration value.
n × R × T P = [4.5] V
Where; n is the number of moles (per m3), R is the gas constant (0.082061 atm mol-1 K- 1), T is the temperature during sampling (°K), and V is the volume of air (m3).
4.2 MATERIALS AND METHODS
4.2.1 Study Sites
Two separate field studies were conducted to measure atmospheric SO2 using passive diffusion samplers. The study sites for the two locations are detailed below;
Chapter Four: Passive Gas Analysis - SO2 57
December 2002
During December 2002, measurements of SO2 were made across five different landuse types within the Tweed and Cudgen catchments. These sites are outlined in Table 4.1, and are summarised below.
Table 4.1. Descriptions of the landuse for the five sample locations – December 2002. Landuse Catchment Description
‘Krasnozem’ Cudgen Lake Non-ASS; Fallow during time of gas measurements
Forest Cudgen Lake ASS; Organic soil within Melaleuca forest McLeods Ck Cane ASS; Under sugarcane cultivation at time of sampling (Tweed) Scald Cudgen Lake ASS; Bare ground inhabited only by clumps of native grass
Pasture Cudgen Lake ASS; Dairy cattle grazing land
The ‘Krasnozem’ soil was located within a hilly upland section of the Cudgen catchment. Although generally regarded as a red well-structured soil of an acidic nature, because of its volcanic parent material and elevation above sea level, no pyritic sediments were present in the underlying profile. Although fallowed at the time of the measurements, the farmland was used in rotation for intensive fruit and vegetable cropping. The ‘Forest’ soil was located within the Cudgen Lake Nature Reserve and was described in more detail in the previous chapter (see Section 3.3.3). The ‘Cane’ soil was located at the Tweed tributary of McLeods Creek. The samplers were set within sugarcane that was < 1.0 m tall, and thus having a very open canopy. The ‘Scald’ and ‘Pasture’ sites were located relatively close to one another within the upper catchment of Cudgen Lake. The exact locations of the five sites are shown in Figure 4.2.
Chapter Four: Passive Gas Analysis - SO2 58
Cane Tweed River Krasnozem
McLeods Creek
Cudgen Lake Pasture
Forest
Scald 1km
Figure 4.2. Location of the passive diffusion samplers during the December 2002 sampling period. Image courtesy of Google Earth - v 4.0 Beta.
At each sample location the ferm tubes were placed 0.2 m above the surface of the ground in a semi-protective casing to minimise meteorological effects such as sunlight, wind velocity, temperature and precipitation. Three sub-sample groups were situated within relatively close distance of each other, with each sub-sample containing 7 ferm tubes. The seven ferm tubes in each sub-sample were pooled together after exposure to ensure that sufficient quantities of SO2 were absorbed. All ferm tubes were exposed for a duration of 7 days.
May 2003 During May 2003, a similar set of experiments was made to that performed in
December 2002, looking at the variation in SO2 emissions across different landuses. However, chambers were used as opposed to the open-air constructions employed during the 2002 sampling period. In addition to this, a more detailed study was performed at the Blacks Drain study site, looking at SO2 emissions under a fallowed sugarcane paddock at differing temporal resolutions.
Chapter Four: Passive Gas Analysis - SO2 59
The five study sites examined are summarised in Table 4.2 below, with their locations (excluding the ‘Forest’ site) depicted in Figure 4.3. At each landuse location, three separate chambers were placed above the soil, each containing two ferm tubes. Although chambers create many artefacts that may not be necessarily valid in real-world situations, they provide a cost-effective, first assessment of gas concentrations (Denmead 1979; Gao and Yates 1998a; b; Rolston 1986). The chambers, being < 0.2 m2, were deemed to be small enough to allow adequate mixing by diffusion rather than by other means. For further discussion on the relative advantages of gas sampling methods, refer to Section 5.1.2. The chambers were placed at random intervals over the land surface for a period of 120 mins.
Table 4.2. Descriptions of the landuse for the five sample locations – May 2003. Landuse Catchment Description
Cane Blacks Drain ASS; Standing cane approx. 2.5 m in height
Fallow Blacks Drain ASS; Fallowed sugarcane paddock
Pasture Blacks Drain ASS; Grazing land infrequently occupied by cattle
Scald Blacks Drain ASS; Devoid of vegetation, previously pasture land
Forest Tweed River ASS; Melaleuca forest
The detailed SO2 measurements were made over a period of thirteen days at the ‘Fallow’ landuse location (see Figure 4.3). As opposed to the differing landuse measurements (described above), the passive diffusion samplers were not located within chambers. Daily SO2 measurements using passive diffusion samplers were made at both 0.5 and 1.5 m above the ground surface, in order to give an approximate indication of the flux of the gas. For each daily measurement, seven ferm tubes were exposed (24 hrs) after which they were grouped together to ensure sufficient quantities of the SO4 were present for analysis. For each weekly measurement, three ferm tubes were exposed (7 days) after which they were grouped for analysis. Both of these measurements were undertaken at the mast heights of 0.5 and 1.5 m.
Chapter Four: Passive Gas Analysis - SO2 60
N
Tweed River
Cane Fallow
Blacks Drain Pasture Scald
300m
Figure 4.3. Location of the passive diffusion samplers during the May 2003 sampling period. (2001 Air photo courtesy of Tweed Shire Council). Note; Forest location not located on map.
4.2.2 Methods: Passive Diffusion Samplers
The method that was used for the setup of the passive diffusion samplers is similar to that outlined by Ayers et al. (1998), and has been summarised below.
Assembly Prior to any field measurements, all components of the passive diffusion samplers (excluding the filter papers) were thoroughly cleaned, and assembled within the laboratory. The cleaning process involved soaking in a detergent bath (5 - 10 % Deacon 90) for > 24 hrs, after which they were rinsed repeatedly with high purity water (Milli- Q). Subsequent to the parts being dried in an oven, the samplers were assembled without incubating the filter papers, and stored in clean (acid and detergent washed) polythene canisters. A PTFE filter paper was used as the outer barrier (∅25 mm; 1.0 )m), whilst the cellulose filter paper (∅24 mm Whatman Grade No. 40) was impregnated with the NaOH trapping solution (both filters from Alltech Associates,
Chapter Four: Passive Gas Analysis - SO2 61
Sydney, Australia). All assembly/disassembly and handling of the ferm tubes required the use of sterile latex gloves and tweezers.
Impregnation Although constructed within the laboratory, the passive diffusion samplers were partially disassembled to allow for the incubation with the trapping solution immediately prior to being set out in the field each day. Additionally, the trapping solution was prepared new each day. The trapping solution consisted of dissolving 0.5 g AR-grade NaOH (UNIVAR Ajax; Sydney, Australia) in a small amount of deionised water, which was then made up to 50 mL with analytical grade methanol (UNIVAR Ajax). A 50 )L aliquot of this solution was evenly pipetted over the paper filter, leaving the passive diffusion sampler ready for measurement.
Analysis Upon the completion of the exposure period of the passive diffusion sampler, the paper filter was removed and placed in a clean polythene container and stored at 4°C for subsequent analysis. When the sample was ready to be analysed, the filter paper was extracted in 2 mL of Milli-Q water (per filter paper) and 20 )L chloroform as a preservative. The extracts were then analysed using a Dionex DX-500 ion chromatograph (Dionex, Sunnyvale, U.S.A) using an AS4A-SC anion column and
ASRS-ULTRA suppressor. The eluent is a mixture of 1.7 mmol/L NaHCO3 and
1.8 mmol/L NaCO3. IC analyses were performed at the Tweed Shire Water Laboratory.
4.3 RESULTS
4.3.1 December 2002
The results from the December 2002 field study across the five sites are shown in Figure 4.4. The concentrations measured by the passive diffusion samplers are surprisingly very low, and show little variation in the SO2 emissions between the five study locations. Even though the ‘Pasture’ site proved the highest and the non-ASS site (‘Krasnozem’) the lowest, there were no statistically significant differences between the
Chapter Four: Passive Gas Analysis - SO2 62
medians of the five study sites (P = 0.268 from Kruskal-Wallis Test; = 0.05). It should be recognised that the scald site was located within 200 m of a major roadway. Gilbert et al. (2007; 2003) found that beyond this distance, there was no influence of traffic on
NO2 measurements. Therefore, motor vehicle emissions cannot be discounted in the measurements of SO2 at the scald site during the 2002 study. All subsequent sample points were located well beyond 200 m from any roadway.
Figure 4.4. SO2 concentrations across different landuses, as measured by passive diffusion samplers, during December 2002. Note; the error bars equate to the standard error of the mean.
4.3.2 May 2003
Landuse Comparison The results from the passive diffusion samplers for the May 2003 study period are shown in Figure 4.5. Much greater concentrations were measured relative to the December 2002 values. The results also illustrate a much wider variation in the concentrations of SO2 across the five landuses when compared to the December 2002 sampling. Statistical analysis using the Kruskal-Wallis test showed that the medians of the five study sites were significantly different from one another (P = 0.032; = 0.05).
Chapter Four: Passive Gas Analysis - SO2 63
Additionally, the variation within each of the landuses is markedly greater as shown by the standard error of the mean. It should be reiterated that none of the locations of the five landuses were the same as that used in the December 2002 study. Therefore, any direct comparison is merely supposition. However, within both the 2002 and 2003 study periods, the ‘Pasture’ landuse ranked as the highest emitter of SO2 as measured by the ferm tubes; even with them being in different catchments and quantified by different exposure techniques. In the May 2003 study, this was followed by the ‘Fallow’ and ‘Forest’ categories respectively, after which came the ‘Cane’ and ‘Scald’ at less than one-tenth of the ‘Pasture’ concentration.
Figure 4.5. SO2 concentrations across different landuses, as measured by passive diffusion samplers, during May 2003. Note; the error bars equate to the standard error of the mean.
Although all five landuses are underlain with pyritic sediments, as with the majority of soil profiles, there is a large variation in the soil material at each location. Figure 4.6 shows the average pH and redox gradients down each of the profiles, along with the general layer boundaries. Large variations are clearly apparent in the boundary depths which may be contributing to variations in the SO2 concentrations that were measured, a point discussed further within Section 4.4.
Chapter Four: Passive Gas Analysis - SO2 64
Figure 4.6. Average soil pH and redox values for the different landuse site during the May 2003 sampling period.
The large disparities in the overall concentrations between the 2002 and 2003 sampling should also be noted at this point. Again, although the 2003 samples were taken within chambers and would be expected to exhibit greater concentrations than ambient sampling, the discrepancy between the two is somewhat curious.
Detailed Fallow SO2 Measurements The daily measurements from the ‘Fallow’ paddock at the two different heights (0.5 and 1.5 m) are shown in Figure 4.7. The negative concentrations shown in the graph are the
Chapter Four: Passive Gas Analysis - SO2 65
result of the blank measurements being higher than that measured when the passive diffusion samplers were exposed to the atmosphere. When the value of the 0.5 m measurement is greater than its counterpart (1.5 m) it can be inferred that there is an emission of SO2 from the ground surface. As can be seen in Figure 4.7, this process only occurs on two of the days that were sampled (14/5/03 and 20/5/03). During the remainder of the time, the concentration at the 1.5 m height was greater; suggesting the source of the measured SO2 was not the underlying ASS. It should be clearly stated that, the concentration difference does not entirely determine the magnitude of a gaseous flux. Rather, it is driven by the transfer coefficients for the gases, and therefore the energy transfer within the measured system. However, as this data was not measured at the time of sampling, this simplification of the data shown still provides useful insight into potential processes occurring.
Figure 4.7. SO2 concentrations at two different heights above the fallow soil surface (0.5 & 1.5 m) over a two-week period in May 2003.
Chapter Four: Passive Gas Analysis - SO2 66
Climatic Interactions
In order to try and understand what mechanisms were influencing the release of SO2 during the two days of positive flux values, the resulting measurements were equated to standard climatic variables. The meteorological data, which includes atmospheric pressure, relative humidity and cumulative daily rainfall, was obtained from the Coolangatta Airport (Australian Bureau of Meteorology) weather station. Figure 4.8 shows the variation in these parameters, along with the daily SO2 flux (bottom), around the time of sampling.
4.3.3 Errors Associated with Passive Diffusion Samplers
Extreme variation in the amounts of sulfate that was being detected, particularly within the samplers that were exposed for only a short period of time (less than a day), was the major concern with the results. Additionally, the concentration of sulfate within the controls often exceeded levels that were measured when exposed to the atmosphere. It is interesting to note that of the two sets of blanks/controls employed as part of this experiment, the ‘no-adsorbent / no-exposure’ blanks were far greater than the ‘adsorbent added / no-exposure’ blanks (range 0.5 - 93.9 %; average 13.1 %).
Therefore, as a consequence of these results, an attempt was made to characterise some of the errors that were possibly being introduced into the experiment. Briefly, the passive diffusion samplers were setup in a laboratory environment by the same procedures as per the two previous field sampling regimes. However, certain parameters were manipulated in an attempt to identify any error-contributing processes that could be eliminated from future sampling work. Factors in the experiment protocol that were looked at included; the atmosphere of the oven in which the filter papers were dried in (anoxic vs. oxic), the type of container used to transport the filter papers to the laboratory for analysis, and the influence of the gloves that were worn whilst the ferm tubes were assembled.
Chapter Four: Passive Gas Analysis - SO2 67
Figure 4.8. Atmospheric pressure, relative humidity, and rainfall for the time period before and after sampling, along with the SO2 flux for the May 2003 sampling period.
Chapter Four: Passive Gas Analysis - SO2 68
The results (Table 4.3) showed that all samples contained levels of sulfate below detection the detection limits of the IC, except for those that were intentionally handled with the latex gloves. These samples had levels of sulfate of between 3 and 4 times the detection limits of the IC. Therefore, whilst the majority of the protocol proved to be robust, clearly, some error is being introduced with the use of the particular latex gloves in the passive diffusion sampler assembly. This result clearly has implications for other aspects of trace-sulfate experimental work in which such gloves are regularly used.
Table 4.3. Average errors intentionally introduced to the passive diffusion sampling method. Parameter Sulfate (mg/L)
Anoxic drying environment < 0.03
Oxic drying environment < 0.03
Transport container 1 < 0.03
Transport container 2 < 0.03
Excessive glove handling 0.09 - 0.12
It is also worth noting, that as part of the detailed measurements made over the fallow area during 2003, weekly measurements at the two different heights (0.5 and 1.5 m; 3 ferm tubes at each) were also made (data not shown). The weekly measurements were employed with the aim of verifying the method through the addition of the daily measurements. However, the weekly samplers showed substantially lower values than the additive daily measurements; only accounting for between 8.6 - 51.1 % (week 1) and 0 - 3.2 % (week 2) of the added daily measurements. Although this goes some way in explaining the variation between the duplicates, it does not account for the reason why the blanks were higher than the exposed samples. It would appear from the results that exposure and in particular, extended exposure, was reducing the concentrations of
SO2 sorbed onto the filter papers. This point is expanded on within the discussion.
4.4 DISCUSSIONS & PRELIMINARY CONCLUSIONS
It needs to be reiterated that the results from both the 2002 and 2003 study were only preliminary measurements of SO2. The experimental rigour and results obtained was such that only qualitative interpretations will be made. However, the combined data
Chapter Four: Passive Gas Analysis - SO2 69
from the sampling periods, would suggest that ASS are a potential source of SO2; with fluxes occurring sporadically in pulses rather than consistently. This is in agreement with the results found on similar soils by Macdonald and colleagues (2004a). As for the delineating the contribution of the varying ASS landuses on the SO2 emissions the results are less conclusive. It would appear though, that grassed pasture overlying ASS could be the major emitter of SO2 within these environments.
The more detailed May 2003 ‘Fallow’ experimental work showed some indication that levels of SO2 are influenced by climatic variables, particularly rainfall and atmospheric pressure. Macdonald et al. (2004a) found that SO2 emissions increased, albeit somewhat variably, with increasing evaporation rates. It was suggested in their study that this could be due to the evaporation of soil waters containing sulfite (as introduced in Section 2.2.1). Their suggestion was accompanied with measurements of trace amounts of sulfite, a transient sulfur species that has been detected by many others in sulfur dominated environments; for example Gagnon et al. (1996) and Thamdrup et al. (1994a).
This process, and hence climatic dependence may account, or at least partially account, for the resultant emissions of SO2 measured in this study. Indeed, the peak emission of
SO2 comes during a period of low rainfall (< 10 mm), a time when the soil profile would be experiencing elevated surface evaporation without being inundated. Sulfite can then accumulate within anoxic areas where molecular oxygen is limited, causing an incomplete oxidation of sulfide (Gagnon et al. 1996), shown in Equation 4.6;
- 3 2- + HS + /2O2 bbb SO3 + H [4.6]
The decrease in the emission of SO2 on the next day is explained by the escalation of the subsequent rainfall event during that day (> 80 mm). This event completely saturated the profile resulting in rainwater pooling on the ground surface for over 24 hr.
SO2 is extremely soluble in water, being over 40 times more soluble than CO2, and 60 times more soluble than N2O at 20°C (Broecker and Peng 1974). Therefore, any SO2 conversion or production would simply be dissolved rather than exist in a gaseous state.
Chapter Four: Passive Gas Analysis - SO2 70
The peak in SO2 emission also occurs during a time when the atmospheric pressure is decreasing from a high of ~ 102.5 kPa to ~ 101.5 kPa (Figure 4.8). The possibility of atmospheric pressure change as a causal influence on SO2 emissions warrants further investigation. Commonly referred to as barometric- or atmospheric pumping, the influence of atmospheric pressure has been shown to be important in gas emissions from both surface (naturally occurring and disturbed) and subsurface soils (Elberling et al. 1998). Although dependent on a variety of different parameters, one early study (Clements and Wilkenin 1974) demonstrated that a shift in atmospheric pressure (~ 1-2 %; in the order of that shown above) caused a 20-60 % flux in 222Rn from the soil profile. Another example (Czepiel et al. 2003) demonstrated that CH4 emissions from landfill sites: showing that gaseous measurements of CH4 were closely, negatively correlated (R2 = 0.95) with surface atmospheric pressure. Is it worth noting that White et al. (2005) found that atmospheric pressure changes affected the elevation of the shallow groundwater in ASS. If a decreasing ambient pressure raises the water table, it logically follows then that this would displace air from above the watertable into the atmosphere.
It also should be acknowledged that even though there appears to be some relationships between the climatic variables and the flux of SO2, the interpretation needs to be taken with extreme caution. The meteorological variables are not nearly detailed enough to provide accurate interpretations, and possibly more so, these measurements were taken from the meteorological station located approximately 20 km northeast of the study site.
Methodological Issues Clearly, another major point of discussion to arise from these results is the suitability of passive diffusion samplers for the analysis of short-term SO2 emissions from ASS. The extreme variability in the chamber results in May 2003 (see Figure 4.5) shows that the experimental setup with the ferm tubes in closed chambers being not well suited to short-term measurements of SO2.
This point is also observed in the literature, with the majority of passive diffusion experimental programs focussing on long-term exposures of between 24 hrs to several months (Ayers et al. 1998; Makkonen and Juntto 1997; Warashina et al. 2001). Data from Ayers et al. (1998) showed a much reduced percentage-error between weekly
Chapter Four: Passive Gas Analysis - SO2 71
exposed pairs of passive diffusion samplers (~ 10 % at 5 ppbv and ~ 25% at 0.5 ppbv). Therefore, within the conducted study the comparison of added daily exposures proving substantially greater than weekly exposures suggests an inordinate level of variability in the method, or as mentioned previously, a possible methodological problem, with increased exposure time somehow reducing the sensitivity of the samplers. In addition to being better suited to longer-term exposures, passive diffusion samplers are also more adapted to higher levels of SO2 exposure, such as their use for monitoring urban air pollution (Plaisance et al. 2002), including personal exposure levels (Lee et al. 2004), as well as shown above by Ayers et al. (1998).
There are several other issues within the methodology that need to be corrected before employing this technique for future measurements of SO2 from ASS. In reference to the chamber measurements, as headspace concentration is time-dependent, the fact that no interval measurements were made precludes the possibility of determining possible flux values from the soils. Also, large temperature fluctuations would be anticipated within the chambers over the 2 hr sampling period. As no temperature measurements within the chambers were made, the extent of this error is uncertain.
The influence of the chamber surface on possible changes to ambient S-gas concentrations (most likely reductions) needs to be assessed. Although not as reactive as reduced sulfur compounds, the measurements of SO2 from closed chambers would be better taken from an inert material such as from a teflon-coated chamber, rather than the
PVC used. The return influence imposed on the soil surface by increasing SO2, and other gas, concentrations within the headspace is also an issue that needs further understanding. The most obvious example of this is that an increase in the headspace concentration would result in suppressing the concentration gradient of the gas, as was identified by Drewitt et al. (2002) with emissions of CO2 from a forest soil. An indirect effect that would also need to be considered is the impact of the decreased gradient concentration within the soil profile. In a study by Conen and Smith (2000), the impact of closed chambers was shown to increase N2O emissions to depths of over 0.3 m in the soil over which the chamber was set. This variation was observed to be primarily dependent on the air-filled porosity / moisture content of the soil. Many of these points are further discussed in the introduction to Chapter 5.
Chapter Four: Passive Gas Analysis - SO2 72
One area that may improve the technique regards the choice of sorbent, which is a critical factor in the performance of passive diffusion samplers (Brown 2000). Indeed, several other compounds, other than NaOH, have been used over the last decade as impregnating agents in the measurement of SO2 using diffusive samplers. Examples of these include 5% Na2CO3 (Cruz et al. 2005; Cruz et al. 2004; Lan et al. 2004;
Warashina et al. 2001) with added glycerol (Lee et al. 2004); NaHCO3 (Perkauskas and Mikelinskiene 1998); and between 10 and 20% triethanolamine (Kasper-Giebl et al. 1999; Krochmal and Kalina 1997; Plaisance et al. 2002), which has the added advantage of adsorbing SO2 and NO2 simultaneously. A comprehensive examination of the effectiveness of the different trapping solutions has been summarised since the field work was undertaken by (Cruz et al. 2005), who demonstrated that Na2CO3 was the most suitable absorbing medium for SO2 in humid environments. Future sampling within sub-tropical ASS of northeast NSW should consider these alternatives.
Other procedural aspects that may improve the accuracy of the method include the use of H2O2 as an extracting agent (post-exposure) to completely oxidise the adsorbed products to sulfate (Cruz et al. 2005; Krochmal and Kalina 1997; Lan et al. 2004; Lee et al. 2004; Plaisance et al. 2002). A greater number of simultaneous sampling measurements would be also beneficial, such as the 600 parallel samples undertaken by (Perkauskas and Mikelinskiene 1998) for the measurement of urban air quality in Lithuania.
Concluding Statements Although the passive diffusion samplers are simple enough to employ in field- monitoring studies, there is a greater need for optimisation and validation than undertaken as part of this study. Any subsequent research on SO2 emissions using the diffusive samplers need to sort out the issues identified above, as well as sufficiently integrating the measurements with appropriate meteorological data, before making any interpretation of the data or identifying possible causal factors.
Despite the variability in the results and the uncertainty of the passive diffusion sampler method, there is an efflux of SO2 coming from the sampled ASS worthy of further investigation. Because the temporal resolution of the samplers is limited, an alternative
Chapter Four: Passive Gas Analysis - SO2 73
method for real-time SO2 concentrations would prove much more beneficial. Additionally this would require analytical instruments with a much lower detection limit.
Chapter Four: Passive Gas Analysis - SO2 74
Chapter Five: ADVANCED FIELD MEASUREMENTS HYDROGEN SULFIDE & SULFUR DIOXIDE
5.1 INTRODUCTION
The quantification of surface-atmosphere trace gas exchange is important for further understanding the role played by the lithosphere and biosphere in regulating the global climatic balance. Because of the turbulent nature of the atmosphere at the air-surface boundary, trace gases are rapidly diffused to and from the surface. Therefore, identification of localised emissions is important in the further understanding of larger budgets of trace gas species. The results from the passive diffusion samplers suggest that ASS are indeed a potential source of SO2, confirming that reported by Macdonald et al. (2004a).
This chapter attempts to quantify these emissions of SO2 on a more accurate scale than that made with passive samplers, by using flux-gradient micrometeorological techniques in conjunction with an ultraviolet (UV) fluorescent analyser. Because of its perceived importance in terrestrial S-gas emissions, along with its ability to be easily used in tandem with the SO2 measurements, H2S measurements were also made using UV fluorescence. The ultimate aim of the chapter is to establish diurnal fluctuations and relate meteorological parameters to high temporal resolution H2S and SO2 concentrations and fluxes from an agricultural ASS under sugarcane cropping.
Therefore, the chapter is structured to include a background and then explanation of the advanced materials and methods used, followed by detailing the measurements from the two field trips undertaken (Nov/Dec 2003 and Oct/Nov 2005), which have been divided
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 75
into two separate results sections. After this is a combined discussion of the real-time concentrations and fluxes of SO2 and H2S, and preliminary conclusions.
5.1.1 Micrometeorological Techniques
Overview According to Arya (1988), micrometeorology is the branch of meteorology that deals with the atmospheric phenomena and processes on the micro- or local- scale. More specifically, micrometeorology is limited to only those phenomena that originate within the shallow layer of frictional influence in contact with the earth’s surface, thus dealing with the exchange processes of radiation, mass and momentum that dictate much of what is commonly referred to as the weather (Arya 1988; Moncrieff et al. 1997). This surface-atmosphere area of interest is commonly referred to as the Atmospheric Boundary Layer (ABL).
The ABL is formed from the interaction of the earth’s surface and the atmosphere over relatively short time-spans (Arya 1988). This contact between the atmosphere and land surface is quickly and efficiently transferred through the mechanism of turbulent diffusion or mixing (Arya 1988). As such, micrometeorological methods used to quantify this turbulent exchange can be divided into two categories; direct methods, which sample the air as it flows past a specific point; and indirect methods, which are based on quantifying the rate of diffusion across a concentration gradient (Moncrieff et al. 1997). Whilst each method has their inherent advantages and disadvantages, the flux-gradient methodology was chosen for this study because of the equipment available and its suitability to the location (homogeneity and lack of vegetation canopy), and the diurnal measurements desired. More detailed information on the alternate techniques can be found in Lenschow (1995).
General Flux-Gradient Theory The principle tenet underlying flux gradient methods is that the gas of interest is transported by a process of turbulent diffusion across a gradient of mean concentration (Arya 1988; Sutton 1953). The vertical flux density of the gas (C) in the lower
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 76
atmosphere, as derived from Fick’s law of diffusion (identified in Chapter 4), can be written as;
Fc = -ρρρ a Kc (dc/dx) [5.1]
Where ρa is the density of dry air, Kc is the diffusion coefficient of the gas, and dc/dx is the vertical gradient in the constant flux layer. The diffusion coefficient can be calculated from the exchange of energy at the ground surface, usually by the simultaneous measurement of a tracer entity whose flux and thus gradient is known, such as heat, mass or momentum (Denmead and Raupach 1993). This model is based on the premise that the transport mechanisms are similar for differing scalar fluxes, including gaseous transport (Denmead and Raupach 1993; Moncrieff et al. 1997).
The theory infers that a lengthy homogeneous exchange surface is required for the air layer of interest to traverse over, whereby small changes in vertical concentrations can be rapidly detected. As such, the technique cannot be applied to measure fluxes in closed-canopy systems because of the counter-gradient transport mechanisms generated (Baldocchi et al. 1988). The height of the upper boundary of the air layer, which is modified by an emission at ground surface, is approximately one-tenth the distance of the fetch (Sutton 1953). Therefore, the general rule of thumb is that the fetch/height ratio should approximate 100:1, i.e. the fetch length should equal 100 times the measurement height (Businger 1982). Although, this ratio has been deemed by Horst and Weil (1994) to be insufficiently small, even in moderately stable conditions. Figure 5.1 diagrammatically represents the fundamental setup behind the flux-gradient micrometeorological technique, which was used for the following experimental work.
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 77
Micrometeorological measurements
Prevailing wind direction
z2 Air Sampling Lines
z1 Gas Analysers Area of interest
Fetch (100 × z2)
Figure 5.1. Diagrammatical simplification of micrometeorological theory. Meteorological instrumentation reproduced with permission courtesy of Campbell Scientific.
This method has been widely used in the measurement of ammonia fluxes from pasture (Denmead et al. 1974) and agricultural fields (Denmead et al. 1978), nitrogen gases
(NOx and N2O) from sugarcane (Denmead et al. 2005), and sulfur gases from pasture (Bartell et al. 1993), to name a few.
5.1.2 Comparison with Alternative Techniques: Chambers
Chamber Theory The underlying principle behind using chambers is to restrict the exchange of air with the atmosphere so that minor changes in the concentration of the emitted gas can be concentrated within the headspace (Denmead and Raupach 1993). When the chamber is closed, the rate of concentration change within the headspace is used to calculate the gas flux to or from the soil system, via the following relationship;
Fg = (V/A) d-g/dt [5.2]
Where Fg is the flux density of the gas, V is the volume of the headspace, A is the area enclosed by the chamber, -g is the density of the gas within the chamber, and t is the
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 78
time. In a closed system, the d-g/dt (change in concentration over time) will increase linearly until reaching a critical level, where the gas within the headspace will begin to influence the flux from the soil.
Relative Merits of Closed Chambers Both Hutchinson and Livingston (1993), and Livingston and Hutchinson (1995), provide excellent reviews on the use of differing chamber systems to measure trace gas fluxes, including associated benefits and drawbacks.
One of the major disadvantages of the closed chamber method, as opposed to dynamic chambers, as described by Reichman and Rolston (2002), are the artifacts imposed on the chamber’s microclimate, affecting the gas exchange of the soil (Denmead 1979; Denmead and Raupach 1993). Chambers, once placed on the ground surface, have the ability to greatly change many climatic variables, including the ambient temperature, humidity, and particularly the natural wind profile near the surface (Jury et al. 1982). Given that the production of many trace gases in soil is the result of biological activity, their production is strongly temperature dependent (Denmead and Raupach 1993). Therefore, equilibrating the internal temperature to the diurnal fluctuations outside the chamber is extremely important, and also very difficult. For this reason, it is much easier to employ several chambers for short periods of time (Denmead and Raupach 1993).
Another source of error encountered with the use of chambers, is due to the inherent heterogeneity in gas emissions from soils. As the chambers usually only occupy less than 1 m2 of ground surface, the variability of soil gas emissions tends to be very large (Denmead 1994; Denmead and Raupach 1993). Folorunso and Rolston (1984) show the extreme variation in nitrous oxide fluxes from chamber methods, on a well-drained alluvial soil. Little difference would be expected from sulfur gas emissions from an ASS.
Another error incurred with the chamber method is the disturbance of the ground surface beneath the chamber. The insertion of a chamber into the ground surface may potentially increase the measured emission of a trace gas such as SO2 or H2S. Matthias
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 79
et al. (1980) found an increase in nitrous oxide concentrations when the chamber was placed over a soil disturbance. Jury et al. (1982) suggests that although this is not representative of an increase in nitrous oxide production, it may create cracks in the soil adjacent to the chamber allowing for an increase in the effective gas diffusion co- efficient through the soil. The errors associated with the chamber methodology have been extensively quantified for nitrous oxide by Jury et al. (1982).
The inherent reactivity of sulfur gases in general poses another problem with chamber methods, with the potential for the gas loss on the chamber surface itself. Kuster and
Goldan (1987) quantified the losses of several sulfur gases including H2S and DMS from different chamber wall materials. They found that all of the materials tested (pyrex, polycarbonate and teflon) resulted in significant losses, with teflon proving the most suitable.
Whilst is it easy to lambaste the deficiencies of chamber methods, they provide an economic alternative to expensive and laborious micrometeorological methods (Hutchinson and Livingston 1993), and should not be discounted for the measurement of trace sulfur gases. Chambers are ideal for process studies as well as for experiments using different treatments (Denmead 1994). Indeed, they would be the preferred technique in natural-state, remote areas such as the Cudgen Lake site.
Summation The micrometeorological method has several inherent problems including those associated with rainfall interference; along with various night-time issues, such as the unreliability of anemometers at low wind speeds and imprecise net radiometers (Denmead and Raupach 1993). Despite these potential problems, however, it was decided that the technique had a unique potential to provide insight into the real-time flux of gases from the soil to the atmosphere. Furthermore, using micrometeorology methods allows for the examination of trace sulfur gas fluxes to be integrated over a relatively large area, without harming the area or vegetative crop being analysed. Consequently, this chapter examines the employment of active gas samplers as part of the micrometeorological measurements of both SO2 and H2S at two separate times.
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5.2 MATERIALS AND METHODS
5.2.1 Study Sites
Two separate field missions looking at the active analysis of SO2 and H2S were performed at the Blacks Drain study site; one in Nov/Dec 2003, and the other in Oct/Nov 2005. Their locations are shown in Figure 5.2.
N
Tweed River Nov/Dec 2003
Oct/Nov Blacks Drain 2005
300m
Figure 5.2. Location of the active gas analysis sites within the Blacks Drain area. 2001 air photo courtesy of Tweed Shire Council.
Nov/Dec 2003 Measurements during the Nov/Dec 2003 interval were undertaken within a single fallowed sugarcane block, across a 16-day period (29th Nov - 14th Dec). This time period is considered to be the beginning of the northern NSW rivers wet season (Dec to Mar), during which the majority of the areas average rainfall (> 1400 mm) occurs (Wilson et al. 1999).
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Oct/Nov 2005 Measurements during the Oct/Nov 2005 interval were also undertaken within a single fallowed sugarcane block, across an 18-day period (16th Oct – 2nd Nov). As can be seen from Figure 5.2, the study sites varied only slightly in location (< 500 m apart).
5.2.2 Gas Analysers
SO2
Active measurements of SO2 were made using a pulsed-fluorescence sulfur dioxide analyser (Model 9850, Monitor Labs, U.S.A). The analyser is an ultraviolet (UV) fluorescence spectrometer designed to continuously monitor low concentrations of SO2 in ambient air (0.001 ppm). The theory of operation follows that SO2 exhibits a strong ultraviolet absorption spectrum between 200 to 240 nm. UV radiation at 214 nm from a zinc discharge lamp within the analyser is separated from the other wavelengths in the zinc spectrum using a UV bandpass filter. The 214 nm radiation is focused into the fluorescence cell where it interacts with SO2 molecules injected into the beam path. The resulting fluorescence is collected and focused onto a photomultiplier tube, measuring the fluctuations from the SO2 interactions. On each occasion, the analyser was calibrated by Ecotech (Knoxfield, Australia), using factory-based calibration equipment.
The USEPA has designated the analyser as an Equivalent Method for SO2 analysis.
H2S
Active measurements of H2S were made using an identical sulfur dioxide analyser
ML®9850 (Monitor Labs), in conjunction with a hydrogen sulfide converter H2S-1100
(Monitor Labs). The theory of operation follows that the H2S is directly converted to
SO2, in the presence of ozone, according to Equation 5.3 below. As the converter operates at only a moderate temperature (275°C), it does not convert many other reduced sulfur compounds to SO2.
2H2S + 2O3 bbb 2SO2 + 2H2O [5.3]
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5.2.3 Methods: Gas Sampling
The concentrations of both SO2 and H2S were measured at heights of 0.5 and 1 m (Nov/Dec 2003), and 1.0 and 2.5 m (Oct/Nov 2005), above the ground surface. Figure 5.3 shows the setup of the gas sampling and meteorological equipment at both sampling times. The gas sampling points and micrometeorological equipment were located on separate structures both times, excluding the rain gauge during the Oct/Nov 2005 period, when it was located on top of the sampling points scaffolding.
Air was drawn from the two different intake heights at 0.5 L/min through PTFE tubing to the gas analysers located downwind of the intake points. The two-height measurements of SO2 and H2S allowed the calculation of the flux of the two gases by vertical gradient micrometeorological techniques. The fetch of both sample locations, extending towards the predominant wind direction (SE), was in excess of the 100 and 250 m (for the two sample periods) required to satisfy a 100:1 fetch-height ratio. Additionally, the fallowed cane field presented an excellent homogeneous surface for gas exchange.
The sampling sequence of the analysers was interchanged every 15 mins using a 2-way solenoid valve controlled by a data logger. This switching of the air intakes further ensured the accuracy of absolute measurements. The construction of the masts and scaffolding attempted to minimise the wake effects generated by such structures, which can be substantial, especially regarding momentum flux (Kaimal 1969).
Micrometeorological measurements varied between the two projects. However, in order to calculate the transfer coefficient (as shown in Section 5.2.4 below), fluxes of heat, water vapour and momentum, as calculated from differences in temperature, humidity and wind speed between the two different heights, plus single height measurements of net radiation, soil heat flux, soil temperature, soil moisture and wind direction were made (Denmead et al. 2006). During the Nov/Dec 2003 sampling period measurements included wind direction and speed (as measured by 3-cup anemometers plus a directional wind vain), and latent heat flux (as determined by soil heat flux plates, net radiometer and temperature and humidity probes). The specific equipment used for these measurements included three-cup anemometers in conjunction with a directional
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wind vain to determine wind speed and direction. The net radiation flux density at the surface was measured using a levelled north-facing net radiometer, mounted 2 m above the ground surface. The heat flux across the soil surface was measured directly using soil heat flux plates (which were buried at 50 mm below the ground surface) in conjunction with a sonic open-path analyser. Soil temperature was determined using buried thermocouples. A sonic anemometer was also used for the combined measurements of turbulent wind fluctuations and heat flux. Precipitation was measured using an electronic rain gauge.
Every attempt to minimise flow distortions created by the positioning of the instrumentation was undertaken. However, certain levelling errors were unavoidable. See Kaimal and Finnigan (1994) for more detailed information on various sensors used for surface and in-situ micrometeorological measurements.
Gas Sampling Mast, air samples drawn at 0.5 & 1.0 m Micrometeorological Sampling Mast
Gas Analysers & Data Logger
Figure 5.3a. Setup of the active gas analysers and micrometeorological equipment during the 2003 sampling period.
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Gas Sampling Mast Micrometeorological Sampling Mast
Sampling Point 1 2.5m
Sampling Point 2 1.0m
Gas Analysers & Data Logger
Figure 5.3b. Setup of the active gas analysers and micrometeorological equipment during the 2005 sampling period.
5.2.4 Calculations
The resulting output from the fluorescence analysers is in ppbv. These concentration values can then be used to calculate the flux (F) of individual gases from the soil profile using the following general equation;
F (ng m-2 s-1) = hDc [5.4]
Where, h is the transfer coefficient (ms-1) across the layer of air between the two mast heights (z1 & z2). Dc represents the difference in concentration as measured by the analysers at the different mast heights, i.e. z2 - z1. The concentration values, however, need to be converted from ppb to ng m-3 prior to this calculation using the Ideal Gas Law; resulting in Equation 5.4 becoming;
F = hDc × 1000 × 0.0409 × mw(gas) [5.5]
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Where, mw is the molecular weight of the gas analysed, and 0.0409 the constant generated from the Ideal Gas Law under standard conditions (25ºC & 1 atm.). The transfer coefficient was calculated from the following equation;
h = ' u* / [ ln (z2 - d) - 1z2 ] – [ ln (z1 - d) - 1z1 ] [5.6]
Where, ' is von Kármán’s constant (taken as 0.41); u* is the friction velocity, a measure of the atmospheric turbulence (measured by the sonic anemometer); d is the zero-plane displacement (0.27); and 1 are adjustments for atmospheric stability at the different mast heights, calculated using average atmospheric pressures, air densities and the sensible heat flux (using sonic temperature values).
This is only a rudimentary explanation of the micrometeorological methods undertaken. Further explanations on the mathematic derivations of these equations and supplementary information can be found in the following references amongst many others (Arya 1988; Denmead 1983; Denmead 1994; Fowler and Duyzer 1989; Kaimal and Finnigan 1994; Sutton 1953; Wesely et al. 1989).
5.3 RESULTS 1 – NOV/DEC 2003
5.3.1 Soil Description
The field description of the soil profile taken adjacent to the gaseous emission sampling equipment is illustrated in Figure 5.4. It showed a brown to black organic topsoil (pH 4.9 to 3.8) overlying an acidic brown peat layer (pH < 3.5). Below this was the oxidised acid sulfate soil layer, followed by the transition layer which experienced both oxidising and reducing conditions depending on the depth of the water table. Within the transition layer there was a rapid increase in the pH from < 3.5 to > 5.3, where it adjoined the unoxidised clay-gel beneath it. The clay gel, which extends beyond the depth of sampling, had a pH that slowly increased from 5.5 to near-neutral > 6.3.
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Figure 5.4. Soil pH and boundary layers at the Nov/Dec 2003 gas analysis location within the Blacks Drain study site.
5.3.2 SO2 and H2S Concentrations
The concentrations of SO2 and H2S for the entire 16-day sampling period are shown in Figures 5.5a & b. Owing to the amount of data within this series it is difficult to discern any patterns that may exist. Indeed, some of the major variations within the sampling period are in fact due to equipment and power malfunctions, and some are partly rain- induced (e.g. Fig 5.5b on the 7th Dec). If individual days are examined in more detail though, some trends in the data become observable.
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Figure 5.5a. SO2 concentrations across the entire 2003 sampling period at 0.5 m above the ground surface.
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Figure 5.5b. H2S concentrations across the entire 2003 sampling period at 0.5 m above the ground surface.
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Of particular interest within this data set is a 32 hr period from the 8th to 10th December, where the H2S concentration reaches a maximum. Figure 5.6 shows the fluctuations in measured SO2 and H2S concentrations at 30 min intervals. During this 32 hr period, a decisive diurnal pattern is evident, with a marked increase in H2S and decrease in SO2 during the night, and the opposite trend observed during the day. This diurnal tendency in H2S emissions has been observed in numerous other studies (Cooper et al. 1987b; Jorgensen and Okholm-Hansen 1985; Servant and Delapart 1982; Sze and Ko 1980), and is confirmed by the correlation coefficient (r2 = 0.80), shown in Figure 5.7. However, this trend is not as pronounced in the remainder of the data. This is evidenced by the fact that the day time (07:30 to 20:30) concentration of SO2 is an average of
8.72% higher than the night time value, and conversely H2S is only 0.80% greater during the night time compared to day time values.
Figure 5.6. SO2 and H2S concentrations across a 32 hr period during the 2003 sampling period. Note; the shading represents night time hours, and the dotted lines the detection limits of the active analysers.
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Figure 5.7. Correlation between concentrations of SO2 & H2S during the 32 hr period in 2003. Note; only concentrations above the detection limits of the equipment were used.
5.3.3 SO2 and H2S Fluxes
The fluxes of the two gases as determined by the flux-gradient method are shown in Figures 5.8 a & b. Once again, it is quite difficult to distinguish overall trends within this dataset.
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Figure 5.8a. SO2 flux across the entire 2003 sampling period.
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Figure 5.8b. H2S flux across the entire 2003 sampling period.
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As a way of summarising the data in the aforementioned figures, the measurements for individual days (00:00 - 23:30) were averaged, the results of which are shown in Figure th 5.9. All of the individual days (except the 7 ) exhibit a greater average flux of SO2 as opposed to H2S, a point obscured within the complete dataset.
Figure 5.9. Mean daily flux values for both SO2 and H2S across the entire Nov/Dec 2003 sampling period.
Returning to the half-hour average flux measurements it is worth looking at the same 32-hr period identified in Figure 5.6. The vertical fluxes (as opposed to concentrations) of the two gases over this time period are shown in Figure 5.10. This diagram shows the emission of SO2 occurs primarily during the daytime period, which is consistent with the measured concentration (Figure 5.6, top), with it peaking at over 40 ng m-2 s-1.
Conversely for H2S, the flux measurements (Figure 5.10) suggest a continual, albeit inconsistent, emission across the sample period, with peaks occurring both during the day and night (~ 25 ng m-2 s-1).
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Figure 5.10. SO2 and H2S fluxes across a 32 hour period during the Nov/Dec 2003 sampling period. Note; the shading represents night time hours, and the dotted lines are the zero reference points.
Although this trend is not as conspicuous throughout the remainder of the data, the mean flux values between day and night were statistically compared. The day time hours were taken as between 07:00 to 19:00, and night time between 19:30 and 06:30.
As can be seen in Table 5.1, the mean flux of SO2 was over 16-fold greater during the day time hours compared to the mean night time flux. This difference was confirmed as statistically significant using the Mann-Whitney U Test ( = 0.01, p ~ 0.0001, n = 304).
Conversely, whilst the mean H2S flux was (marginally) greater during the night time, this proved not statistically significant either = 0.01, or at = 0.05 (p ~ 0.089, n = 298).
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Table 5.1. Diurnal flux averages for SO2 and H2S for the 2003 sample period. All values in ng m-2 s-1. Errors represented as standard error of the mean.
SO2 H2S
Day Time Mean 36.66 ± 2.78 1.67 ± 3.37
Night Time Mean 2.22 ± 1.67 2.13 ± 2.33
5.3.4 Micrometeorological Interactions
As the micrometeorological measurements made for the 2003 sampling period were not as comprehensive as the 2005 period, not as much detail will be pursued into the relationships between these and the gas measurements. The large volume of data points generated by automated gas sampling and associated climatic measurements, means that some important relationships between certain variables may be lost. As a result, some statistical analysis is required to help decipher these relationships.
As a first step, a correlation matrix between the climatic parameters and the two gases (concentration) was generated, as shown in Table 5.2. Spearman’s rank correlations were used to determine the relatedness between climatic variables and the SO2 and H2S measurements. A non-parametric statistical test was chosen because the vast majority of the data did not display a normal distribution. All data were tested for normality using Shapiro-Wilk, Lilliefors and Jarque-Bera tests; with homoscedasticity checked using Bartlett’s specificity test. Although the parametric equivalent (Pearson’s correlation coefficient) is sometimes used for climatic data that does not display a typical normal distribution, the use of Spearman’s coefficient is commonly used in the statistical examination of non-parametric gas emission and climatic parameter comparisons, for example Holtgrieve et al. (2006), Maljanen et al. (2006) and Pinto et al. (2002). Therefore, in relation to the gas and climatic measurements, any further mention of correlation will refer to Spearman’s correlation.
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Table 5.2. Correlation coefficients for SO2 and H2S concentration using Spearman’s Correlation. Values in bold are significant at = 0.01 and values with ** indicating a significance also at = 0.05.
Variable SO2 Conc. H2S Conc.
Latent Heat 0.137** 0.156
Instantaneous Precipitation -0.026 0.097
Cumulative Precipitation -0.170 0.059 Air Temperature 0.175 0.337 Dew Temperature 0.229 0.394 Relative Humidity 0.061 0.146** Av. Barometric Pressure -0.144** -0.170 Wind Direction -0.098 -0.195
As Table 5.2 shows, there are significant correlations between the gas concentrations and certain climatic variables. It should be emphasised at this point, however, that correlation does not imply causality, but it does give an indication of the climatic parameters influencing the gas emissions. In order to explain the covariance within the data, principal components analysis (PCA) was used, which linearly simplifies the data by applying two arbitrarily chosen components (set as axes) to account for the variance- covariance in the dataset. As is shown in Figure 5.11, the amount of variance captured by the PCA of all variables is relatively low at ~ 50 % for both SO2 and H2S. It does, however; highlight the positive correlation between SO2 concentration and air temperature, latent heat and dew temperature, and the negative correlation with barometric pressure. H2S displays similar relationships to the climatic variables.
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Variables (axes F1 and F2: 50.87 %) Variables (axes F1 and F2: 50.16 %)
1 1 Air Temp Air Temp Latent Heat Latent Heat 0.75 0.75
SO2 Conc
0.5 0.5 H 2 S Conc Dew Temp Dew Temp 0.25 0.25
0 0 Bar. Press Bar. Press Rel. F2 (23.33%) Rel. Humidity (22.33%) F2 -0.25 Wind Dir -0.25 Instan. Precip Humidity Instan. Precip Wind Dir Cumul. Precip -0.5 Cumul. Precip -0.5
-0.75 -0.75
-1 -1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 F1 (27.55 %) F1 (27.83 %)
Figure 5.11. PCA of SO2 (left) and H2S (right) concentrations and measured climatic variables during the 2003 sampling period.
The flux values for both gases were also compared in the same way as described above. The correlation matrix (Table 5.3) shows that overall there were stronger correlations between SO2 flux and the climatic variables as opposed to SO2 concentration and the climatic variables (Table 5.2). Conversely, the significant relationships with H2S concentration (Table 5.2) are no longer observed when comparing H2S flux (Table 5.3), with no significant correlations at either 1 or 5 %.
Table 5.3. Correlation coefficients for SO2 and H2S flux using Spearman’s Correlation. Values in bold are significant at = 0.01 and values with ** indicating a significance also at = 0.05.
Variable SO2 Flux H2S Flux
Latent Heat 0.485 0.018
Instantaneous Precipitation -0.047 0.121
Cumulative Precipitation -0.169 0.006 Air Temperature 0.530 0.042 Dew Temperature 0.045 -0.007 Relative Humidity -0.297 -0.027 Av. Barometric Pressure -0.007 0.085 Wind Direction -0.089 0.030
In order to identify covariance in the dataset, a PCA for both gas fluxes were constructed (Figure 5.12). Again these manipulations only account for just over half the
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variance in the dataset. Nevertheless, they confirm the strong correlation of SO2 flux and latent heat and air temperature, and the lack of correlation between H2S and any of the other parameters.
Variables (axes F1 and F2: 52.42 %) Variables (axes F1 and F2: 50.59 %)
1 1
Air Temp 0.75 Dew Temp 0.75 Air Temp Latent Heat Latent Heat Dew Temp 0.5 0.5 SO 2 Flux
0.25 Rel. Humidity 0.25
0 0 Instan. Precip Rel. Humidity F2 (23.81%) Cumul. Precip (18.03%) F2 Instan. -0.25 -0.25 Cumul. Precip Wind Dir Precip H S Flux Bar. Press 2 -0.5 -0.5 Wind Dir Bar. Press -0.75 -0.75
-1 -1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 F1 (28.61 %) F1 (32.56 %)
Figure 5.12. PCA of SO2 (left) and H2S (right) fluxes and measured climatic variables during the 2003 sampling period.
Since the micrometeorological method of flux determination is dependent on the air sample of interest flowing over a designated area of ground surface, it is logical to distinguish the data based on the wind direction. Therefore, if only the measurements made whilst the prevailing wind flowed over the selected fetch for the micrometeorological equipment, then more informed correlations with the gas fluxes would be anticipated. The correlation matrix for this data is shown in Table 5.4. As suggested, stronger positive correlations are observed with SO2 flux and air temperature and latent heat. Relative humidity is also moderately negatively correlated with SO2 flux. Although weaker correlations are observed with the H2S flux, some still proved significant at 5%.
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Table 5.4. Correlation coefficients for SO2 and H2S flux, restricted by ideal wind direction (80-160°), using Spearman’s Correlation. Values in bold are significant at = 0.01 and values with ** indicating a significance also at = 0.05.
Variable SO2 Flux H2S Flux
Latent Heat 0.613 0.370**
Instantaneous Precipitation 0.027 0.310
Cumulative Precipitation -0.108 -0.015 Air Temperature 0.633 0.222 Dew Temperature -0.101 0.108 Relative Humidity -0.428 -0.025 Av. Barometric Pressure -0.055 -0.397** Wind Direction 0.022 -0.125
These stronger relationships are also confirmed in the PCA (Figure 5.13). The SO2 PCA (Figure 5.13, left) shows the (not unexpected) relationship between wind direction and other climatic variables (relative humidity and precipitation). This could indicate that the observed correlation between SO2 and relative humidity is only due to the manipulation of the data based on wind direction rather than an inter-relationship between the two.
Variables (axes F1 and F2: 63.26 %) Variables (axes F1 and F2: 63.44 %)
1 1 Latent Latent Heat 0.75 Heat 0.75 Air Temp Dew Temp Air Temp 0.5 0.5 H 2 S Flux Dew Temp
SO2 Flux 0.25 0.25 Rel. Humidity Wind Dir Instan. Precip 0 0 Instan. Cumul. Rel. Humidity F2 (24.89%) F2
F2 (23.61%) F2 Precip -0.25 Precip -0.25 Cumul. Precip
-0.5 -0.5 Wind Dir
-0.75 Bar. Press -0.75 Bar. Press
-1 -1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 F1 (39.65 %) F1 (38.56 %)
Figure 5.13. PCA of SO2 (left) and H2S (right) fluxes and measured climatic variables, based on wind direction (80-160°) during the 2003 sampling period.
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As suggested above though, interpreting too much into the relationships between the gas measurements and climatic variables is fraught with uncertainties in this instance because of the reservations regarding much of the climatic data. As more comprehensive measurements were made during the 2005 sample period, a more detailed analysis is subsequently made.
5.4 RESULTS 2 – OCT/NOV 2005
5.4.1 Soil Description
The pH, redox and soil boundary layers are shown in Figure 5.14. The 2005 soil description shows a similar profile to that recorded during the 2003 sampling program, with the only major difference being the absence of the black organic topsoil layer. The redox profile (not measured in 2003) illustrates that the unoxidised material is quite reducing, being < 100 mV (standard hydrogen electrode).
Figure 5.14. Redox, pH and soil boundary layers for a profile taken at the study site during the 2005 gas sampling period.
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5.4.2 SO2 and H2S Concentrations
The concentrations of both SO2 and H2S over the entire sampling period are shown together in Figure 5.15. Again, specific trends within the data set are hard to discern when the data is shown in its entirety. However, it is clear that the major peaks in SO2 occur during the middle of the day, whilst H2S (although more irregular) occur during the night time.
When comparing the values between the 2003 and 2005 sample periods, it is immediately noticeable that concentrations of SO2 are greater than H2S in the 2005 sample set, which is the opposite to 2003. Furthermore, looking at the concentration maxima across the two periods, although the SO2 levels are approximately equal between 2003 and 2005 (~ < 3 ppb), the H2S concentrations are considerably reduced within the 2005 period (~ < 1 ppb).
5.4.3 SO2 and H2S Fluxes
As can be seen in Figure 5.16, the fluxes of SO2 are substantially greater than H2S. This is the case for both positive and negative peak fluxes (SO2 flux peaks at -2 -1 -2 -1 314.53 ng m s and H2S at 101.99 ng m s ; raw data – not shown) , as well as in terms of the overall average, which over the entire sampling period is nearly 5-times greater for SO2. This disparity can be seen in Figure 5.17 which shows the average daily flux of the two gases.
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Figure 5.15. SO2 and H2S concentrations across the entire 2005 sampling period.
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Figure 5.16. SO2 and H2S fluxes across the entire 2005 sampling period.
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Figure 5.17. Averaged daily fluxes of SO2 and H2S throughout the 2005 sampling period.
As illustrated with the 2003 data, the mean flux values between day and night also show interesting results. The mean fluxes of the two gases for the day time (07:00 - 19:00) and that for the night time (19:30 - 06:30) are shown in Table 5.5. For both SO2 and
H2S, the day time mean fluxes were greater than that measured during the night time hours, with the fold increase being 9.3-fold for SO2, and 3.3-fold for H2S.
Table 5.5. Diurnal flux averages for SO2 and H2S for the 2005 sample period. All values in ng m-2 s-1. Errors represented as standard error of the mean.
SO2 H2S
Day Time Mean 10.20 ± 3.83 1.79 ± 1.11
Night Time Mean 1.10 ± 1.34 0.54 ± 0.41
Analysis of the data using the Mann-Whitney U Test ( = 0.01) confirmed that the SO2 daily average was significantly different to the nightly average (p ~ 0.001, n = 358).
However, the perceived difference in the H2S proved not statistically significant at either = 0.01, or at = 0.05 (p ~ 0.085, n = 358).
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5.4.4 Micrometeorological Interactions
The statistical analysis of the 2003 results was able to give some indication of the possible interactions between gas concentration / flux and climatic variables. Although these were the focus of the 2005 sample period, other previously-unmeasured parameters were also considered. Trends in the raw data were considered in the first instance.
Atmospheric Pressure As was demonstrated in the results from the 2003 sample period, there appeared a potential link between atmospheric pressure and the flux of SO2, and particularly with
H2S. The atmospheric pressure along with the flux of SO2 across the entire sample period is shown at the top of Figure 5.18. An arbitrary flux line was chosen to separate the ten largest flux values across the sample period. These peaks were then enlarged along the bottom of Figure 5.18. As can be seen, of the ten principal flux peaks during the sampling period, it would appear that all of them (except number 6) are associated with decreases in atmospheric pressure.
The H2S fluxes are also compared in a similar manner to the atmospheric pressure in Figure 5.19. As can be seen in the enlarged sections of the graph at the bottom of Figure
5.19, of the eight major peaks in H2S flux, all would appear to be associated with a decrease in atmospheric pressure.
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Figure 5.18. Comparison between SO2 flux and atmospheric pressure during the 2005 sampling period. Enlargements of these ten peaks are shown beneath. The flux lines are represented as raw, 3-run smoothed and 5-run smoothed data. Note; the dotted line indicates an arbitrarily chosen flux to segregate the ten largest SO2 flux peaks.
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Figure 5.19. Comparison between H2S flux and atmospheric pressure during the 2005 sampling period. Enlargements of these eight peaks are shown beneath. The flux lines are represented as raw, 3-run smoothed and 5-run smoothed data. Note; the dotted line indicates an arbitrarily chosen flux to segregate the eight largest H2S flux peaks.
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This would then suggest that there is an inverse link between atmospheric pressure and the major fluxes of the two measured gases. Whether this link is causative or simply systematic of other variables (climatic or other) cannot be established by examining trends in the two variables in isolation.
Rainfall
A comparison of the SO2 and H2S fluxes with the measured rainfall across the sampling period is shown at the top of Figure 5.20. The enlarged sections below this show the five rain periods, varying from falls of 5.6 mm (across 6 hrs) to 48.8 mm (across 2.5 hrs). One major point to observe from the graphs is the reduction in gas fluxes almost immediately following the precipitation events. It would appear that the flux of
SO2 is influenced more so than H2S, with periods of ‘near-zero’ flux extending on average from between 8 to 14 hrs after the rainfall event. These periods are highlighted by brackets in Figure 5.20. Again, it needs to be acknowledged that, although the rainfall events show a degree of correlation to the flux of the two gases, they do not indicate any causative effect.
Air Temperature The changes in gas fluxes were also compared to the air temperature. Figure 5.21 (top) shows the comparison with SO2 across the entire sample period, and the ten major peak enlargements underneath. The majority of the peaks in SO2 flux are observed at the same time when temperature is also at a maximum, although this trend is not as well defined as observed with the atmospheric pressure data.
H2S fluxes were also compared to air temperature (Figure 5.22) over the entire period (top) with enlargements of the eight major fluxes (beneath). Again, although there appears to be a recurring trend between large fluxes of H2S and air temperature maxima, it is not as pronounced as that seen with SO2 and air temperature, or atmospheric pressure.
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Rainfall Events 1 & 2
Rainfall Events 3 & 4
?
Rainfall Event 5
Figure 5.20. Comparison between SO2 / H2S flux and rainfall during the 2005 sampling period. The entire 2005 sample period shown at top. Enlargements of the five rainfall events in chronological order (beneath).
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Figure 5.21. Comparison between SO2 flux and air temperature during the 2005 sampling period. Enlargements of these ten peaks are shown beneath. The flux lines are represented as raw, 3-run smoothed and 5-run smoothed data. Note; the dotted line indicates an arbitrarily chosen flux to segregate the ten largest SO2 flux peaks.
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Figure 5.22. Comparison between H2S flux and air temperature during the 2005 sampling period. Enlargements of these eight peaks are shown beneath. The flux lines are represented as raw, 3-run smoothed and 5-run smoothed data. Note; the dotted line indicates an arbitrarily chosen flux to segregate the eight largest H2S flux peaks.
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5.4.5 Statistical Analysis
As with the 2003 dataset, there is a necessity to statistically analyse these observed trends. Again, the first step is to look at the correlation matrix to see possible influencing factors on the gas emission. Both the correlation matrices for gas concentration and flux, and climatic variables are shown in Table 5.6. Immediately obvious from Table 5.6, are the number of significant correlations between the gas concentrations and the climatic variables. Conversely, with regards to the flux correlations, SO2 flux is significant only with barometric pressure ( = 0.01).
Table 5.6. Correlation coefficients for SO2 / H2S concentration and flux using Spearman’s Correlation. Values in bold are significant at = 0.01 and values with ** indicating a significance also at = 0.05.
Variables SO2 Conc. H2S Conc. SO2 Flux H2S Flux
Av. Barometric Pressure 0.012 -0.415 0.124 0.020 Av. Air Temperature 0.316 -0.415 0.116** 0.046 Precipitation 0.084 0.031 0.045 0.096** Av. Net Radiation 0.067 -0.589 0.107** 0.084 Av. Water Vapour 0.101 0.205 0.002 -0.079 Wind Direction -0.113** -0.212 0.111** 0.032 Soil Temp @ 5cm 0.387 -0.384 0.101** -0.012 Soil Temp @ 10cm 0.334 -0.166 0.088 -0.094** Soil Temp @ 20cm 0.230 -0.138 0.064 -0.061 Soil Moisture @ 5cm -0.178 0.124 -0.061 0.019 Soil Moisture @ 10cm -0.296 0.211 -0.085 0.072 Soil Moisture @ 20cm -0.174 0.249 -0.064 0.006 Latent Heat 0.239 -0.459 0.096** 0.068
The data within the correlation matrix was then simplified using PCA, with the PCA vector diagrams for SO2 and H2S concentrations shown in Figure 5.23. The PCA would suggest two actions at work in this system graphically represented by factor 1 along the x-axis. In the case of the SO2 concentration (Figure 5.23, left); to the positive side are the temperatures (soil and air), whilst the negative side is dominated by the soil moisture and precipitation. The underlying structure of the H2S concentration PCA
(Figure 5.23, right) is almost inversely identical to the SO2 PCA. To simplify, if both concentrations were plotted on the same graph, they would lie either side of the y-axis.
This positive correlation between SO2 and the temperatures, and negative towards to the
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moisture (and inverse for H2S), is confirmed by the above shown correlation matrix and also by the corresponding high factor loadings (> 0.9; not shown).
Variables (axes F1 and F2: 57.07 %) Variables (axes F1 and F2: 54.88 %)
1 1
Soil M oist. Soil Temp Av. Bar. Pres. Soil M oist. 0.75 20cm 0.75 20cm 5cm Latent Heat Av. Air Temp. 0.5 0.5 Soil Temp Av. Net 5cm 0.25 Radiation 0.25 Soil Temp Wind Precipitation 10 c m So il M o ist. SO 2 Conc Direction H 2 S Conc 10 c m 0 0 Wind Av Water Vap. Direction
F2 (23.09%) F2 Soil M oist.
F2 (23.09%) F2 Soil Temp -0.25 -0.25 Precipitation 5cm Av Water Vap. 10 c m Soil Temp 10 c m Av. Net -0.5 -0.5 Radiation So il M o ist. Soil Temp 20cm Soil M oist. 20cm Latent Heat Av. Air Temp. -0.75 5cm -0.75 Av. Bar. Pres.
-1 -1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 -1 -0.75 -0.5 -0.25 0 0.25 0.5 0.75 1 F1 (33.97 %) F1 (31.79 %)
Figure 5.23. PCA of SO2 (left) and H2S (right) concentrations and measured climatic variables during the 2005 sampling period.
The PCA for the fluxes of SO2 and H2S showed a similar distribution in the climatic variables, and in terms of where the vector lay for each gas flux in relation to these (not shown). The first and second principle components in this case explained a total of
54.75% and 53.94% of the variation in the SO2 and H2S flux dataset respectively.
In order to clarify the covariance within the dataset a little further, agglomerative hierarchical clustering (AHC) of the datasets were undertaken, using Spearman’s dissimilatory algorithm. Clustering is a technique that can be used to help partition the different variables into smaller groups based on the similarity of specified characteristics (Hair et al. 2006). With respect to AHC, each observation starts off as its own cluster before being grouped with next most similar cluster, thereby continually reducing in number to a singular cluster (Hair et al. 2006), and is used primarily as an exploratory data analysis tool in atmospheric sciences (Wilks 1995).
The results of this further simplification for SO2 and H2S concentrations are shown as dendrograms in Figure 5.24. As can be seen in Figure 5.24 (top), AHC confirms that the
SO2 concentration is primarily related to the climatic variables soil temperature (5 cm),
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air temperature, latent heat and net radiation (red lines). It is also related, but to a lesser degree, to the other soil temperatures and water vapour (purple lines), and least related to the moisture-related (blue lined) variables. The inverse trend in correlation between
SO2 and H2S concentrations continues, with the results of the AHC for the H2S concentration (Figure 5.24, bottom) showing that it is most closely related to soil moisture (10 and 20 cm) (blue lines), followed by precipitation and soil moisture at
5 cm (purple lines); or the variables to which SO2 concentration is least related.
Figure 5.24. AHC of SO2 (top) and H2S (bottom) concentrations and measured climatic variables during the 2005 sampling period. Grouped variables are dependent on relation with gas concentration below 0.4 dissimilarity.
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The results of the AHC for SO2 and H2S flux samples (Figure 5.25) do not show as close an association with the climatic variables, but do confirm the overall trend shown throughout the statistical analysis of the data.
Figure 5.25. AHC of SO2 (top) and H2S (bottom) fluxes and measured climatic variables during the 2005 sampling period. Grouped variables are dependent on relation with gas concentration below 0.4 dissimilarity.
Interestingly, when looking at the samples that corresponded to those in which the wind direction was flowing over the ideal fetch, the AHC shows a reduced correlation with the variables identified above (Figure 5.26). Figure 5.26 (top) shows that the SO2 flux is most closely related to soil moisture (5 cm), soil temperature (20 cm) and barometric pressure. The wind-directed H2S flux (Figure 5.26, bottom) shows that it is not closely
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 116
related to any of the climatic variables, with the closest being wind direction, precipitation and soil moisture (10 and 20 cm).
Figure 5.26. AHC of SO2 (top) and H2S (bottom) fluxes, flowing over the ideal fetch, and measured climatic variables during the 2005 sampling period. Grouped variables are dependent on relation with gas concentration below 0.4 dissimilarity.
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5.5 DISCUSSIONS & PRELIMINARY CONCLUSIONS
5.5.1 H2S Measurements: Formation and Consumption Processes
Contrary to the comments from Hansen et al. (1978) and Jaeschke et al. (1978) that for
H2S to be released, the sulfide zone must extend to the very profile surface, these ASS have been shown to emit H2S where the zone of sulfate reduction is almost 1 m below the ground surface. Dissolved H2S reacts rapidly with any nearby dissolved iron and other metals to form various metal sulfides, in accordance with favorable thermodynamics and kinetics. It has been shown in aquatic sediments, however, that this only accounts for small amounts (~ 10 %) of the sulfate reduced in these systems (Elsgaard and Jorgensen 1992; Lin and Morse 1991; Swider and Mackin 1989). In addition to the formation of metal sulfides, the diffusive movement of H2S to the oxidising zone and to the atmosphere would place it in contact with oxygen and iron oxides. These strong oxidising agents react much faster with sulfide to form elemental sulfur and sulfate (Canfield et al. 1992). So much so, that iron (oxy)hydroxides are being trialed as detoxification agents in minimising the adverse impacts of dissolved sulfide in aquaculture pond, e.g. Lahav et al. (2004). H2S can also be retained within the sediment through the formation of diagenic organic sulfides (Bruchert 1998). This, along with the chemical oxidation processes described above are exemplified by
Kristensen et al. (2003), who found within a sandy Danish sediment, that H2S was prevented from reaching the surface by the total metal pool during periods of intense anoxia.
Although anaerobic sediments commonly do not completely retain dissolved sulfides by precipitation and / or burial (Chanton et al. 1987), they are readily consumed by bacterial action. Indeed, H2S provides an excellent substrate for microbial oxidation for specific sets of bacteria to obtain energy (Andreae 1990). The combined chemical and bacterial action often results in up to 90 % of dissolved sulfide being recycled within sediments (Jorgensen 1977b; Thamdrup et al. 1994b). In marine sediments, Jorgensen et al. (1990) have reported that between 68 to 96 % of H2S present in porewaters (North / Baltic Seas) is re-oxidised within the sediment.
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Nonetheless, the H2S flux measurements are real and others have shown gaseous release, including from seawater, which Elliott and Rowland (1990) suggest is a dependence on S saturation levels as well as the availability and binding properties of organic matter and metals. Exactly how H2S can be emitted to the atmosphere without being oxidised by the surface layers, through which it must first pass in the ASS profile, is something that still needs to be further examined.
Formation The bacterial decomposition of organic matter has been shown to be vital during the early diagenesis of pyrite, with hydrogen sulfide reacting with detrital iron minerals.
The formation of H2S can therefore be accounted for by the dissimilatory reduction of sulfate in the presence of decomposing organic matter (see Equation 5.7). The predominant bacterial assemblage associated with this process is the genus Desulfovibrio, which can operate at a pH from between 4.2 to 9.9, although preferring a pH close to neutral (D. desulfuricans), see Langmuir (1997) and references cited therein.
2- + SO4 + 2CH2O(s,aq) + 2H bbb H2S + 2CO2(aq,g) + 2H2O [5.7]
The dissolved sulfide primarily reacts with forms of reduced iron to eventually produce insoluble iron monosulfides, and potentially pyrite (Berner 1970). However, sulfate reduction can also result in the occurrence of large concentrations of ferrous iron and
H2S without their conversion to iron monosulfides, at relatively low pH (van Breemen 1988b). As not all of the iron in the sediment is readily reducible (to Fe2+), nor is it reactive towards H2S, therefore the concentrations of H2S can build up and diffuse away from the source of organic matter degradation; see Figure 5.27, from Berner (1980). This is most likely to occur in the transition layer, which in the sampled profile is > 1.2 m from the ground surface (Figure 5.4 & 5.14). It is within this layer that the pH increases above 5.5, a fundamental requirement for the growth of Desulfovibrio desulfuricans. If the reduction continued, H2S would be able to diffuse upwards towards the sediment-air interface along concentration gradients, and be emitted to the atmosphere. If the H2S reacts with excess oxygen, then a large concentrations of 0 0 elemental sulfur (S 8) would be expected (Jorgensen 1983). The presence of S 8 has
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been identified within drain samples at the Blacks Drain study site, in proportions of up to 50 % of the reduced inorganic sulfur content of the sediment (Burton et al. 2006b). However, it is more likely to be associated with high concentrations of AVS, primarily 0 as an intermediate of AVS oxidation. Therefore, even though S 8 was not quantified as part of these experiments, it is anticipated that it would contribute only a minor fraction of the reduced inorganic sulfur fraction, because of the limited concentrations of Fe(III) and AVS, shown in Chapter 3. This would suggest the H2S was not impeded in its diffusive movement towards to the sediment-air interface through transformations to 0 S 8.
Potential Export?
Area of localised H2S build-up
FeS2
2- 2- SO4 H2S SO4 H2S Organic Matter Aggregation
Fe2+ Fe2+ H2S
FeOOH FeS 2 FeOOH
Figure 5.27. Formation and possible export of H2S from the decomposition of organic matter. Adapted from Berner’s (1980) depiction of the formation of pyrite concretions around a clot of organic matter.
The process of atmospheric H2S emissions is more characteristic of lake and estuarine sediments where the zone of active sulfate reduction occurs close to the air surface e.g. tidal flats such as in Jaeschke et al. (1980). Ingvorsen and Jorgensen (1982) showed that within inter-tidal sediments, less than 0.001 % of the total sulfide resulting from dissimilatory sulfate reduction escapes to the atmosphere. This small figure is due to sulfides propensity for microbial oxidation (Visscher 1996), as well as its short chemical half-life (Zhang and Millero 1993).
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Whilst the reduction of sulfate is the more likely cause of H2S present in the profiles measured, alternatives should also be considered. The underlying pyritic substrate cannot be discounted as a potential source, as it is possible for gaseous H2S to form in acidic (pH < 6.5) conditions from the dissociation of aqueous iron sulfide clusters (Equation 5.8), or similarly from the dissociation of iron sulfide complexes. These have been demonstrated in the laboratory and within an anoxic man-made sulfidic lake by Luther III et al. (2003; 1996), see Equation 5.9. These reactions only proceed though, when the reduction of Fe(III) is superior to the reduction of SO4, i.e. when the system is sulfide rather than iron limiting.
+ 2+ FeS(aq) + 2H bbb H2S + Fe [5.8] 3+ + 2+ Fe2SH + H bbb H2S + 2Fe [5.9]
H2S can also be generated from the decomposition of sulfur-containing amino acids such as methionine and cysteine (Grinenko and Ivanov 1983). One example of this is the reaction between cysteine and cysteine desulfhydrase forming H2S and ammonia (Segal and Starkey 1969), as represented by Equation 5.10;
HS-CH2-CH(NH2)-COOH + H2O bbb H2S + NH3 + CH3-CO-COOH [5.10] cysteine pyruvic acid
Another process of H2S formation that could also be responsible for our measurements 2- is the reduction of sulfur intermediates, particularly thiosulfate (S2O3 ). Thiosulfate has been shown to be a key intermediate in redox sulfur cycling (Fossing and Jorgensen 1990), as well as being a product of pyrite oxidation (Goldhaber 1983; Luther III 1987). Importantly, it can be readily reduced (Equation 5.11) or disproportionated (Equation
5.12) to H2S by sulfate reducing bacteria (Jorgensen 1990).
2- - + - S2O3 + CH3COO + H bbb 2HS + 2CO2 + H2O [5.11] 2- 2- - + S2O3 + H2O bbb SO4 + HS + H [5.12]
H2S also has been shown to be emitted from higher plants in response to the uptake of sulfate, SO2, and L- and D-cysteine (Rennenberg 1989). Although from that study, emissions of VSCs (H2S and MSH) from higher plants occurred during the daytime
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rather than during the night, the opposite of that measured at Blacks Drain. Additionally, because the paddock was fallow for the 2003 measurements, and newly growing in the 2005 study, any possible effects from this would be negligible.
Microbial Degradation
The microbial degradation of H2S also needs to be considered (somewhat briefly) as part of this discussion. There are two physiologically different groups of bacteria which are specialised in the oxidation of H2S; colourless- and phototrophic- sulfur bacteria (Jorgensen 1983). The colourless sulfur bacteria (e.g. Acidithiobacillus and Beggiatoa spp.) require oxygen to oxidise sulfide or other reduced sulfur compounds, and as such would most likely be located at the transition of aerobic / anaerobic conditions where O2 and H2S gradients converge (Kuenen 1975).
The second group, phototrophic green and purple sulfur bacteria, are anaerobic organisms specialised in the consumption of H2S for CO2 assimilation (Jorgensen 1983). However, as this is a photosynthetic process requiring light, they usually occur within the photic zone very close to the sediment-air surface (Schlegel 1974). As such, they are more relevant within aquatic systems, where the sediment-air surface is usually the lake bottom where light, but not oxygen, is able to penetrate. They would only be of importance at the very soil surface. More detailed information on the microbial cycling of sulfur can be found in the above mentioned papers as well as in the following reviews (Bruser et al. 2000; de Wit 2000; Germida et al. 1992).
5.5.2 H2S Release Mechanisms
Although many studies have quantified the release of trace gases including H2S from soil and marine systems, there is little evidence on the fundamental controlling processes; see review by Upstill-Goddard (2006). As it would appear that the majority of sulfide production occurring at Blacks Drain is doing so from depths > 1 m, the possible mechanisms leading to its release to the atmosphere need to be considered. Most soil-gas transport studies through porous media have been shown to occur through a process of molecular diffusion (or turbulent diffusion) and to a lesser extent,
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advection, through the soil porespace along a concentration gradient (Aachib et al. 2004).
An additional phenomenon that may contribute to emissions of H2S, specifically in areas such as Blacks Drain where there is > 1 m of unsaturated soil above the zone of reduction, is enhanced diffusional transport via shear planes. Shear planes are structural deformations with the soil matrix occurring primarily because of reductions in the water content and reactions with the soil’s exchange complex associated with soil ripening. See White et al. (2003) for a full explanation of the dewatering and hydraulic properties of ASS. As a result, shear planes occur within the upper, oxidised ASS. Therefore, shear planes may provide another means by which H2S (and possibly other gases formed at depth) can exit the soil profile, and bypass the mediating effects of the oxidised zone. It should be noted however, that as is the case with other disruptions in the continuity of soil profiles such as bioturbation, oxidation processes occur along their surface edges, potentially negating the effects of mass H2S transport from deeper sediments.
Bubble Ebullition As an alternative to diffusion along concentration gradients, within coastal marine and freshwater ecosystems, reduced gas exchange between the subsurface and atmosphere can also proceed by bubble ebullition (Chanton and Whiting 1995; MacIntyre et al. 1995; Mitsch and Gosselink 2000). Bubble ebullition is a one-way transport of gases from reducing sediments to the atmosphere or overlying water, principally as a way for aquatic plants to ventilate the water-logged sediments in which they are rooted (Chanton and Whiting 1995). In transporting gases directly to the atmosphere, bubble ebullition bypasses the oxidative impacts of overlying sediment, as discussed previously. As Chanton and Whiting (1995) further identify, for bubbles to form, the concentration of the gas of interest does not need to exceed the solubility of the gas in water, rather it simply needs to accumulate sufficiently to generate bubbles under hydrostatic pressure. Previous observations have shown that methane emissions via bubble ebullition routinely exceed that by molecular and turbulent diffusions, within lakes (Ostrovsky 2003) and from sub-tidal freshwater sediments, where Chanton et al. (1989) found that 50 % of the total methane emissions were due to ebullition.
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Even though ebullition is primarily associated with menthanogenesis and within saturated profiles, in rice paddies for example (Rothfuss and Conrad 1998), it still may be relevant to the release of H2S from ASS. Roden and Tuttle (1992) discuss the possibility of dissolved sulfide being flushed from the aqueous system (as H2S) by gas bubble ebullition, driven by CH4 production. In essence, the less soluble CH4 forms cohesive bubbles which force the H2S (and other dissolved gases) to the surface. It is also worthy to note that ebullition bubbles enhance the process of diffusion of other gases (and dissolved species) towards the air-surface interface, by creating air-pockets through which the bubbles exit the sediments (Chanton and Whiting 1995; Kipphut and Martens 1982).
Although ebullition can occur at any point in time, which causes problems with static sampling programmes (Mattson and Likens 1990), it is often initiated by some forcing factor such as wind speed, causing greater physical disturbance of the sediments, or shifts in hydrostatic, atmospheric pressure or tidal gradients (Carroll et al. 1986; Chanton et al. 1989; Chanton and Whiting 1995; Cooper et al. 1987a; MacIntyre et al. 1995; Martens and Valklump 1980). Therefore, if bubble ebullition is a contributing factor to the release of H2S from the Blacks Drain site, then we would expect a correlation between its emission and a triggering factor.
Previous evidence shows that the total flux of dissolved sulfide from ebullition is minimal, with Chanton et al. (1987) showing inconsequential levels of release from near-coast sediments via this process. However, non-saturated ASS, such as those at the Blacks Drain study site have been shown to emit large quantities of methane (Denmead et al. 2006). Although this may suggest a possible connection between the emissions of methane and H2S, not all sediments have the cohesiveness to support this air-pocketing (Kelley et al. 1990). It is suspected then, that ebullition at Blacks Drain is at most a minor contributor to the measurements of H2S. The fact that releases of H2S do not appear in the periods after rainfall events (during the 2005 measurements) when the soil profile would still be saturated, also agrees with this assertion. The process should not be discounted, however, as an influence within permanently saturated and reducing ASS profiles such as at the Cudgen site.
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Micro-niches
The previous reasoning’s have inferred that the formation of H2S takes place within the reducing zone of the profile and therefore needs to be transported over 1 m to the atmosphere. An alternative theory is that the H2S is actually formed at or very close to the sediment-air interface. During the night, reducing conditions may form in these near-surface sediments as a result of high metabolic activity, with respiration in the absence of photosynthesis inducing anoxia. Localised areas of sulfate reduction may be occurring in the oxic zone around reducing micro-environments of organic matter or areas of bioturbation, in a similar manner to that occurring in the reducing zone (Aller 1982; Devai and DeLaune 1995; Francois 1987; Froelich et al. 1979). As oxygen is consumed by aerobic chemoheterotrophic bacteria the conditions become locally 2- anaerobic, allowing for the use of SO4 as the terminal electron acceptor. The explanation of micro-niches have been used by Jorgensen (1977a) to account for the action of bacterial reduction of 35S in the oxidised surface layers of marine sediments, and also by Canfield (1989) to potentially account for the presence of free Fe(II) in sulfide-dominated marine sediments.
This hypothesis is partially supported by some of the work being undertaken at the Blacks Drain study site by Dürr (in prep.), and colleagues. Figure 5.28 shows the extractable DNA, isolated and then measured by realtime PCR, with the data and methods contained within Dürr (in prep.). If it is assumed that there is a certain amount of DNA per microorganism, then the following measure gives a rough indication of the microbial activity. However, it should be noted that it is entirely possible that much of the DNA may not be related to microbial activity (especially S microbial activity) and / or the DNA may be non-functional. Further information, such as functional gene quantification, would be necessary for the assessment of the active genes related to key biological processes, such as sulfate reduction. Furthermore, as Blacks Drain is an agricultural soil fertilised with urea, its microbial biomass is expected to be enhanced within the topsoil, owing primarily to N-related processes. Nevertheless, if the above concerns are taken into consideration, and Figure 5.28 shows an approximation of the microbial activity, then we can see that by far the largest activity occurs within the upper 0.5 m of the Blacks Drain profile. This goes a small way to support the micro- niche hypothesis, but clearly further work is necessary before drawing any concrete
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conclusions. Interestingly, the Cudgen site profile provides a contrasting measurement (as per the same technique) as an undisturbed soil. The values of extractable DNA at depths below 0.5 m correspond to the values measured at Blacks Drain supporting the chemical evidence in Chapter 3 of the relative similarities between the two sites’ subsoils.
Figure 5.28. Extractable DNA as a measure of soil biomass at the Blacks Drain and Cudgen study sites. All data and methods attributable to Mira Dürr’s PhD thesis (in prep). Note; dashed lines represent soils layers for the Blacks Drain site, and not the Cudgen Lake site.
Another aspect of near-surface sulfate reduction, not necessarily occurring within micro-niches, is also worth considering. SRB have generally been thought of as strict anaerobes. However, contrary to this belief, although not ideal for their growth, several studies have measured active bacterial sulfate reduction in the presence of limited oxygen (and also nitrate) concentrations (< 20 % air) in the natural environments of microbial mats (Canfield and Desmarais 1991; Jorgensen 1994; Teske et al. 1998; Visscher et al. 1992), aerobic and periodically aerobic sediments (Bak and Pfennig 1991; Jorgensen and Bak 1991), as well as in laboratory cultures (Dilling and Cypionka 1990; Eschemann et al. 1999; Marschall et al. 1993). This process is thought to occur as a response by the SRB, including D. desulfuricans and Desulfobacterium spp., to reduce oxygen concentrations as a way of establishing anoxic conditions for continued
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anaerobic growth (Eschemann et al. 1999). As a consequence of this, the aerobic respiration rates (as opposed to growth rates) are substantially higher in the aerobic zone when compared to their anaerobic functioning, in an attempt to rapidly consume nearby
O2.
Whilst this aerobic SRB respiration may not contribute greatly to H2S production, it may be important in the creation of micro-niches, where areas of localised anoxia are established by the O2-tolerant SRB, and H2S production is more prevalent. Indeed,
Teske et al. (1998) suggests that more O2-sensitive SRB shield and protect themselves from full oxygen exposure by potentially forming co-cultures or particle associations and aggregations. If this is actually the case, then it is not necessary for a rainfall (or such other) event to induce anoxic conditions in the profile before near-surface colonization of SRB, and potentially H2S emissions, occur. This then is a plausible explanation for the near-continuous H2S flux observed during the sampling periods.
Regardless of the exact nature of its formation, it is more conceivable that H2S is being emitted to the atmosphere if the origin of sulfide production is much closer to the surface.
Vegetative / Animal Transport Another factor that needs to be considered as part of this hypothesis is the role of any overlying vegetation, as Chanton and Whiting (1995) identify that most vascular plants also have a role in the emission of sulfur gases, ammonia and hydrocarbons via gas pressurised flow-through or diffusive transport processes. A study by Lee (2003), found that certain colonising species of estuarine cordgrass (S. anglica) were able to transport
H2S between the rhizosphere and atmosphere, and vice-versa. Similarly, emergent aquatic plants (Eleocharis sphacelata) in freshwater wetlands (S.E. Australia) have been shown to release CH4 from sediments in quantities of up to 15 times when compared to that released by bubble ebullition (Sorrell and Boon 1994). This process is complicated by the inverse transport of oxygen to the sediment (controlled by plant photosynthesis) which in turn effects the redox and thus, the microbial processes occurring at the rhizosphere (Brix et al. 1996; Ding et al. 2005).
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The relevance of this to the specific measurements at the Blacks Drain site is limited as the study site was fallowed during the 2003 measurements, and minimally covered by cane during the 2005 period. However, under extensive cane growth, and more pertinently at the Cudgen Lake site, this is potentially an important regulator in the transport of VSCs between the soil and atmosphere. Similarly, there is a case for an examination of the role played by biogenic macropores (such as old root channels) as well as earthworm or other mesofaunal burrows in the enhanced release of VSCs particularly from naturally occurring ASS.
Summation
From the above reasoning the most logical conclusion to the measured H2S emissions is that they are a product of high rates of sulfide production in the reducing zone, resulting in the rapid build-up of hydrogen sulfide, leading to its discharge into the atmosphere, primarily by molecular diffusion with the possible assistance of shear planes and bubble ebullition. Periods where production of H2S exceeds its consumption within these sediments are certainly occurring, most likely during the night time. The sedimentary and porewater chemistry indicates that the system is not sulfur-limiting. Also, the constant emission of H2S suggests that ebullition has minimal influence, which would normally be characterized by intense fluxes of relatively short duration on a daily basis (Chanton et al. 1989). Although, as mentioned, the emission process is favoured during the night-time, the variability of these measurements shows that definitive conclusions cannot be reached on the factors influencing its emission.
Regardless of the method of formation and release, the data shows (although not consistently) a diurnal shift in H2S concentrations, with levels increasing during the night time. This is a pattern that has been previously well established in various coastal marshes and tidal flats (Aneja 1984; Azad et al. 2005; Bodenbender et al. 1999; Goldberg et al. 1981; Hansen et al. 1978). However, variations to this trend have also been observed, with an inverse pattern in H2S emissions found by Carroll et al. (1986). Cooper et al. (1987b) found that highest emissions occurred during the early- to mid- afternoons and lowest in the mornings from wetland soils in Florida. This diurnal trend is discussed in further detail with reference to the SO2 measurements in the section directly following.
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5.5.3 SO2 Measurements
In discussing the measurements of SO2 made by this study it is pertinent to first return to the previously summarised argument (refer to Section 2.2.1), that SO2 concentrations 2- from ASS were possibly due to the evaporation of soil waters containing SO3 , since atmospheric measurements of SO2 were closely linked to surface soil evaporation (Macdonald et al. 2004a). Again, this is clearly relevant within a S cycling system such 2- as in ASS, where SO3 is an important intermediate in the oxidation of reduced sulfides, being formed through the partial oxidation of elemental sulfur (Equation 5.13);
0 2- + S + O2 + H2O bbb SO3 + 2H [5.13]
Whilst this may be one factor in SO2 measurements, the results from this study indicate an alternative explanation; whereby the emissions of SO2 are directly linked, via photo- oxiation, to the formation of H2S. Volatile sulfur gases engage in a variety of chemical and photochemical oxidation reactions in the atmosphere (Plane 1989). The literature shows that it is unlikely that H2S would oxidise to SO2 via direct photolysis, as it does not absorb UV radiation at wavelengths greater than 260 nm (Cox and Sandalls 1974). However, this reaction will proceed in the presence of photochemically produced free radicals such as OH, NO2 and O3, which explain why tropospheric chemical conditions are mildly oxidising. The resulting atmospheric lifetime of H2S is relatively short, being usually between 4.3 to 12.3 hrs (Jaeschke et al. 1980; Spedding and Cope 1984; Wang and Howard 1990).
The stoichiometry of these reactions is shown in Figure 5.29. The hydroxyl radical will remove most gases that are emitted into the atmosphere, despite relatively low tropospheric concentrations (Crutzen and Zimmermann 1991; Hynes and Wine 1989).
The rate-determining step is therefore the oxidation of H2S by OH (Wang and Howard
1990), after which the product [HS] can be oxidised by O3 or NO2, with NO2 reacting
20 times faster than with O3 (Wang et al. 1987). The oxidation kinetics of reduced sulfur gases have been extensively reviewed both experimentally and theoretically; see
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 129
the references above in addition to Lovejoy et al. (1987), Tyndall and Ravishankara (1991), and Wilson and Hirst (1996).
OH
NO2 H2S + OH b HS + H2O O3
b b HS + NO2 HSO + NO HS + O3 HSO + O2
HSO + NO2 b HSO2 + NO HSO + O3 b HSO2 + O2
b HSO2 + O2 SO2 + HO2
Figure 5.29. Reaction pathways involving the photo-oxidation of H2S by the free radicals OH, NO2 and O3, ultimately forming SO2. Adapted from Berresheim et al. (1995) and Shooter (1999).
As the free radicals are photochemically produced, they follow a diurnal pattern similar to the SO2 concentrations, peaking during the day time and declining during the night time. This can explain the diurnal pattern in H2S concentrations measured in this and a variety of other studies, as mentioned previously. This reasoning could also possibly explain why we haven’t been able to measure sulfur dioxide within fully-grown cane, where sunlight penetration is restricted (unpublished data). However, the presence of a cane canopy also infers a decreased rate of soil surface evaporation. Additionally, measurements of sulfite in the surface soils by (Macdonald et al. 2004a) could simply 2- be due to the oxidation of H2S in surface sediments by dissolved O2 yielding SO3 , 2- S2O3 and other polysulfides (Luther III et al. 2001).
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However, whilst this trend is particularly evident during the chosen 32 hr period in the 2003 dataset, it is interesting to note that it is only partially present throughout the remainder of the dataset, suggesting a combination of factors at work, rather than simply a photo-oxidative effect. Also providing further insight are the average day versus night time fluxes compared to peak concentrations. During the 32 hr period in
2003 identified in Figure 5.6, a large concentration of H2S at the 0.5 m sampling point is clearly evident, yet a relatively low flux exists at the corresponding time point (Figure
5.10). Conversely, the SO2 concentrations are relatively low (compared to H2S), but the corresponding fluxes over those particularly time points are far greater than H2S, particularly during the day time. This point is played out in the average flux values for the two gases in both sample periods. The average day time flux for SO2 is ~ 22 times larger than the H2S flux during the 2003 period, and ~ 6 times greater in the 2005 set.
Conversely during the night time, the H2S flux is approximately equal to that of the SO2 flux during the 2003 period, and half the SO2 flux average during the 2005 period. To theorise then, if H2S photo-oxidation was accounting for the observed daytime increase in SO2, then a night-time dominance of H2S flux would be expected. This is clearly not shown in the data, suggesting two points to consider. First, that the daytime production of H2S is greater than the night time, leading to large SO2 fluxes via photo-oxidation; and second, that photo-oxidative effects of H2S are not necessarily the dominant cause of the SO2 fluxes observed. These points are not mutually exclusive and the measurements made during both field trips could conceivably be a combination of both factors.
Another aspect to the SO2 measurements aside from the photo-chemical effects, is the role of microorganisms in the production of SO2, as several heterotrophic groups of bacteria are capable of SO2 or sulfite production (Bremner and Steele 1978). Alexander
(1977) suggested that SO2 can be produced from the decomposition of sulfur-containing amino acids. Specifically, the desulfination of cysteine sulfinic acid (a bacterial intermediate in the conversion of cysteine to sulfate within soil) can produce SO2 and alanine (Alexander 1974; 1977; Freney 1960), the reaction stoichiometry of which is shown in equation 5.14 below;
O=HO-S-CH2-CH(NH2)-COOH bbb SO2 + CH3-CH(NH2)-COOH [5.14]
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 131
Although this bacterial activity is important in the wine production industry, little research has been pursued within the cycling of sulfur in sediments.
5.5.4 Climatic Variable Interactions
It is logical to conclude that concentrations and flux of gases from sediments are influenced by climatic variables. However, the results from this work give an insight on the specific variables which have the greatest control over SO2 and H2S emissions. A greater emphasis will be placed on the interpretations of the 2005 results owing to its more detailed measurements.
The results, particularly from the statistical analyses, suggest that both the concentrations and fluxes of the two gases are primarily influenced by two competing groups of climatic factors. On the one hand, is the influence of temperature which includes primarily air and soil (5, 10 and 20 cm) temperature, and to a lesser degree latent heat, and net radiation. On the other hand there are the moisture related variables, with soil moisture (5, 10 and 20 cm) and precipitation being the principle agents. SO2 (flux and concentration) is consistently correlated in the positive direction with the temperature, and negatively with the moisture related climatic variable. The inverse is true for H2S concentration and flux.
SO2
This general interpretation brings up several interesting points. Firstly, the idea that SO2 is negatively correlated to the moisture variables is not all that unexpected. SO2 is an extremely soluble gas (9.4 g / 100 g water; see Chapter 2), so it is not surprising that its concentration and flux would be reduced under increased-moisture conditions. Additionally, bulk gas diffusion transport is suppressed under high soil water contents, possibly leading to a build-up of gases until porespace changes, due to evaporation / evapotranspiration, result in their release. This effect is best shown in Figure 5.20, where the five rain-events during the 2005 sampling period are succeeded by extremely low SO2 flux periods of 8 to 14 hrs, after which relatively high fluxes are observed.
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 132
Secondly, the positive correlation between SO2 and the temperature variables supports both hypotheses regarding the possible evolution of SO2 from ASS. The idea that SO2 emissions from ASS could be the result of the evaporation of sulfite-containing porewater (Macdonald et al. 2004a) is completely plausible based on this data.
However, so is the possibility that the measurements of SO2 are from the photo- oxidation of H2S, which would also be represented by positive correlations of SO2 with temperature and net radiation. Indeed, the strong diurnal fluctuation in SO2 flux and concentration does not really distinguish between these two possible mechanisms.
Also of note is the influence of barometric pressure. Clearly, there would appear to be a relationship between the major fluxes of SO2 (and H2S) and with the atmospheric pressure (Figure 5.18 & 5.19). However, the lack of consistent communality between the gases and atmospheric pressure in the statistical analysis would suggest no definite causative effect on SO2 (and H2S) flux. Despite this, it is an area worthy of continued investigation owing to other research which has shown increased gas emissions from reductions in atmospheric pressure, as detailed in Chapter 4. This is especially the case regarding the possible enhancement of this phenomenon in the presence of fractures (Auer et al. 1996) or shear-planes, as described previously. On a related point, the pressure changes induced by high wind velocity also need further investigation. High wind speed at the ground surface increases surface turbulence and thus decreases the immediate pressure above the ground surface (via the Bernoulli Principle). Flechard et al. (2007) suggested that this effect was partially responsible for the temporal changes in CO2 storage within the soil profile.
It is interesting to observe that the wind directed SO2 flux (Figure 5.26, top) did not show the same covariance as indicated in the overall sampling dataset. It should be noted also that although there are many algorithms for cluster analysis, there is no generally accepted best method, introducing a rather large subjective component to the assessment (Manly 2005). The clustering of the variables will ultimately depend on the algorithm used, suggesting a possible reason for the difference for the clustering of the wind-directed samples.
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 133
H2S
In the case of H2S, it is expected that there is no negative parallel with the moisture variables (as was the case with SO2), as H2S(g) has a far lower solubility in water; 0.33 g / 100 g. However, the results showing the opposite, positive, correlation with these moisture variables is a curious finding. Interestingly, H2S correlates to a greater degree with the soil moisture levels at 10 and 20 cm than it does with the surface soil
(5 cm). With the generation of H2S most certainly due to bacterial action, it could be suggested that an increase in soil moisture, especially at these depths, stimulates the bacterial production of H2S. This finding would suggest that aerobic bacteria, below
10 cm, might play an important role in the production of H2S from ASS.
The zero-flux periods observed after the rainfall events can be explained by excessive soil moisture restricting the movement of air through the profile. At a simplistic level, any precipitation would have the effect of reducing the oxygen level within the soil profile. Indeed, it has been demonstrated that periodic water saturation enhanced another reduced gas flux (CH4) from agricultural soils in North America (Chan and Parkin 2001a). Although the decrease in oxygen was undoubtedly a factor, that study established that changes in CH4 flux were reliant also on changes associated to the nitrogen bacterial community (Chan and Parkin 2001b), which also have the ability to metabolise CH4 (Hanson and Hanson 1996). How this translates in the case of H2S
(instead of CH4) is still unclear, as little research of this nature has been performed within agricultural soils. However, one relevant point that has been demonstrated, is that the presence of free sulfide inhibits the reduction of N2O, allowing for its accumulation in surface sediments (Senga et al. 2006; Senga et al. 2001). This interaction could contribute to the emissions of N2O observed from sugarcane cultivated ASS (Denmead et al. 2005; Weier 1999). Although this is largely beyond the scope of this thesis, there are seemingly many questions still to be answered regarding the interactions between the biogeochemical element cycles within ASS.
The final aspect of the two-way, inverse correlation hypothesis between H2S and moisture / temperature variables is the negative relationship between H2S and temperature. It is commonly accepted that bacterial growth is a function of temperature, with bacterial activity approximately doubling for every 10°C rise in temperature. The
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 134
effect of temperature on sulfur gas emissions from sediments is equally as important. A variety of studies have shown that H2S concentrations and fluxes are strongly dependent on radiation and soil temperature (Bodenbender et al. 1999; Delmas et al. 1980; Harrison et al. 1992). Although it would be easy to suggest that this factor was more profound in cold-climate areas (and not as marked in the sub-tropical soils measured here), the studies included a variety of coastal and agricultural soils in France and Ivory
Coast in addition to German and English sediments. Emissions of H2S, DMS and CS2 in a S. alterniflora wetland were also primarily explained by changes in sediment temperature (Cooper et al. 1989; Cooper et al. 1987a). These findings lend weight towards the notion that H2S is being photo-oxidised to SO2, with the temperature rise (causing increased bacterial action) being insufficient to account for its loss due to oxidation. This is partially supported by the soil temperature measurements which showed diurnal variations from between; 3.18 - 7.90°C at 5 cm depth, 1.61 - 4.14°C at 10 cm, and 0.37 - 1.41°C at 20 cm.
Something that also needs to be considered in any extension of these measurements is the influence of any vegetation on emissions, and climatic interactions on vegetation. Indeed, emissions of VSCs from plants are also strongly dependent on light, with Fall et al. (1988) showing a tripling in the DMS flux (from corn) with a transition from dark to light conditions, the opposite effect of what is shown regarding H2S in this study.
Similar results were also obtained for H2S, MSH and CS2 for alfalfa and wheat (Fall et al. 1988).
5.5.5 Implications for Measurements
The flux values from the data presented, clearly indicates that SO2 and H2S are being emitted from coastal ASS, with measured values exceeding the background levels of between 0.05 and 1.2 ppb for SO2 (Seiler and Sigel 1988), and 0.11 and 0.33 ppb for
H2S (EPA 1993). Any extrapolations or generalisations of trace gas measurements from such a study are fraught with inaccuracies; owing largely to wide variations in both temporal and spatial scales (Chanton and Whiting 1995; Stewart 1989). Even with detailed remotely sensed data, the extrapolation of single measurements cannot be extended to beyond anything other than the local scale. It is more relevant to look at the
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 135
factors controlling trace sulfur gas emissions from ASS for a superior guide as to regional and global gas budgets.
Nonetheless, the novelty of the measured H2S flux value from ASS is worth extending to a wider context, as it is a long-held view that significant H2S measurements come only from marine environments, or limited to near-shore environments such as estuaries and salt-marshes (Adams 1996; Andreae 1990; Giblin and Weider 1992). As the lifetime of many reduced sulfur gas species is in the order of seconds to hours, real-time measurements are important to regional budgets, which in turn aid the interpretation of the global sulfur cycle.
Indeed, averaging the flux values over the sample period shows that the emission of H2S from the agricultural ASS is comparable to that emitted by some salt- and freshwater systems (Table 5.7). Although tremendously variable in nature (see values below), these are generally classified as high-productivity sediments which are major contributors to the global biogenic H2S budget, with emissions of gaseous S compounds at least 10- to 100-fold greater than emissions from oceans or inland soils (Steudler and Peterson 1984). The Blacks Drain ASS, in any gas species inventory, would most likely be mapped as agricultural lands rather than drained backswamps, and therefore assigned an
H2S flux value of zero. Clearly then, there is a case to extend these novel measurements of H2S fluxes to other areas to see if the values are consistent across differing coastal agricultural ASS locations.
As shown at the bottom of Table 5.7, the measurements from Blacks Drain are extremely close to previous measurements of inland agricultural and urban soils in
France and the USA. However, these two other measurements of H2S were made using chamber methods, the problems with which have been described previously. Therefore, on balance, those measurements were more likely an overestimation of the actual average emissions of H2S from agricultural soils. Little work has subsequently been performed on the emission of H2S from agricultural soils, despite early indications that inland soils appeared to contribute significantly to total sulfur fluxes, see Adams et al. (1981a).
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It should be noted that many measurements of H2S from agricultural soils have been shown to be well in excess of the measurements made as part of this study; e.g. Goldan et al. (1987), Lamb et al. (1987). However, some of these studies included measurements over different vegetation types / crops, which have been shown to emit
H2S (and DMS) (Aneja 1990). A study by Aneja et al. (1981) showed that inland soils -2 -1 (U.S.A) had an average H2S flux of 0 0.013 g S m yr .
Table 5.7. Variations in average biogenic H2S emissions from various wetlands/coastal marshes.
Location Emission Rate Reference (g S m-2yr-1)
Saltwater *Saline marshes (Louisiana, USA) 0.02 - 601.6 (Adams et al. 1981b)
Saline marshes (Florida, USA) 0.003 - 0.015 (Castro and Dierberg 1987)
Tidal mud-flats (Sumiyoshi, Japan) 0.021 - 0.079 (Azad et al. 2005)
*Salt-marsh (Louisiana, USA) ~ 0.022 (DeLaune et al. 2002)
Brackish-marsh (Louisiana, USA) ~ 0.186 (DeLaune et al. 2002) Freshwater
Swamp (Florida, USA) 0.001 - 0.006 (Castro and Dierberg 1987)
Marsh (Louisiana, USA) ~ 0.003 (DeLaune et al. 2002)
Other
Equatorial forest (Ivory Coast) 0.30 - 0.88 (Delmas et al. 1980)
Wet meadow (Germany) 0.21† (Bartell et al. 1993)
Agricultural
Inland agricultural & urban soils (France) 0.04 (av.) (Delmas et al. 1980) (Adams et al. 1979; Adams et al. Inland agricultural soils (USA) 0.06 (av.) 1981b) Blacks Drain (NSW, Australia) ~ 0.056 2003 Measurements Blacks Drain (NSW, Australia) ~ 0.036 2005 Measurements * denotes study sites were dominated by Spartina alterniflora † denotes peak flux values
As demonstrated by the results presented, the flux of SO2 was much greater than for -2 -1 H2S, being approximately 6 times larger during the 2003 period (~ 0.31 g S m yr ), and 2½ times larger during the 2005 sampling period (~ 0.095 g S m-2yr-1). As the biogenic emissions of SO2 are somewhat limited within the literature, comparisons to other measurements are also restricted, as shown in Table 5.8. As is the case with H2S, the flux of SO2 from the agricultural ASS would appear approximately equal to some saltwater tidal flats. Again, although these measurements are temporally isolated and
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 137
have a limited representation towards other agricultural ASS, the measurements still warrant further examination.
If these two measurements are extrapolated to the geographical extent of ASS within Australia (conservatively estimated at 9 Mha), then as a first estimate, it would suggest that ASS potentially contribute around 0.009 Tg S yr-1 (2005 measurement) to 0.028 Tg S yr-1 (2003 measurement). It is somewhat difficult putting these numbers into any context, as there is little data to compare them against. An earlier figure by Ayers and Granek (1997) indicated that terrestrial sources (vegetation and soils) contributed 0.09 Tg S yr-1 to Australia’s atmospheric sulfur budget. It should be noted that the flux estimate by Ayers and Granek was deemed highly uncertain by the authors, as it was based on previous South American and African measurements. However, such a value indicates that Australian ASS contribute in the region of 10 to 31 % of the terrestrial biogenic sulfur emissions to the atmosphere based on these two sets of measurements.
Table 5.8. Variations in average biogenic SO2 emissions from available data.
Location Emission Rate Reference (g S m-2yr-1)
Saltwater
Tidal mud flat (Sumiyoshi, Japan) 0.043 (Azad et al. 2005)
Tidal sandy mud flat ( “ ) 0.38 ( “ )
Tidal sand flat ( “ ) 0.40 ( “ )
Agricultural ASS Blacks Drain (NSW, Australia) ~ 0.31 2003 Measurements Blacks Drain (NSW, Australia) ~ 0.095 2005 Measurements
5.5.6 Preliminary Conclusions
This study confirms the long experienced notion that variations in the natural emission of sulfur gases are extremely high. Whilst this is only a preliminary study into H2S and
SO2 emissions from coastal ASS, it clearly shows that both gases are being emitted from such an agricultural soil. The diurnal pattern as well as the relationship between the two gases suggests that the H2S is a possible source of the measured SO2 through a photo-oxidative effect, with the SO2 ultimately oxidised to sulfate.
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 138
The measured flux values of H2S and SO2 indicate that drained acid sulfate environments, such as those under sugarcane cropping, are a potentially significant unaccounted source in the biogenic atmospheric sulfur cycle. The area sugarcane occupies in Australia is over 500,000 ha, of which approximately 30 % is on ASS (CSIRO 2002). This indicates that there may well be a substantial regional issue in terms of crop sustainability, as well as environmental and human health. It should be remembered though that this particular study was conducted on a fallow ASS field, and potential impacts from growing sugarcane are not accounted for.
Chapter Five: Advanced Gaseous Field Measurements - H2S & SO2 139
Chapter Six: LABORATORY-BASED SULFUR GAS MEASUREMENTS GAS CHROMATOGRAPHY
6.1 INTRODUCTION
The results from the previous chapter show that despite the oxidising conditions within the uppermost metre of the profile at Blacks Drain, it is still able to emit a reduced sulfur gas, H2S, as well as SO2. This finding presents a number of potential research directions when looking at such systems. This includes the determination of other possible VSCs being emitted from ASS, and the identification of in which depths their formation occurs. In pursuing these questions, an alternative analytical technique was sought to the field-based micrometeorological approach. In order to delineate between the different sulfur gases potentially present, a laboratory-based gas chromatography method sampling the headspace above ASS samples was employed.
The detection of trace levels of VSCs, even using gas chromatography is a difficult process. Therefore, there is a strong focus within this chapter, on the introduction / background and methodology used, as aspects of the technology employed are still being refined within the field of environmental analysis. This is followed by sample depth-measurements from both the Blacks Drain and Cudgen Lake study sites, and a detailed interpretation of the results.
Chapter Six: Laboratory-based Sulfur Gas Measurements 141
6.1.1 Gas Chromatography
Gas Chromatography (GC) is a separation technique in which volatile compounds pass through a column containing a stationary phase at rates proportional to their volatilities, or their boiling points (Kealey and Haines 2002). For the following experiments the GC was configured with a flame photometric detector (FPD). The FPD is a highly selective detector with low detection limits for the determination of analytes containing sulfur or phosphorus. It consists of a hydrogen-rich flame which excites the molecular species of interest to higher electronic states, which on returning to their ground state emit light at characteristic wavelengths (Westmoreland and Rhodes 1989). The optical filter isolates the specific wavelength, which in the case of sulfur compounds is at 394 nm due to the formation of the activated S2 molecule (Gilbert 1970; Westmoreland and Rhodes 1989). The photomultiplier tube then detects the specific light wavelength where it is converted to an electrical signal, and then amplified by the electrometer. A detailed schematic of the FPD used for this study is shown in Figure 6.1.
Figure 6.1. Schematic representation of the Shimadzu FPD 17A. Image courtesy of the Shimadzu flame photometric detector 17 Ver. 2; User’s Manual.
Chapter Six: Laboratory-based Sulfur Gas Measurements 142
6.1.2 Solid-phase Microextraction
Solid-phase microextraction (SPME) has been shown to be a simple and effective adsorption and desorption technique with many advantages over traditional analytical methods. The major advantage is its ability to combine the sampling, preconcentration, and transfer of analytes into a single procedure, thereby negating the need for complicated equipment before transfer into a standard gas chromatograph (Alpendurada 2000; Koziel et al. 2000). Although SPME was only developed by Pawliszyn and his colleagues in the late 1980s, it has been successfully applied in numerous fields, including; food and beverage aromas and flavourings, forensic, pharmaceutical and clinical applications; examples of which are shown in Table 6.1.
Table 6.1. Some examples of non-environmental SPME applications. Application Analyte Author(s)
Food & Beverage Aroma
Volatile sulfur Wine (Fang and Qian 2005; Mestres et al. 1998) compounds
(Hill and Smith 2000; Miracle et al. 2005; Beer “ Scarlata and Ebeler 1999)
Milk “ (Vazquez-Landaverde et al. 2006)
Strawberry Puree “ (Schulbach et al. 2004)
Turkey Breast “ (Fan et al. 2002)
Medical / Forensic Testing
Blood & Urine Antihistaminics (Nishikawa et al. 1997)
Volatile compound Urine (Mills and Walker 2001) metabolism Pharmaceuticals, Review articles (Theodoridis et al. 2000; Ulrich 2000) forensics, illicit drugs
Within the environmental field, research utilising SPME has primarily centred on the analysis of pesticides and organic contaminants in soil and groundwaters, and various air pollutants. Some research using SPME has focused on soil gas emissions (Fuster et al. 2005), although this has mainly been in the field of soil fumigation and odour research. Some examples of these applications are shown in Table 6.2.
Chapter Six: Laboratory-based Sulfur Gas Measurements 143
Table 6.2. Some examples of environmental-based SPME applications. Application Analyte Author(s)
Chemical Contamination
Groundwater Pharmaceutical (Moeder et al. 2000) Contamination Contaminants
Chlorobenzenes, (Florez Menendez et al. 2000; Fromberg et Soil Contamination Nitroanilines, BTEX al. 1996; Santos et al. 1997)
Review article Multiple Herbicides (Krutz et al. 2003)
Air Pollution
Volatile Sulfur Coal Stockpiles (Kozinc et al. 2004) Compounds Trimethylamine, Agricultural Odours (Kim et al. 2002) volatile S compounds Volatile Sulfur Industrial Effluent Odours (Lestremau et al. 2004) Compounds VOCs, formaldehyde & General Air Pollution (Koziel et al. 1999) particulate matter
The articles listed in the two tables above, as well as reviews by Alpendurada (2000) Pawliszyn (1997; 1999) and Vas and Vekey (2004), highlight the various applications in which SPME now plays a vital role in extraction procedures, many of which include the SPME of volatile sulfur compounds from liquid and solid samples.
SPME Apparatus The SPME device itself resembles a modified syringe (Figure 6.2), consisting of a separate fibre holder and fibre assembly. The fibre assembly, wholly encased within the fibre holder, contains a 1-2 cm long retractable SPME fused-silica fibre which is coated with a thin polymer film conventionally used as a coating material in chromatography (Vas and Vekey 2004). The fibre assembly acts as protection for the fragile fibre, allowing for it to be exposed to the environment. The fibre coating acts like a sponge, by concentrating the analytes of interest by adsorption/absorption processes (Vas and Vekey 2004). Several fibre coatings are available depending on the analytes of interest, with different fibres offering different selectivity. One of the most common materials used as the thin polymeric film is polydimethylsiloxane (PDMS). Coating thickness and volume will determine the necessary extraction time for the fibre. Some of the
Chapter Six: Laboratory-based Sulfur Gas Measurements 144
commercially available coatings include; polyacrylate, and mixed phases of PDMS- divinylbenzene (DVB), Carboxen-PDMS, Carbowax-DVB and Carbowax-templated resin (Alpendurada 2000). A full discussion of the properties of commercial SPME coatings can be found in Mani (1999).
The carboxen-PDMS fibre that was used as part of this study (see section 6.2), is graphically represented within Figure 6.2. As shown, the carboxen particles are immobilised on the fused silica fibre with the PDMS acting as glue to hold them together. The high capacity of the carboxen-PDMS fibre to concentrate specific analytes is due to the carboxen (mixed carbon) component having a surface area ~ 1000 m2 g-1 with small micropores (Vas and Vekey 2004).
PDMS fibres have been used widely in SPME applications because of their ability to withstand high injector temperatures (< 320ºC), as well as their versatility in extracting both polar and non-polar compounds (Alpendurada 2000). As a result, carboxen-PDMS fibres have been the SPME of choice in the analysis of volatile compounds such as H2S; see: Burbank and Qian (2005), Fang and Qian (2005), Haberhauer-Troyer et al. (1999), Hill and Smith (2000), Mestres et al. (1999b), Miracle et al. (2005), Nielsen and Jonsson (2002b), Popp et al. (1999), Scarlata and Ebeler (1999).
Chapter Six: Laboratory-based Sulfur Gas Measurements 145
Figure 6.2. Schematic representation of the SPME device, and a magnified representation of carboxen- PDMS fibre. Diagrammatical representations modified from the Sigma-Aldrich website, and Alpendurada (2000).
General SPME Theory The general theory of operation behind SPME involves immersing the fused silica fibre into a liquid sample or gaseous headspace above a liquid or solid. After a period of time, a concentration-equilibrium is established between the sample matrix and the extraction phase, beyond which no further analytes can accumulate on the fibre (Lord and Pawliszyn 2000). The concentration-equilibrium conditions can be represented as;
fs ⋅⋅⋅ ⋅⋅ sf ⋅⋅⋅ CVVK 0 n === fs ⋅⋅⋅ +++ VVK sf [6.1]
Where n is the number of moles extracted on the SPME coating, Kfs is the fibre coating distribution constant, Vf is the fibre coating volume, Vs is the sample volume, and C0 is the initial concentration of a given analyte in the sample. The analytes of interest are absorbed onto the fibre and then thermally desorbed within the injection port of the gas chromatograph.
Chapter Six: Laboratory-based Sulfur Gas Measurements 146
As with all newly applied methods, there are some limitations to SPME. These include the fragile nature of the fibres themselves, which can be easily broken as part of the sampling and injection procedures. Another potential problem is the carry-over of analytes on the fibre, which can be difficult to eliminate even at the highest temperatures as set down by the manufacturers (Alpendurada 2000). Such issues merely heighten the need for extensive control samples as well as the continual conditioning of the fibres to ensure accurate reproducibility and integrity of results.
Another problem that is harder to counter is competitive interference from other compounds not necessarily being analysed, which may displace the target analyte from the surface of the fibre. This can be a potential issue with the carboxen-PDMS fibre, as it has only a limited number of adsorption sites, and higher molecular compounds can displace their lower molecular weight compound counterparts (Lestremau et al. 2003b; Murray 2001). Murray (2001) further identifies that the relative proportions of the components adsorbed onto the fibre depend on their ratio in the headspace, meaning that any internal standard should contain the same relative concentrations of all analytes present in the samples. The quantification of analytes within complex matrices can only be achieved under non-equilibrium conditions, that is, short extraction times (Lestremau et al. 2003a; Tuduri et al. 2002).
6.2 MATERIALS
6.2.1 SPME
The SPME fibre assembly and holder used for this study were from Supelco (Bellefonte, U.S.A). The fibres used were Carboxen™/Polydimethylsiloxane (CAR/PDMS) 75 )m (Cat. No. 57318). The SPME fibre was conditioned according to the manufacturer’s guidelines prior to its use (300°C for 1-2 hours).
Chapter Six: Laboratory-based Sulfur Gas Measurements 147
6.2.2 Chromatography
A Shimadzu GC-17A (Ver. 3; Shimadzu Oceania, Sydney) gas chromatograph equipped with a (FPD-17 Ver. 2) flame photometric detector, with a sulfur filter (394 nm), was used for the GC analyses. The injection was in the splitless mode. Ultra- high purity helium was used as the mobile phase at a pressure of 80 kPa (1.0 mL/min column flow), whilst the zero-grade air and high purity hydrogen were set to 60 kPa. The separation was performed using a Supel-Q™ PLOT fused-silica capillary column (30 m, 0.32 mm ID; Supelco, Bellefonte). The software interface; CLASS-VP Chromatography Data System (Ver. 4.3; Shimadzu Oceania, Sydney), was used for system control and data acquisition. A schematic diagram of the setup is shown in Figure 6.3. Blanks were run frequently (< every 10 samples) to ensure that there was no carry-over contamination on the SPME fibre and / or within the column.
Figure 6.3. Schematic representation of the fundamental GC setup employed for this study.
Chapter Six: Laboratory-based Sulfur Gas Measurements 148
A list of the specific temperature parameters employed for the method used here is shown below in Table 6.3.
Table 6.3. Temperature parameters used in the GC / FPD method. Component Temperature
Injector 290°C Detector 1 200°C Detector 2 280°C
Oven Temperature Program
Initial temperature 45°C Initial hold time 1.0 min Temperature rate increase 15°C / min Final temperature 250°C
Calibration An external calibration procedure was carried out using both dimethylsulfide (DMS) (Fluka, > 99.0% puriss) and ethanethiol (ESH) (Fluka, > 97.0% purum). All dilutions were made using Hamilton teflon-barrelled microlitre syringes in absolute analytical- grade ethanol (UNIVAR Ajax). Standards were also made up in ethanol, contained in crimp-top 20 or 100 mL glass vials, with aluminium tear-away seals and PTFE / silicone liners (all from Alltech Associates, Sydney, Australia). The external calibration technique was used owing to its simplicity of use, and because the standard addition technique could not be applied as the ASS sample containers were sealed at the study site on immediate extraction from the ground.
The calibration of H2S and SO2 was attempted using solid chemicals (Na2S and NaSO3) to create both gases. Although the reactions produced the desired gases, the method was quantitatively unsuccessful owing to the inability to achieve consistent concentrations of both gases. Even though this meant that no standard curve could be generated, the identification of the peak retention times enabled the presence / absence determination of H2S and SO2 from the sediment headspace samples. Because of their low molecular weight, specifically H2S, their calibration using GC and SPME have frequently proved troublesome; see Nielsen and Jonsson (2002a).
Chapter Six: Laboratory-based Sulfur Gas Measurements 149
Quantification The response of the FPD to a sulfur compound is not linear and varies between different compounds, in accordance with the following relationship (Farwell and Rasmussen 1976);
n I ∝∝∝ IO [S] [6.2]
Where I is the observed intensity of the optical emission due to the S2 species, [S] is the concentration of the sulfur atoms, and IO and n are constants under set experimental conditions. In the case of this experiment the value of n was taken as 2, a common approximation; for example Chin and Lindsay (1993), Karvaly et al. (2005), Lestremau et al. (2004). This assumption implies that the energy emitted from the decay of the excited S2 molecule will be approximately related to the square root of the S concentration. Even though it has been shown that this is not always the case (Burnett et al. 1977; 1978; Farwell and Barinaga 1986; Farwell and Rasmussen 1976; Gaines et al. 1990), the linear regression showed it to be a very close approximation. It should be noted that quantification using the external calibration technique described above could potentially lead to an overestimation of the concentration in the soil samples, because of the matrix changes (Lestremau et al. 2004; Zeng and Noblet 2002).
6.2.3 Sampling Containers
Septa jars (125 mL I-CHEM / Thermo Fisher Scientific, Waltham, U.S.A) were used as the incubation storage containers. They were chosen because of their large volume and open-top polypropylene closures bonded with 1.143 mm thick PTFE / silicone septa. The large volume made the jars ideal for soil sampling, and the septa provided the inert surface necessary when dealing with volatile and semi-volatile compounds.
6.2.4 Sampling Location
Samples were taken during May 2006 from both the Blacks Drain and Cudgen sites previously described in Chapter 3. The Blacks Drain samples were from the same sugarcane paddock as was sampled during Oct-Nov 2005 (see Figure 5.1), whilst the
Chapter Six: Laboratory-based Sulfur Gas Measurements 150
Cudgen samples were from the backswamp area shown in Figure 3.7 (July 2004 location). Sediment samples were also taken from within Blacks Drain itself and along the lake edge at Cudgen. As these sediments were partially submerged, they were extracted using a sleeve corer with only the uppermost section used.
Samples (~ 30-60 g wet weight) from the two study sites were carefully inserted into the wide-mouth glass jars whilst wearing latex gloves. They were immediately sealed and transported back to UNSW for analysis, at ambient temperatures (18 - 22°C) and sealed from any source of light. They were then analysed within 48 hrs of sampling.
6.2.5 Extraction Procedure
The optimisation of the adsorption and desorption times were based on the response of the various gases to their standards and interpretation of residual values from the blanks, respectively. Adsorption time was therefore set at 2 mins (at ambient laboratory temperature; 22°C) and desorption to 5 mins. Samples adsorbed to the fibre were injected directly into the GC within seconds of the completion of the extraction period. A diagrammatical and pictorial representation of this procedure is shown in Figure 6.4.
Chapter Six: Laboratory-based Sulfur Gas Measurements 151
SPME Holder
PTFE / Silicone Septum 125 mL Septa Jar
Exposed Car/PDMS Fibre
Headspace
Soil Sample
Figure 6.4. Schematic and photograph demonstrating the use of SPME with the septa jars.
6.3 RESULTS
6.3.1 Calibration
The results of the external calibrations of DMS and ESH are shown in Figure 6.5. The relationship between the square root of the response of the GC (peak area) and the standards is very good, with both correlations > 0.99. The errors bars (S.E. of mean) also highlight the reproducibility of the method. It would appear from the current method that the limit of detection (LOD) is approximately 1-2 )g/L for both DMS and ESH.
Chapter Six: Laboratory-based Sulfur Gas Measurements 152
Figure 6.5. Calibration curves for DMS (top) and ESH (bottom), and correlation coefficients.
6.3.2 Blacks Drain Samples
The levels of DMS and ESH measured at the Blacks Drain site are shown in Figure 6.6. The concentration of DMS was approximately 100-fold greater than ESH at their respective maxima, with DMS peaking at 353.63 )g/L in the 1.25 m sample, and ESH peaking at 3.94 )g/L within the surface (top 50 mm) sample. As can be seen in Figure 6.6, there was a clear switch over in the dominance of the two gases down the profile.
Chapter Six: Laboratory-based Sulfur Gas Measurements 153
ESH was present (albeit in only trace concentrations) within the organic topsoil and oxidised layer, elevations where DMS was not detected. At around the beginning of the transition zone, which periodically experiences reducing conditions, DMS increased from below its LOD, to its maximum at the depth of the watertable (~ 1.25 m), before decreasing slightly within the completely unoxidised sulfidic layer. The 0.5 m sample was taken specifically within the brown peat layer identified in Section 3.2, and notably as can be seen in Figure 6.6, this was where the second peak in ESH occurs.
Figure 6.6. DMS and ESH concentrations down a soil profile taken at Blacks Drain during May 2006.
Chapter Six: Laboratory-based Sulfur Gas Measurements 154
Sediment samples were also taken from within Blacks Drain itself, immediately below the water surface. Only DMS was detected within these samples. Figure 6.7 shows the chromatograph of one of the samples, along with its calculated concentration (18.27 )g/L) based on the calibration procedure. The average concentration for the samples taken within Blacks Drain itself was 17.49 )g/L.
Peak Area = 18.27 )g/L
Figure 6.7. Sediment sample chromatograph from within Blacks Drain taken during May 2006, and the calculated DMS concentration.
6.3.3 Cudgen Lake Samples
Figure 6.8 shows the concentrations of DMS measured at the Cudgen Lake site. No ESH was detected in these samples. The levels of DMS were near the LOD within the brown oxide surface layer, after which they rapidly increased after ~ 0.1 m within the unoxidised fibrous layer. The maximum concentration (324.01 )g/L) was reached with the unoxidised layer (~ 0.45 m), the deepest sampling point.
Chapter Six: Laboratory-based Sulfur Gas Measurements 155
Figure 6.8. DMS concentrations down a soil profile taken at Cudgen Lake Nature Reserve during May 2006.
Portions (10-15 mins) of the chromatographs for the Cudgen surface layers are shown in Figure 6.9. As can be seen from all three chromatographs, numerous larger molecular weight sulfur compounds are also present within the brown oxidised surface samples. Even though they are unable to be identified or quantified, they are clearly present in the sample, with some exceeding the peak areas compared to DMS concentrations (see Figure 6.9; c).
a. b. c.
Unknown higher molecular weight DMS DMS DMS S compound
Figure 6.9. Sample chromatographs from the Cudgen Lake site surface sample.
Chapter Six: Laboratory-based Sulfur Gas Measurements 156
An additional sample was taken from the surface of the Cudgen study site (i.e. the surface sample as described previously). Before sealing the septa jar, sodium azide
(NaN3; 0.5 % w/w) was added to the sample. NaN3 has been used previously as a general inhibitor of metabolic activity (Gaillardon 1996; Ning et al. 1996; Rugge et al.
1999; Sijm et al. 1998; Taylor and Viraraghavan 1999). Even though NaN3 has been shown to react with some inherent compounds (Goel et al. 2003), it was thought interference would be minimal, and it would be suitable for a first analysis. The chromatographic result of the NaN3 addition is shown in Figure 6.10. The only peak that can be positively identified from the standards tested is ESH. The first peak is tentatively identified as SO2, based on the retention time of the compound when compared to the dissolution of NaSO3 and HCl. The second peak is more loosely identified (as MSH), based on the retention time obtained from literature provided by SUPLECO regarding their use of the column.
SO2? MSH? ESH
Figure 6.10. Chromatograph of a surface sample taken from the Cudgen Lake site with the addition of NaN3.
Samples were also taken from the edge of Cudgen Lake itself. The location of these samples is described in detail by (Macdonald et al. 2004b). These sediment samples, as was the case with the Blacks Drain / drain samples, were from immediately below the water surface. Figure 6.11 shows the chromatograph of one of the samples, along with its calculated concentration of DMS (2.23 )g/L). As shown, no other compounds were detected, with the average DMS concentration for the Cudgen Lake-edge samples being 1.41 )g/L.
Chapter Six: Laboratory-based Sulfur Gas Measurements 157
Peak Area = 2.23 )g/L
Figure 6.11. Sediment sample chromatograph from Cudgen Lake edge taken during May 2006, and calculated DMS concentration.
6.4 DISCUSSION & PRELIMINARY CONCLUSIONS
6.4.1 Dimethylsulfide
DMS is a major contributor to the global oceanic sulfur budget, but research into its contribution from terrestrial (especially agricultural) sources has been limited. The presence of DMS with marine and freshwater locations (as discussed previously in Chapter 2) has long been recorded, occurring from: the degradation of the tertiary sulfonium compound DMSP; the metabolism of methionine; or through the bacterial methylation of thiols, especially in anaerobic sediments. The formation of DMS has also been observed from the bacterial reduction of dimethylsulfoxide (DMSO) by a variety of micro-organisms in marine (Simo et al. 2000) and coastal environments (Kiene and Capone 1988; Lopez and Duarte 2004). It has been shown that the phylogenetic cluster Roseobacter, appears to play an important (yet also not exclusive) role in not only the reduction of DMSO in marine environments, but also more generally regarding DMSP turnover (Gonzalez et al. 1999; Simo 2004; Zubkov et al. 2001). Alternatively, DMSO can be reduced to DMS within marine or hypersaline environments, with certain sulfate-reducing bacteria (e.g. Desulfovibrio desulfuricans) using it as a terminal electron acceptor (Jonkers et al. 1996). DMS has also been produced from DMSO through the reductive action of purple non-sulfur bacteria (e.g. Rhodovulum spp.) and green sulfur bacteria (e.g. Chlorobium spp.); see Vogt et al.
Chapter Six: Laboratory-based Sulfur Gas Measurements 158
(1997). With DMSO contributing around 20% (and up to 50%) of the algal dimethylated sulfur pool (Simo and Vila-Costa 2006), the role it plays in sulfur inter- conversions in coastal marine environments clearly needs greater clarification (Lee and de Mora 1999). It has also been recently suggested that intracellular DMS is directly released to the (marine) environment from phytoplankton under oxidative stress, e.g. ultraviolet radiation, CO2 or Fe limitations (Sunda et al. 2002).
The catabolism of DMS also needs considering in this discussion, with both aerobic and anaerobic bacteria able to metabolise DMS. The aerobes capable of catabolising DMS include Acidithiobacillus (De Zwart and Kuenen 1992; Kanagawa and Kelly 1986) and Hyphomicrobium (Suylen et al. 1987). Importantly, as identified by Visscher and Taylor (1993a; 1993b), some specific strains of Acidithiobacillus are able to catabolise DMS by methyl transfer (rather than oxygenases) by using nitrate as an electron acceptor. This enables either the aerobic or anaerobic metabolism of DMS to proceed under fluctuating oxygen conditions, making it more common (than the oxygenase pathway) in natural environments (Visscher and Taylor 1993b), such as in inter-tidal, or other sediments with shifting watertable levels.
Blacks Drain Site Previous studies of porewater DMS concentrations have primarily been conducted in coastal or marine sediments rather than agricultural soils. The concentrations measured at the Blacks Drain and Cudgen sites are compared with other research values in Table 6.4. Table 6.4 shows the substantial variation in DMS concentrations from the previously published studies, owing primarily to natural variations as well as in part to the different methodologies used. The measurements from the ASS are towards the upper limit of this range. Interestingly though, the maximum concentrations of DMS measured by the other studies shown in Table 6.4, occurred immediately below the sediment-air interface, usually within 10 mm of the surface. The peak of DMS at Blacks Drain was at 1.3 m.
Chapter Six: Laboratory-based Sulfur Gas Measurements 159
Table 6.4. DMS porewater concentrations of this study compared to other published research. Max. Conc. Location Method Used Reference (μM)
Salt marsh (S. alterniflora In-situ porewater 0.08* (Howes et al. 1985) dominated), USA sampler Salt marsh (S. alterniflora Core-cut sampling / 20* (Howes et al. 1985) dominated), USA centrifugation Sediment Continental shelf marine, Peru 0.12* (Andreae 1985) headspace Freshwater estuarine sediment, 0.1 Pressure filtration (Sorensen 1988) Denmark Cyanobacterial mat on salt marsh, (Visscher et al. 20 Slurry headspace USA 1995) Diatom-covered inter-tidal sand flat, (Visscher et al. 8 Slurry headspace USA 1995) (Visscher et al. Carbonate sediment, unknown < 1 Slurry headspace 1995) Agricultural ASS, Blacks Drain, Sediment 5.69 This study Australia headspace / SPME Coastal Lake backswamp, Cudgen Sediment 5.22 This study Lake, Australia headspace / SPME * Approximate values only from graphical representations
The lack of repeats and variation in the samples make interpretation of the results somewhat speculative. It would appear, judging by the depth-concentration profile (Figure 6.6), that the DMS measurements are the result of the anaerobic, bacterially- driven methylation of thiols. Although this process is less well understood, compared to DMSP degradation (Stets et al. 2004), it is thought to occur from the methylation of
H2S and methoxylated aromatic compounds (Bak et al. 1992; Drotar et al. 1987; Finster et al. 1990) or from the addition of a methyl group to methanethiol, which can form from the degradation of methyl amino acids, e.g. methionine (Kiene and Capone 1988). The stoichiometries of these reactions are shown in Equations 6.3 and 6.4.
R-O-CH3 + H2S bbb R-OH + CH3SH [6.3]
R-O-CH3 + CH3SH bbb 2R-OH + CH3SCH3 [6.4]
Alternatively MSH can also be transformed to DMS (Equation 6.5), which has been demonstrated with incubations of MSH within wastewater sludge sediments (Sipma et al. 2002; van Leerdam et al. 2006). However, its formation in these studies appeared
only as an intermediate, with the primary product of the mechanism being CH4.
Chapter Six: Laboratory-based Sulfur Gas Measurements 160
2CH3SH bbb CH3SCH3 + H2S [6.5]
The almost ubiquitous presence of heterotrophic H2S-methylating bacteria in soil (Jorgensen 1983) along with the abundance of free sulfide within ASS (as suggested by the consistently high DOP values at both study sites), lends support to the hypothesis that DMS formation is being driven by thiol methylation. Our previous measurements of H2S emissions from agricultural ASS also affirm this concept. If this process dominates the formation of DMS within the Blacks Drain cane field site, then the mineralisation of organic matter (producing methyl group donors) would be the controlling factor in determining its formation. Whilst this process may be accumulating DMS (and potentially MSH) it would also be reducing the concentrations (and therefore the toxicity towards the sugarcane) of sulfide within the profile. Although these are interesting issues, without any further measurements of the potential precursors of DMS, this conclusion is only speculative.
Visscher et al. (1995) found that the presence of nitrate increased levels of DMS under anoxic conditions, which suggested an influence of nitrate respirers. Visscher et al. (1995) went on to suggest, that nitrate respirers may continuously consume DMS under fluctuating oxygen concentrations, provided there is a constant supply of nitrate. Lomans et al. (1999a), however, found that within the freshwater sediments measured, owing to restrictions in nitrate concentrations (< 10 )M), nitrate degradation (via nitrate-reducing bacteria) of DMS is limited. Whilst this may be true within the Cudgen Lake system, the application of nitrogen fertilisers to sugarcane areas such as at Blacks Drain may be causing the lack of DMS within the surface samples. Indeed, it has been shown that nitrogen fertiliser application can alter soil microbiological activity, including measured decreases in methane consumption rates owing to a population switch towards ammonium oxidisers (Steudler et al. 1996). With nitrogen fertiliser (in the form of urea) being applied to sugarcane crops at average rates of between 120 kg N/ha to 160 kg N/ha (Thorburn 2004), the alterations to the bacterial community may be influencing DMS concentrations in surface soils. Similarly, Bremner and Bundy (1974) showed that the nitrification of ammonium in soils was retarded by the production of thiols (MSH), DMS and H2S. From this perspective it would seem that
Chapter Six: Laboratory-based Sulfur Gas Measurements 161
the creation of VSCs in ASS have a beneficial side, with its reduction of nitrogen losses in their gaseous forms.
Also of note is that the decomposition of 3-MPA, a bacterially driven product of DMS demethylation in anoxic sediments, can be demethiolated to methanethiol and subsequently methane (van der Maarel and Hansen 1997). Additionally studies by Kiene et al. (1986) have shown that increases in concentrations of DMS increased production of methane, with Lomans et al. (1999a; 1999c) suggesting that methanogenesis was the major mechanism for the consumption of DMS (and MSH) in freshwater sediments, as the rates of degradation for DMS were similar to that of methylotrophs, and only three strains of sulfate-reducers have been identified as capable of growing on a DMS (or MSH) substrate (Tanimoto and Bak 1994). It is interesting to note the large decrease in DMS concentration (by ~ 250 μg/L) at the deepest sampling point (1.7 m) within the Blacks Drain profile (Figure 6.6). This reduction in DMS concentration could be due to its use as both a competitive and / or non-competitive substrate for the formation of the methane concentrations at Blacks Drain measured by Denmead et al. (2006). The stoichiometry of degradation of DMS can produce methane as well as H2S (with MSH as an intermediate), as indicated by Lomans et al. (1999b) in freshwaters (see Equation 6.6). Interestingly, a highly-diverse community of methanogenic archaea have been identified at the Blacks Drain study site (Durr et al. 2006), providing supporting evidence for this process.
- + 2CH3SCH3 + 3H2O bbb 3CH4 + HCO3 + 2H2S + H [6.6]
Whilst it is suggested that DMS may be a substrate for methane production, the reverse may also be true with methane acting as a substrate for sulfate-reducing bacteria at the sulfur and carbon redox interface. Methane is oxidised microbiologically both under aerobic and anaerobic conditions (Nedwell 1996). This process may be enhancing the formation of DMS, occurring anaerobically within the sulfate-reducing zone via the following process;
2- - - CH4 + SO4 bbb HS + HCO3 + H2O [6.7]
Chapter Six: Laboratory-based Sulfur Gas Measurements 162
Cudgen Lake Site The marine-dominance of the Cudgen Lake backswamp samples, would suggest that these measurements of DMS were likely as a result of the degradation of DMSP (see Equation 2.19), arising from the decomposition of sedimented phytoplankton material. This process has been observed in other marine-dominated locations (Cerqueira and Pio 1999; van der Maarel and Hansen 1997). The other marine-associated processes of DMSP/DMS formation, as described previously, may also contribute to its measurements. Examples of this are phototrophic purple and green bacteria which could be important contributors to DMS emissions from areas such as the Cudgen Lake backswamp, because of their ability to use DMSO as an electron acceptor from highly sulfidic habitats.
However, the DMS profile (Figure 6.8), with concentrations increasing with depth, could suggest that DMS production is occurring below 1.0 m and is subsequently being depleted (microbial and / or chemically) at shallower depths. Indeed, the depth profiles from other estuarine (and marine) sediment porewater studies show a DMS peak at (or just below) the oxic surface, followed by a rapid decline within the sulfide production zone (Andreae 1985; Howes et al. 1985; Sorensen 1988). The measurements would suggest then that there is a greater role played within the Cudgen backswamp samples by the bacterial methylation of thiols, as described previously for Blacks Drain. Generally speaking, the broad geomorphological differences between the two sites are not greatly disparate, making the assumption that similar processes regarding DMS formation between sites, a plausible conclusion. Both sites are underlain by sulfidic sediments which are Holocene-age estuarine deposits, with the Blacks Drain samples differing by being progressively overlain by alluvial inputs with an increasingly fluviatile composition. However, these hypotheses need to be substantiated by ancillary information, in particular regards to the dimethylated sulfur pool; DMSP + DMSO.
Regardless of the processes forming DMS it would appear that the concentrations of DMS are controlled by the watertable elevations and thus redox boundary at both study site locations. This point has been demonstrated also by Cerqueira and Pio (1999), who found that the main controlling influence determining DMS emissions into the atmosphere was surface water concentration and thus redox delineations. Similarly,
Chapter Six: Laboratory-based Sulfur Gas Measurements 163
Bodenbender et al. (1999) identified that DMS concentrations showed a distinct dependency on the tidal pattern within a tidal mudflat in Germany.
6.4.2 Ethanethiol
Thiols can be formed by a number of distinct pathways, with those in sediments principally derived from the microbial conversion of methionine (Mopper and Taylor 1986; Zinder and Brock 1978c); but they can also form via the successive microbial demethylations of DMSP (Yoch 2002), or alternatively from the abiotic methyl addition to sulfides / polysulfides / methoxy-S-compounds (Mopper and Taylor 1986; Vairavamurthy and Mopper 1987). However, these processes generate only the more dominant thiols found in marine and freshwater sediments, and not ESH.
Blacks Drain Site While a number of studies have detailed emissions of DMS from coastal areas, the formation of trace thiols (especially ESH) in agricultural sediments has been an unexplored area of research. Therefore, the measurement of ESH within only top layers of the agricultural ASS is an interesting finding. Again, this is primarily because it is not considered one of the dominant thiols, with methanethiol (MSH) and 3- mercaptopropionate the most commonly measured in marine and freshwater sediments (DeLaune et al. 2002; Sorensen 1988). However, trace concentrations of ESH have been identified within marine sediment slurries (Mopper and Taylor 1986). ESH has also been measured in incubated samples of soil and leaf litter from a deciduous forest within the southern Appalachians (U.S.A) and a tropical rainforest (Cost Rica) (Haines 1989; Haines et al. 1987). Levels of ESH (cumulatively measured with DMS) were suggested to be the result of plant community establishment; acting as anti-microbial or anti-herbivory agent. ESH has also been measured in ambient air with large concentrations existing within geothermally active areas, and lesser concentrations within the city of Rotorua, New Zealand (Li and Shooter 2004). It has also been measured in higher concentrations from anthropogenic wastewater sediments (van Leerdam et al. 2006).
Chapter Six: Laboratory-based Sulfur Gas Measurements 164
One possibility that has been suggested, is that ESH may form in a similar manner to that exhibited by MSH formation from methionine (Lynch and Harper 1980; Oremland et al. 1988; Visscher and Taylor 1993a). This occurs through the microbial metabolism of larger, ethylated macromolecules, and has been demonstrated with ESH shown to increase in sediments amended with ethionine or S-ethyl cysteine under both aerobic and anaerobic conditions (Banwart and Bremner 1975; Oremland et al. 1988). Interestingly, a budding species of yeast (Saccharomyces cerevisiae) which underpins the most common type of fermentation, catalyses the production of ESH from ethionine and MSH (Cherest et al. 1970), as well as producing sulfite (Alexander 1974).
Another possibility is that the measurements of ESH could be due to the catabolism of DMS. Visscher and Taylor (1993b) found that when DMS (and ethyl methyl sulfide) was metabolised by a particular strain of Acidithiobacillus (thioparus T5, isolated from a marine microbial mat), ESH was formed in trace levels. However, the concentrations of ESH were only transient, being enhanced when in the presence of tributylphosphine, a chemical oxidation retardant of ESH. Levels of ESH were ultimately degraded to sulfate and CO2.
An alternative scenario that may account for the presence of ESH within the topsoil of the agricultural ASS is specifically related to the sugarcane crop under which the measurements were made. Indeed, it is possible for ESH to be formed by the reaction of
H2S and ethanol (Equation 6.8). Even though this process has only been documented within the wine production industry (Amerine 1980; Rauhut 1993), the presence of ethanol from fermenting sugarcane trash, and H2S from sulfate reduction, is a feasible combination within this environment. Additionally, ethanol production has been observed within stands of S. alterniflora, an alcoholic fermentation process induced by root oxygen deficiencies (Mendelssohn et al. 1981). However, this stress response would be more applicable within the anaerobic zone and not within the topsoil where the ESH was measured, suggesting it not to be the precursor in this instance. Another more likely source of ethanol is from decaying plant matter (Seco et al. 2007; Warneke et al. 1999).
2C2H5OH + 2H2S + ½O2 bbb 2C2H5SH + 3H2O [6.8]
Chapter Six: Laboratory-based Sulfur Gas Measurements 165
If fermenting sugarcane trash is the precursor of ESH measurements, then there are certain implications in the shift from the traditional burning of the crop prior to harvest, to the new green harvest program currently being undertaken for co-generation of electricity by the NSW sugar milling industry. The decline in burning of the residual green trash could potentially alter the amount of fermentable sugars being left after the mechanical harvesting process, leading to variations in ESH production, and therefore, trace metal export / retention within agricultural ASS (discussed in section 6.4.4).
Another possibility that may factor into the ESH measurements within only the upper horizons could be a result of a biotic process with the sugarcane rhizosphere. Although there is only a minimal amount of research in this area, studies have shown that the volatilisation of sulfur compounds from soils may be assisted by rhizosphere microorganisms (Siman and Jansson 1976).
The degradation of ESH within these systems presents another unknown aspect to ASS research. Its anaerobic degradation has been shown to be quite difficult (van Leerdam et al. 2006). Oremland (1981) measured trace amounts of ethane (in addition to larger quantities of methane) in anoxic estuarine sediments in San Francisco Bay, California. Its production appeared to be stimulated by the presence ESH, a process which would be dependent on methanogenic flora at the specific location (Oremland et al. 1988). However, the very low production rates would suggest the process was not responsible for the majority of ESH present within the system. Nonetheless, there is the possibility of these sediments producing, and potentially emitting ethane, as a process of ESH degradation. A more likely outcome for the measured ESH is its oxidation to the corresponding disulfide, a process discussed in more detail below, along with its absence from the Cudgen Lake samples.
Cudgen Lake Site As mentioned above, the absence of ESH in the Cudgen Lake backswamp samples is an unusual discovery. Reasons why ESH (and other thiols) were not measured in the Cudgen samples could be due to microbial turnover, consumption or sedimentary interactions, which often results in their effective removal from porewaters (Kiene 1991). In a study at Massachusetts marsh, Luther III et al. (1986a) found no measurable
Chapter Six: Laboratory-based Sulfur Gas Measurements 166
levels of thiols within the profile, which the authors suggested may be due to the lack of sulfide oxidation, and sulfate depletion at the site. The reducing nature of the majority of the Cudgen profile would also suggest that this may be a possibility.
Alternatively, its absence may be the result of changes to the organic-rich sediment matrix of the Cudgen samples. This may have been a factor with the headspace SPME technique, with Ng et al. (1999) illustrating the matrix dependence of organophosphorus pesticide concentrations, with the technique experiencing a much lower affinity for retaining the compounds when working in humic substrates. In regards to thiols, they have been shown to readily react with sediment surfaces, a process which effectively adsorbs or binds the organic sulfur compounds (Kiene 1991). Kiene (1991) went on to postulate that the bound thiols may be exchangeable, implying an availability for reaction and metabolism within porewaters, a point which is dependent, however, on the individual thiol.
More specifically, regarding the absence of ESH, it has been shown to be a substrate for Acidithiobacillus bacteria (methanethiol oxidase) under laboratory conditions (Gould and Kanagawa 1992; Suylen et al. 1987), which allows for its oxidation to diethyldisulfide (not analysed for) and eventually sulfate. This bacteria, since isolated from a marine microbial mat, has also been demonstrated to oxidise ESH itself using oxygen or nitrate as its electron acceptor (Visscher and Taylor 1993a). Sipma et al. (2004) was able to show in biomass reactors that in the presence of MSH, ESH was successfully converted to sulfate. This process has yet to be shown in sediments though.
The measurement of ESH over MSH within these sediments is somewhat surprising though, as MSH is readily produced from bacteria found in soil, and in addition to being present in marine sediments is the most abundant methylated sulfide in freshwater systems (Caron and Kramer 1994; Kadota and Ishida 1972). MSH can form either through the bacterially induced enzymatic cleavage of methionine (Kadota and Ishida
1972), or from the reduction of sulfate and subsequent methylation of H2S (Drotar et al. 1987). It may very well be the case that MSH was present in the samples at the time of extraction, but as identified by Chin and Lindsay (1994) and Haberhauer-Troyer et al. (2000; 1999), MSH readily oxidises to dimethyldisulfide and dimethyltrisulfide in the presence of oxygen, transition metals or free radicals. In addition to it being consumed
Chapter Six: Laboratory-based Sulfur Gas Measurements 167
by sulfate-reducing bacteria, methanogenic bacteria can also compete for MSH as a substrate for the production of methane (Kiene et al. 1986). Along a similar rationale, it was demonstrated by Devai and DeLaune (1995) within an anaerobic salt marsh soil, that the composition of individual sulfur gases emitted was strongly dependant on the redox condition of the sediment.
It is interesting to note, that in the sample to which NaN3 was added, a peak tentatively identified as MSH was apparent (refer back to Figure 6.10). Although conclusions based on a single sample are extremely hazardous, there is merit in the further examination of the notion that MSH is present within the sediments sampled and is being microbially transformed prior to analysis.
6.4.3 Emissions of Volatile Sulfur Compounds to the Atmosphere
Unlike the production of VSCs in surface waters, the formation of these compounds within sediments is expected to be strongly affected by microbial metabolism before emission to the atmosphere (Sorensen 1988). On a general level though, where concentration gradients do not reach the sediment-air boundary, it would be unlikely that any emissions to the atmosphere of VSCs would be occurring. In reference to this study then, as the concentration profile for DMS does not reach the surface it would be unlikely that DMS would be emitted from the Blacks Drain site sediments.
Under laboratory conditions, the catabolism of DMS has been shown within pure cultures by a variety of anaerobic (sulfate-reducers and methanogens) and aerobic bacteria (Acidithiobacillus) (Finster et al. 1990; Kiene 1993; Visscher et al. 1991). Therefore, it would be anticipated that the DMS measured at depth would be transformed within the oxidised layer, and also within the reduced zone. This has been confirmed in other sediments where even though there appears to be a continuous production of DMS and thiols, their steady-state concentrations are relatively low due to the microbial affinity for these compounds (Lomans et al. 1999a; Lomans et al. 2002; Lomans et al. 1997). The initial results would therefore only confirm that these acid sulfate sediments have a high capacity to produce DMS and ESH.
Chapter Six: Laboratory-based Sulfur Gas Measurements 168
However, it is entirely plausible that DMS at Cudgen Lake (and ESH at Blacks Drain), where concentrations reach the profile surface, could be emitted to the atmosphere. Kiene and Hines (1995) were able to show that approximately 30 % of the DMS produced within the low-pH Sphagnum peat-bog (New Hampshire, USA) was emitted to the atmosphere. In addition, the production of DMS within the top layers of the sediment was enough to explain these emissions to the atmosphere.
The emission of DMS from the Cudgen samples would also likely be a seasonal phenomenon, as standing water retards the emission of slightly soluble gases such as DMS (MacIntyre et al. 1995). This would suggest that during the summer period, when rainfall is at its peak and the area sampled most likely underwater, that DMS emissions to the atmosphere would be minimal.
This notion of pressure forcing reduced S gases through the oxidised horizon to the atmosphere has been witnessed in a number of tidal sediments (see Chapter 5 Discussion). The basic proposition is that the incoming tide creates an hydrostatic pressure, forcing the emitted gases to the surface by convection, after which emissions often drop owing to presence of a water column exerting pressure downwards. This effect can manifest itself as short, intense pulses of gas release (Cooper et al. 1987b; Jorgensen and Okholm-Hansen 1985).
An alternative mechanism causing a potential release of DMS or ESH to the atmosphere can occur where the volatile compounds are generated from microbial DMSP degradation. Van Bergeijk et al. (2002) found that disturbances which cause osmotic shocks in micro-organisms, such as a rainfall event, can lead to transient accumulations of DMS (and possibly ESH under the appropriate conditions) in porewaters, leading possibly to an atmospheric flux.
Another factor which needs to be considered when discussing possible emissions to the atmosphere is the influence of the overlying vegetation. Indeed, as mentioned previously, DMS emissions in coastal wetlands can largely occur from certain vegetation types, such as S. alterniflora leaves rather than from the sediment itself (Dacey et al. 1987; Morrison and Hines 1990), indicating that biomass type is also an important factor. Not only do emissions of DMS occur from plants that can accumulate
Chapter Six: Laboratory-based Sulfur Gas Measurements 169
DMSP, but DMS has also been shown to be the predominant sulfur gas being emitted from most plants (Lamb et al. 1987); with specific measurements from trees (Lamb et al. 1987); and the crops maize, corn, soybeans, rice and wheat (Fall et al. 1988; Goldan et al. 1987; Kanda et al. 1995; Lamb et al. 1987; Nouchi et al. 1997). As such, DMS has been observed to occur as the primary sulfur gas emitted from vegetated wetlands
(de Mello et al. 1987), whilst H2S is the predominant species from unvegetated tidal flats (Jorgensen and Okholm-Hansen 1985; Wakeham and Dacey 1989). Its role in plant functions is still somewhat unclear (Yoch 2002).
In terms of sugarcane, there has been little definitive evidence to show the crop contributes substantial DMS emissions to the atmosphere. However, Paquet et al. (1994) found that under favourable conditions, sugarcane (Saccharum spp.) accumulated DMSP. This led Paquet et al. (1994) to suggest that sugarcane may make a significant contribution to terrestrial DMS emissions. Although many reasons have been suggested for its accumulation in most species of sugarcane; e.g. osmoregulation (Paquet et al. 1994), detoxification of excess sulfur (Stefels 2000), its actual function remains unclear (Colmer et al. 2000; Otte et al. 2004). Regardless of the means, DMS accumulation occurs primarily in sugarcane leaves, with it being enzymatically released when leaf tissue is damaged (Godshall 1988). More recently though, DMSP accumulation has been measured with sugarcane roots in addition to that in leaf and stem tissue (Colmer et al. 2000). Therefore, there is the possibility that this process could account for the presence (or part thereof) of the measured DMS. The concentrations of DMSP measured as part of that particular study (from Saccharum officinarum), were between 4.25 - 4.85 )mol/g (dry weight) within the sugarcane roots (Colmer et al. 2000). The levels of DMSP would suggest that it plays only a minor role, especially at concentrations equal to that of coastal marine environments as measured within the ASS at Blacks Drain. As such, the results would therefore indicate that any release of DMS to the atmosphere would more likely come from diffusion/ebullition rather than plant-mediated release.
There also exists an argument for the presence of vegetation in the formation of thiols in this agricultural topsoil, from the microbial degradation of DMSP. Further to the information provided above, DMSP has only been measured in two other higher plant generas; Spartina (cordgrass) and a dicotyledonous strand plant Wolastonia biflora
Chapter Six: Laboratory-based Sulfur Gas Measurements 170
(Hanson and Gage 1996; Otte et al. 2004; Paquet et al. 1994). Otte et al. (2004) suggests that its presence in sugarcane could be as a herbivore deterrent, sink for excess sulfur, or more likely as an osmoregulator in response to drought conditions. However, the absence of MSH, in conjunction with / or in place of, the ESH measurements would suggest that the sugarcane is an unlikely contributor to the measured thiol values.
6.4.4 Volatile Sulfur Compounds & ASS
The formation and turnover of volatile sulfur compounds as components of the biogeochemical cycling within ASS has yet to be fully addressed. These initial measurements of DMS and ESH would suggest that they may be important for several reasons aside from the fact that these measurements show that ASS are a potential unaccounted global source of DMS and ESH; a point discussed in more detail in Chapter 7.
Clearly, through possible processes of formation, both volatile compounds are integral in the cycling of sulfur between the organic and inorganic phases during early diagenesis (Mopper and Taylor 1986; Vairavamurthy and Mopper 1987). Measurements by Luther III et al. (1986b) of thiol cycling in salt-marshes suggested that sulfur is transformed from inorganic to organic species during warmer, more productive seasons, and conversely from organic to inorganic during cooler months. The authors of that paper go on to suggest that part of this cycling may include the reaction with, and formation of, solid inorganic and organic sulfur phases. One particular pathway for the formation of thiols in such an environment is the reaction of organic compounds with sulfur anions (such as thiosulfate), as shown in Equation 6.9;
2- - - ROH + S2O3 →→→ RS + HSO4 [6.9]
Research has also shown that pyrite oxidation may be an important factor in the production of thiols, potentially serving as a link between the organic and inorganic sulfur pools in some salt marsh sediments (Luther III et al. 1986a). Interestingly, Luther III et al. (1986b) found that thiol production peaked at the same depth as for
Chapter Six: Laboratory-based Sulfur Gas Measurements 171
optimal pyrite oxidation, suggesting pyrite is a possible starting material for thiol production (Equation 6.10);
2- 2+ FeS2 + R (organic matter) →→→ RSH + SO4 + Fe [6.10]
Alternatively, polysulfides, an important intermediate species in pyrite oxidation (Luther III et al. 1986b), have been shown by Vairavamurthy et al. (1992) to react with organic carbon to form organic polysulfides, which may act as precursors to volatile organic sulfur compounds (Visscher 1996). The above points are especially important in areas where reactive iron is limiting, as large concentrations of organic sulfur would be expected to dominate the sedimentary sulfur phase (Aizenshtat et al. 1981; Filley et al. 2002; Francois 1987; Hartgers et al. 1997). An example of this is within the upper half metre of a S. alterniflora marsh in Delaware (U.S.A), where Ferdelman and colleagues (1991) found that up to 60 % of the total S inventory was present as organic S. The relative abundance of reactive iron at the Blacks Drain site would suggest that organic sulfur incorporation was a more important process at the Cudgen site, where DOP values exceed 90 % and sulfate reduction rates are most likely higher (Chapter 3). The organic S would be effectively competing, not necessarily exclusively, with pyrite as a long-term sedimentary sink for reduced sulfide.
As is pointed out by Luther III et al. (1986b), Equation 6.10 could account for the high turnover of pyrite in some sediments and for the accumulation of ferrous iron at depths concurrent with optimal pyrite formation. Using polarographic methods, Luther III et al. (1986a) found that thiols occurred exclusive to inorganic sulfide, peaking above the sulfidic zone where (in that particular study) the pH was at its minimum along with excess interstitial sulfate concentrations. This phenomenon was also noted by Zhang et al. (2004), who also identified that no correlation existed between thiols and DOC within selected freshwater and saltwater wetlands in Canada. Although the data from Blacks Drain shows that the peak in ESH is at the surface, the secondary peak (at ~ 0.5 m) could be of a similar nature, where the pH is rapidly decreasing alongside a high turnover of sedimentary S. However, the fact that the peak in ESH measurements is occurring well away from the area of sulfide production would suggest the abiotic
Chapter Six: Laboratory-based Sulfur Gas Measurements 172
pathway for thiol formation is less important unless micro-niches of reducing conditions are present in the upper horizons within the Blacks Drain sediment profile.
The exact role that pyrite formation and oxidation plays in the creation of these compounds remains unclear. It has been documented that in sulfide-rich sediments (such as within the reducing zones of ASS), the formation of VSCs is significantly higher than sulfide-poor sediments (Lomans et al. 2002). This suggests that the anaerobic methylation of sulfide is the primary driver for VOSC formation in sulfide/organic-rich sediments (Lomans et al. 2002). The high levels of sulfide in ASS would also point towards sulfur being bound to organic matter during the early stages of diagenesis (Hartgers et al. 1997), meaning organic S could therefore be a competitive source of sedimentary sulfur within ASS, as suggested above. However, as highlighted by Yang et al. (1996) in a S fertiliser addition experiment, there is no simple correlation between the emissions of VSCs and total soil sulfur.
Other studies have also demonstrated the importance of thiols, in particular, in the cycling of sulfur between organic and inorganic phases. Adam et al. (1998) demonstrated the possibility of photochemically inducing the rapid formation of sedimentary organo-sulfur compounds (including thiols) from inorganic sulfur 0 precursors (H2S and S ) under natural conditions. Roy and Trudinger (1970) also suggested that the thiol glutathione is the actual enzymatic intermediate for S2- and S0 oxidations in the formation of sulfate.
The measurement of thiols within ASS has implications for the cycling of trace metals, particularly Cu. Leal and van den Berg (1998) have demonstrated that thiols (cysteine and glutathione) form stable complexes with Cu(I) in seawater. Similarly, in anaerobic salt marsh sediments, it has been suggested that > 90 % of total interstitial Cu exists as Cu-organic S complexes (Boulegue et al. 1982). The implications of these measurements suggest that organic S ligands (in the form of thiols) could be a controlling factor in the cycling and release of heavy metals such as Cu in agricultural ASS, such as at Blacks Drain. Research into the toxicity of these metals has largely focussed on the reactivity and dissolution of inorganic sulfur species, e.g. Ankley et al. (1991) and Di Toro et al. (1992). This is an unexplored area in ASS research which requires further investigation to help elucidate potential release mechanisms and the
Chapter Six: Laboratory-based Sulfur Gas Measurements 173
bioavailability of heavy metals in discharge waters. This is countered by the fact that the reactivity of thiols (including ESH) is pH dependent being usually higher at elevated pH, with ESH having a pKa value of approximately 10.2 (van Leerdam et al. 2006).
The roles that DMS and its precursor, DMSP, play in assimilatory sulfur chemistry within sugarcane, especially in relation to ASS, are areas that also require further research. As mentioned previously, sugarcane is only one of the relatively few higher plants able to accumulate DMSP as part of the assimilatory uptake of sulfate. It has been offered that the production of DMSP within higher plants could be a way of dissipating excess reduced sulfur as a method for keeping cysteine and methionine at low equilibrium concentrations (Stefels 2000). Therefore, it is tempting to hypothesise that in areas of high reduced sulfur soil content (such as in ASS or other marine-derived landscapes), that DMS emissions would be elevated from plants able to accumulate DMSP. However, this oversimplifies what is a complex relationship between sulfide concentrations and plant growth. There is little evidence to show that there is any relationship between sulfide and DMSP accumulation, with one study by Morris et al. (1996) on S. alterniflora, showing that additional sulfide had only marginally significant effects on DMSP concentrations in leaf tissues and no effect in roots.
Furthermore, the accumulation of DMSP within sugarcane is potentially dependent on nitrogen availability (Colmer et al. 2000) in a manner similar to that observed with S. alterniflora (Colmer et al. 1996). If indeed the accumulations of DMSP are a method for regulating sulfur, then an understanding of the interactions between sulfur and nitrogen would be especially beneficial to the agricultural management and fertiliser application of sugarcane cultivation on ASS. This is particularly the case with sugarcane on ASS, since sugarcane does not appear to suffer from the effects of sulfide toxicity. Again, further research into the interactions is necessary to help elucidate some of these questions. One area that may shed some light on the issue is the condition of the rhizosphere in sugarcane, and whether it maintains an oxidising or reducing interface within the sulfidic sediments. Whilst oxidation around pre-existing sugarcane root channels is widely apparent in the ASS of eastern Australia, whether this is the case when the plant is living may help in the understanding of not only sulfur uptake (i.e. preference for sulfide or sulfate uptake), but also on the issue of broadscale oxidation of these sediments.
Chapter Six: Laboratory-based Sulfur Gas Measurements 174
6.4.5 Methodological Issues / Difficulties
The measurement and quantification of low concentrations of highly volatile sulfur compounds, even using SPME enrichment, is a difficult task (Popp et al. 1999). The major problem associated with these experiments is the ex-situ nature of the analysis, with direct in-situ measurements being more desirable. Howes et al. (1985) found that concentrations of DMS in sediment porewaters increased when the samples were disturbed, specifically as a result of root damage. As the study by Howes and colleagues was within a S. alterniflora salt marsh, an indirect comparison of the results can only be made. However, if sugarcane roots and rhizomes are important in the production of DMSP, then disturbance of the soil would most likely, result in the increase of DMS production within the upper sections of the profile where the majority of the roots are located. The results here did not demonstrate this (Figure 6.6), suggesting that root disturbance does not influence DMS production under sugarcane. Yang et al. (1996) also discovered that DMS concentrations, in particular, increased after incubation, peaking around 2-3 days after sampling. It should be pointed out that organic manure and chemical fertilisers were added to the soil within the study by Yang and colleagues.
In contrast to the enhancement of DMS, is the possible degradation of VSCs between sampling and analysis. Other studies (Lomans et al. 1999a; van Bergeijk et al. 2002) found the degradation of DMS (and MSH) from incubations mimicking aerobic in-situ conditions to be low in freshwater sediments due to oxygen limitation. This was not the case though for anoxic samples which showed high aerobic DMS consumption rates. The primary mechanism identified by Lomans et al. (1999a) for the degradation of DMS within these samples was methanogenesis, once again as a result of the limitation of oxygen. Variations in the sulfur porewater and solid-phase speciation will also have an impact on gaseous emissions as a result of changes to preferred pathways (Kiene and Hines 1995). There is the possibility that some of the in-situ aerobic samples (i.e. from the upper horizons) became anaerobic in the sample jars before analysis through the bacterial consumption of any remaining oxygen. This is simply an inherent problem with analysing the samples a long distance from their origin, and could be addressed by
Chapter Six: Laboratory-based Sulfur Gas Measurements 175
analysing samples in the field. These points further underscore the need to avoid pH and redox alterations induced from ex-situ analysis.
It also needs to be considered, especially in regards to ESH, that the measurements were potentially an artefact of the sampling process. As mentioned previously, there is little evidence in the literature of the compound being routinely present in sediment samples, and some debate exists as to whether it is a naturally occurring sulfur compound in sediments (Oremland et al. 1988). In the study by Haines et al. (1987), the experimental work that derived values of ESH were similar to those performed here, in that soil and leaf litter samples (including living plant materials) were sampled in enclosed polypropylene containers. There is a possibility that within the upper, aerobic Blacks Drain samples a shift towards anaerobic conditions in combination with the organic matter may have accentuated the concentrations of ESH. Unfortunately, no evidence in the current literature can shed light on this issue. The notion that this type of enclosed incubation could induce the formation of ESH in both studies needs to be further examined.
In terms of the materials used for analysis; some problems with the carboxen-PDMS fibres in the measurements of certain sulfur compounds have been highlighted by previous studies. One such issue involves the decomposition of analytes during sample preparation and subsequent injection into the GC, as shown in the study by Haberhauer- Troyer et al. (1999) where DMS was oxidised to DMSO. In a separate study, Lestremau et al. (2004) found that using the carboxen-PDMS fibre for malodorous sulfur compounds in the analysis of industrial effluent by pulsed-FPD, ESH was completely dimerised into diethyldisulfide. Notably though, no other sulfur compounds in their study were affected by this reaction.
It is very difficult to estimate the sources of error from this limited data set. However, the results from this study are, if anything, more likely to be an underestimation of the concentrations present in the ASS sampled. The fact that they were kept at a constant, ambient temperature out of the light, certainly would have minimised any degradation of DMS (van Bergeijk et al. 2002). However, the degradation of VSCs between sampling and analysis more than likely occurred, meaning the measurements should
Chapter Six: Laboratory-based Sulfur Gas Measurements 176
therefore be considered within the lower limits of that present from undisturbed conditions.
Calibration Issues The limit of detection for ESH is comparable with that measured by Mestres et al. (1999a), using a similar laboratory process along with a cryogenic trap. It is anticipated that the LOD for these experiments could be reduced substantially if certain parameters were altered. This is particularly relevant regarding the temperature of extraction; as an increase would amplify the concentration of volatiles within the container headspace in accordance with Henry’s Law. Also, Abalos et al. (2002) showed that although the addition of NaCl to wastewater samples did not increase the headspace concentration of alkyl-sulfides, they suggested that it could affect the extraction of other analytes by altering the solubility of humic acids. In terms of the experiments conducted in this study though, these alterations are impractical; firstly regarding the use of sediment (i.e. non-aqueous) samples; and secondly because the aim was to identify potential gaseous emissions of such volatiles under natural conditions rather than induced releases. However, these changes may be a way of decreasing the LOD for porewater samples, and/or determining the total concentrations of certain analytes by removing interactions with colloidal phases, for example.
Because the SPME is still a somewhat emerging technique in environmental assessment, its calibration is still not completely error-free. Aside from the assumption that the response of the GC to the analyte concentration is by a factor of n2 (as discussed in section 6.2.2), there is another calibration issue that needs to be considered; the use of external standards. The standard addition technique (using multiple measurements) in conjunction with external standards would be most appropriate for the calculation of the quantities of VSCs present in the samples (Ezquerro et al. 2004; Zeng and Noblet 2002). Alternatively, several internal standards would also have proved a better measure, allowing for the different standards and heterogeneous matrices. However, because the samples were sealed onsite in order to prevent losses of VSCs following extraction from the sediment profile, the approach taken was deemed most appropriate under the circumstances.
Chapter Six: Laboratory-based Sulfur Gas Measurements 177
Absence of SO2 & H2S in Samples
Numerous studies have identified that the measurements of SO2, and particularly H2S, are extremely difficult, in both ambient air samples (Hines and Morrison 1992; Lau 1989), as well as in concentrated headspace analyses (Miracle et al. 2005; Nielsen and Jonsson 2002a). Although extreme care was taken to ensure all gaseous sulfur species remained as was sampled, changes to the chemistry between sampling and analysis undoubtedly occurred. As mentioned previously, the composition of VSCs emitted from coastal sediments is strongly influenced by the redox condition (Devai and DeLaune
1995). This study by Devai and DeLaune (1995) showed that H2S was not detected within the coastal marsh sediments until the redox potential was below -100 mV.
Additionally, the inertness of the sampling containers (glass and PTFE) would suggest that adsorption to the surfaces would not have been primarily responsible for any losses. An alternative possibility is that the detection limits of the GC and column were not sufficiently sensitive to register those components. Again, this is unlikely, as the literature set out for the Shimadzu GC-17A-FPD in conjunction with SPME, states that the detection limits for SO2 and H2S are 17.8 and 9.4 ppbv respectively. Based on the field concentrations measured, as well as the enclosure and enrichment processes, the incubated concentrations of the two compounds should be sufficiently above such detection limits.
Therefore, it is felt that the most likely reason for the inability to detect either SO2 or
H2S was due to transformations of the two gases by changes in the redox chemistry of the samples. H2S was most likely oxidised (to SO2) or dissolved, between sampling and analysis, and any SO2 would most likely have dissolved in the moist samples (as shown in the equations below).
SO2(g) + H2O ↔↔↔ SO2.H2O(l) [6.11] + - SO2.H2O(l) ↔↔↔ H + HSO3 [6.12] - + 2- HSO3 ↔↔↔ H + SO3 [6.13]
The absence of other VSCs is also an interesting point arising from these experiments. The presence of ethanethiol over methanethiol has already been discussed, but also the
Chapter Six: Laboratory-based Sulfur Gas Measurements 178
absence of COS in the samples is worthy of note, as Cutter and Radford-Knoery (1993) measured COS concentrations in excess of 7000 nmol L-1 in anoxic estuarine porewaters.
6.4.6 Combined Interpretations and Conclusions
This is the first time to the author’s knowledge that DMS and ESH have been measured in an agricultural and naturally occurring ASS. The presence of these reduced VSCs in addition to H2S does place a question over the previous chapter’s interpretation of a constant emission of H2S which is photo-oxidised during the daytime to SO2. DMS can also oxidise via the action of free radicals to form SO2, as well as methanesulfonic acid, and minor amounts of dimethyl sulfoxide and sulfuric acid (Ferek et al. 1986; Finlayson-Pitts and Pitts Jr 2000; Fried and Tyndall 1999; Grosjean 1984; Grosjean and Lewis 1982; Hatakeyama et al. 1985; Hatakeyama et al. 1982; Heicklen 1985; Jensen et al. 1992; Murrell et al. 1996; Tyndall and Ravishankara 1991; Wilson and Hirst 1996).
Thiols can also photo-oxidise in the presence of OH and NOx to form SO2 (Grosjean 1984; Wilson and Hirst 1996), or alternatively their microbial oxidation (methanethiol) can produce H2S (Suylen et al. 1987).
The diurnal pattern partially evident in the H2S and SO2 from Blacks Drain has also been observed in DMS concentrations attributed to this photo-oxidative effect, within upland soils in Germany (Staubes et al. 1989), salt marshes in Florida, U.S.A (Cooper et al. 1987a; Cooper et al. 1987b), and rice paddy fields in Japan (Kanda and Minami 1992). Jorgensen and Okholm-Hansen (1985) also found that DMS emissions peaked during the daytime within Danish coastal marsh soils suggesting DMS may also be contributing to SO2 measurements in addition to that from H2S.
The fact that a reduced sulfur compound (ESH) was measured within the oxidised sulfuric layer of the agricultural ASS gives valuable information on the emission measurements of H2S (Chapter 5). This would support the idea of micro-niches allowing for the formation of thiols via the reaction of upwardly diffusing H2S (from bacterial sulfate reduction) with organic matter, a point also noted by Francois (1987). The micro-niches facilitate the localised reduction of sulfate and consumption of Fe2+
Chapter Six: Laboratory-based Sulfur Gas Measurements 179
(within the oxic zone) which would also allow for an export of H2S. This pool of H2S has the ability to interact with humic materials through their oxidation and cleavage to polysulfides to form organic polysulfides and eventually thiol groups (Francois 1987).
Undoubtedly, DMS and ESH are linked within this system as part of the biogeochemical sulfur cycle. Evidence from Kiene and Hines (1995) has shown that the methylation of methanethiol was the major pathway leading to DMS formation in anaerobic (Sphagnum spp. dominated) peat / wetlands. The links between these sulfur gases and other biogeochemical cycles (esp. carbon and nitrogen) also need further exploration. Kiene and Hines (1995) went on to express the idea that the formation of DMS within the low-pH environment was more closely tied to the carbon metabolism than to inorganic sulfur cycling. Preliminary measurements of a diverse range of methanogenic archaebacteria within the ASS at the Blacks Drain study site (Durr et al. 2006), would also support an exploration into the possible biogeochemical linking between S and C cycles.
Chapter Six: Laboratory-based Sulfur Gas Measurements 180
Chapter Seven: DISCUSSION OF COLLECTIVE RESULTS AND CONCLUDING REMARKS
7.1 CONCLUDING REMARKS
7.1.1 Introduction
This thesis brings together novel research aspects on the measurements of VSCs within ASS on the NSW north coast. Several important findings are reiterated within this chapter, in an attempt to create a holistic discussion of the issues involved. As a broad summary of the experimental work, a schematic diagram from both the Blacks Drain (Figure 7.1) and Cudgen Lake (Figure 7.2) study sites have been constructed. These diagrams outline the important (and potentially important) pathways of VSCs transfers within the two ASS locations. Whilst many of these reaction pathways have been previously well established, there are some that require further supporting experimental evidence, and are merely hypotheses at this stage. These points are discussed in more detail below.
7.1.2 Summary
This thesis further demonstrates that the cycling of sulfur within ASS is not a simple process of its reduction from sulfate to iron sulfides, after which it can be oxidised, back to sulfate. But rather, that sulfur is in a complex equilibrium with the other elemental cycles (particularly C and N), of which, the formation (and release) of volatile compounds are an integral part. As such, it is difficult to discuss the cycling of a single element such as sulfur in isolation from others. The interlinking of elemental cycles and the production of VSCs have received little attention within ASS research.
Chapter Seven: Combined Discussion and Conclusions 181
Initial results using passive diffusion samplers confirmed that ASS are a varying source of SO2, with emissions most likely occurring as pulses relating to climatic variables.
The results indicate that there may be some variation in the emission of SO2 from different ASS landuse, although this point still requires further clarification. The use of passive diffusion samplers is clearly more suitable to longer term emission studies, in accordance with their original experimental design. Although, the more concentrated chamber measurements showed that there is the possibility to expand such a sampling regime looking at shorter timescales. The methods also demonstrated the need for careful handling of the samplers and a need for substantial repetitions owing to large variability in the results.
The results of the micrometeorological field studies, showed for the first time that ASS are a source of H2S (in addition to SO2) with flux values in the range of that shown by many highly productive marine and freshwater wetlands and marshes. The data also shows an inverse diurnal fluctuation between H2S and SO2 during parts of the sample periods suggesting a potential photo-oxidative conversion of H2S to SO2 during daylight hours. Analysis of the climatic data shows that the emissions of SO2 and H2S are influenced primarily by two opposing groups of climatic parameters; temperature (soil, air, latent heat and net radiation) and moisture (soil and precipitation). SO2
(concentration and flux), positively correlates to the temperature variables, whilst H2S, positively correlates to the moisture variables. These observations support the photo- oxidative hypothesis as well as displaying congruence with the solubility of the gases and the induced effects of soil porespace saturation.
An interesting aspect to the measured emissions of H2S is that it challenges a pre- existing view that agricultural soils are a sink rather than source of reduced sulfur gases.
Although the same can be said for SO2, it is particularly curious as to the method by which the reduced gas is escaping to the atmosphere. With the uppermost metre of the Blacks Drain site profile being strongly oxidising, it still remains unclear whether the driving force is simply diffusion, or whether there are significant contributions made by micro-niches, macropores / shear planes / bioturbation, and / or bubble ebullition.
The measurements of H2S also have implications regarding our understanding of ASS functioning within such a system. High concentrations of H2S within the soil profile
Chapter Seven: Combined Discussion and Conclusions 182
generally confer that higher concentrations of pyrite are able to form, and that there is abundant organic matter to act as the energy source for the reduction of sulfate (Berner
1970). Therefore, the emission of H2S (along with the CRS measurements of < 1.6%) would indicate that there are insufficient oxidants (especially reactive iron, shown by the high DOP values) within the profile to initiate the conversion of H2S to pyrite.
Excess H2S would also hinder the persistence of iron monosulfides within the profile, as sufficient elemental S and thus pyrite would be plentiful under such conditions. This point is supported by the AVS measurements being below detection limits.
The use of SPME as an extraction and preconcentration step prior to using GC/FPD demonstrates that ASS also have a large capacity to produce other low molecular weight sulfur compounds such as DMS and ESH. Although frequently measured in coastal and marine ecosystems, the measurements of DMS within Australian ASS are novel, especially within agricultural ASS. Although it was beyond the scope of the project, it would appear that the presence of DMS was most likely due to the anaerobic methylation of thiols at both study sites, with a degree of influence of DMSP interconversion within the Cudgen Lake site.
The measurement of ESH within the agricultural ASS is another novel finding in ASS research, and indeed global VSC measurements, primarily because it is not considered one of the dominant naturally occurring thiols. Although present in substantially lower concentrations compared to DMS, its formation could result from the degradation of ethionine (or S-ethyl-cysteine), the catabolism of DMS, or the reaction of ethanol and
H2S. The concentration gradients for the two compounds also suggest a possibility for DMS (at the Cudgen site) and ESH (at the Blacks Drain site) to be released to the atmosphere.
As mentioned above, Figures 7.1 and 7.2 conceptually summarise some of the processes thought to be important in the cycling of sulfur and the production of VSCs at the two study site locations, Blacks Drain and Cudgen Lake, respectively.
Chapter Seven: Combined Discussion and Conclusions 183
BLACKS DRAIN: Agricultural ASS
Potential influence of cane crop
Aerosol formation 2- Photo-oxidation Photo- SO2 SO4 SO2 oxidation Evaporation VSCs of soil H S Ground H2S 2 porewater Surface
SO 2- 3 C2H5O ESH Oxidised ASS H2S Biological Micro-niche Upward diffusion / oxidation 2- H2S formation bubble ebullition / SO4 shear planes CH2O Oxidation S0 Org. S Front DMS Unoxidised H2S H S H S ASS 2 2 FeS FeS2
Processes relating to VSC formation and potential emissions:
Atmospheric emission of H2S derived from depth, within the unoxidised ASS layer
Emission of H2S from oxic soils, derived from localised areas of sulfate reduction around organic matter aggregations, according to; 2- - SO4 + CH2O b H2S + 2HCO3
Formation and potential emission of DMS via the sequential methyl addition to H2S and then methanethiol; R-O-CH3 + H2S b R-OH + CH3SH R-O-CH3 + CH3SH b 2R-OH + CH3SCH3
Formation and potential emission of ESH via the addition of ethanol derived from sugar fermentation and H2S from either the surface layers (as described in process 2 above), or from the anaerobic sediments (link not shown); 2C2H5OH + 2H2S + ½O2 b 2CH2H5SH + 3H2O
Note: Dashed lines represent uncertain pathways.
Figure 7.1. Schematic summary of the processes important in the formation and release of VSCs from agricultural ASS.
Chapter Seven: Combined Discussion and Conclusions 184
CUDGEN LAKE: Undisturbed ASS
Δ Cloud albedo
Sulfate aerosols
SO2
DMS / H2S? H2S? Marine Sediment Input Surface DMSO DMSP
Oxidised Zone
2- Upward S2O3 Oxidation FeS S0 diffusion Fibrous /bubble 2- SO4 ebullition / Reducing Downward shear Zone diffusion planes
Strongly DMS MMPA Reducing H2S Zone FeS2 pH > 7
H2S + CH4 MSH 3-MPA H2S
Processes relating to VSC formation and potential gaseous emissions within Cudgen Lake:
In anoxic sediments, DMSP can either be cleaved to form DMS and acrylate , or demethylated to MMPA (which subsequently can form DMS, MSH or 3-MPA and H2S). See Chapter 2 for equation stoichiometries.
Formation of H2S and DMS (also through reduction of DMSO) could potentially lead to emission of the gaseous forms , which alter global climate by contributing to changes in cloud albedo.
Alternatively, these reduced compounds have the ability both in their reduced and oxidised form to contribute to enhanced pyrite and monosulfide formation
Note: Dashed lines represent uncertain pathways.
Figure 7.2. Schematic summary of the processes important in the formation and release of VSCs from undisturbed ASS.
Chapter Seven: Combined Discussion and Conclusions 185
7.2 IMPLICATIONS FOR COASTAL LOWLAND ACID SULFATE SOILS
Although there is significant dispute as to the global estimates of ASS distribution, they are not all that extensive. Nevertheless, they are frequently located in sensitive ecological areas, particularly in the tropics, which are exhibiting rapid population expansions (Ritsema et al. 2000). Therefore, it is worth looking at the potential implications and impacts of VSCs from ASS, particularly regarding the context in which the measurements were made.
7.2.1 Are Concentrations of Volatile Sulfur Compounds an Issue in Acid Sulfate Soils?
Although the concentrations and fluxes of SO2 and H2S measured coming from ASS would suggest only a limited impact in terms of global sulfur budgets, they may be important on a regional or local scale as suggested in other small scale S gas emission studies (Steudler and Peterson 1984). This is especially the case where ASS occupy large fractions of the overall land areas in specific regions, such as in coastal areas within the tropics. Therefore, ASS may have the potential to affect regional atmospheric chemistry and sulfur systems in a manner similar to that identified by Nriagu et al. (1987), who found that the release of DMS from boreal wetlands (14 % of total land area) accounted for 30 % of the acidifying atmospheric sulfur burden in remote areas of wetlands. Similarly, Fletcher (1989) found the oxidation products of DMS could account for 30 - 50 % of the total sulfur acids in Scandinavian air during the summer phytoplankton blooms. It should be noted that on a global scale though, the oceans contribute a far greater proportion of sulfur compounds compared to coastal wetlands (~ 2 %). An anecdotal observation from Prof M. Lou (pers. comm.), President of South China Normal Univ., Guangzhou, is that one area in southern coastal China that has significant acid rain occurrence is not nearby any large industrial area or in the wind fetch of such an area.
It should be acknowledged that a major problem in calculating regional budgets for individual gases or sulfur emissions as a whole, is the inherent uncertainty of results
Chapter Seven: Combined Discussion and Conclusions 186
derived from the extrapolation of limited field data, and their relation to counterparts on even a catchment-wide basis. Of particular concern, as highlighted by this research, is the diurnal variation in SO2 and H2S concentrations, let alone potential seasonality impacts that the sampling regime unfortunately could not consider. For example, see the results presented by Bodenbender et al. (1999), which showed a distinct summer increase in emission of gaseous sulfur compounds. Additionally, the SO2 and H2S measurements in my study were made only over fallow or newly growing cane. Undoubtedly, changes in emissions would occur with the establishment of a canopy, meaning these figures would be non-representative of total crop cycles for sugarcane overlying ASS. This final point though is somewhat debatable (at least in terms of H2S flux), as seasonal changes also bring with it alterations to concentrations of free radicals. Crutzen and Fishman (1977) showed that OH levels in the troposphere increase approximately 3-fold during summer, which would have the potential to balance a 5- fold increase in H2S flux. Additionally, sulfur gases can be substantially dispersed in the atmosphere (Rodhe 1972), meaning deposition can occur well beyond the source.
It could be assumed that this level (for H2S) would represent an above-average limit for this particular area, as both sampling periods were undertaken during late spring / early summer when daily temperatures, and therefore bacterial productivity, are higher than in winter months (Ingvorsen and Jorgensen 1982). Although seasonal changes are particularly evident in cooler-climate locations, they have also been shown to effect reduced S-gas emissions from sub-tropical areas; e.g. Louisiana, see DeLaune et al. (2002).
Another factor limiting the assessment of the possible role of VSCs in the global S cycle is the definition of the area of ASS. Global ASS area values published by some (Brinkman 1982) appear to be based upon earlier soil mapping with only about 12- 13 Mha indicated. Such early assessments showed zero area of ASS in Australia, but a recent assessment by Fitzpatrick et al. (2006) suggests there are > 9 Mha in coastal Australia alone. Obviously the problem here is one of definition of ASS and also that of the method of their assessment. Fitzpatrick et al. (2006) have used remote sensing and geomorphic models in their assessment but they indicate this value may need to be increased to incorporate non-coastal ASS. A value of 100-200 Mha globally, of ASS and potential ASS (PASS) has been made (Dent 1986), but the ASS Working Group of
Chapter Seven: Combined Discussion and Conclusions 187
the International Society of Soil Science is encouraging Fitzpatrick’s assessment techniques to provide a new value for global ASS.
Another issue in assessing global VSCs emissions is the uncertainty of how ASS of different characteristics behave under various landuse management regimes. It is already clear that the global population is concentrated close to coastlines where ASS often occur, and any problems will be accentuated with further population growth. Indeed, the health impact of long-term exposure to naturally occurring trace sulfur gases on human populations is an area that needs further exploration (see below).
7.2.2 Vegetation / Animal Impacts
SO2
Assessments on the phytotoxicity impacts of SO2 concentrations (both directly and indirectly) have been in decline since the stabilisation and decrease of western anthropogenic emissions since the 1980s. However, with the more recent increase in emissions from developing nations (primarily China and India) resulting from rapid industrialisation and greater demand for energy and transport, this area of research may escalate in the near future; see Agrawal (2005) and Yang et al. (2002). Nonetheless, a multitude of research papers have shown variable effects of simulated SO2 concentrations on plant yield, leaf weight, grain weight, etc. (Adaros et al. 1991; Ashmore et al. 1988; De Temmerman et al. 1992; Fowler et al. 1988; Weigel et al.
1987). However, the majority of studies have used mixtures of pollutants (NO2, O3), often resulting in the increased prevalence of adverse effects (WHO 2000). This makes the interpretation and any generalisations of threshold values difficult to put into this context.
Some interesting points to draw from the literature include, that lichens have been found to be sensitive to atmospheric SO2, with early studies showing levels of < 0.1 ppm caused growth inhibition (Baddeley et al. 1973; Richardson and Puckett 1973; Saunders and Wood 1973). More recent work suggest this value is much lower, see review article by Nash and Gries (2002). However, lichens and mosses do not have a protective cuticle, making them extremely sensitive to ambient SO2 concentrations. Therefore, this
Chapter Seven: Combined Discussion and Conclusions 188
is less of an issue in the case of sugarcane, and other higher plants. However, there is some evidence that levels of SO2 (yearly means of between 25 and 830 ppb from an industrial source) can also reduce total crop volume (Garcia-Huidobro et al. 2001).
From an indirect perspective, chronic, low-level exposures can also result in an increased plant susceptibility to other biotic and abiotic stresses including drought, pathogen attack, and frosts (Garcia-Huidobro et al. 2001). This final category is particularly relevant to the NSW sugar industry, as frost absence is one of the primary determinants for crop location within NSW since the tropical crop is growing towards its southernmost climatic limit. It is interesting to note a study by Baker et al. (1982), which demonstrated that levels of SO2 of between 19.25 - 56 ppb resulted in a wheat crop suffering more frost damage than those that were SO2-free. It was speculated that this was due to SO2 weakening membranes or reducing photosynthetic rates / increasing respiratory rates (Baker et al. 1982). Also regarding SO2 concentrations specifically, preliminary studies suggest that increased concentrations can result in the stimulation of pest populations such as aphids and fungal pathogens, potentially when synergising with certain nitrogen gases (Bell et al. 1993), which have been measured at the Blacks
Drain study site (Denmead et al. 2005). Also, the wash-out of gaseous SO2 can result in localised acid rain problems. However, the tolerance of sugarcane to extremely acidic soil conditions, would suggest levels of acid-rain deposition generated from SO2 would be inconsequential.
The concentrations of SO2 measured at Blacks Drain (< 3 ppb) are clearly not at a level that would be considered dangerous in any way towards primary plant functioning, with natural background concentrations in remote Australian locations averaging 0.3 ppb (Ferrari and Salisbury 1999). Adapting research from van der Eerden et al. (1988) would suggest that the SO2 concentrations measured from ASS would result in an approximate reduction in crop volume of < 0.36 %. This almost-negligible decrease is complicated by any beneficial effects observed by a concurrent decrease in pest populations in an elevated SO2 environment. It would be anticipated that native plant species, although being more susceptible to pollutants compared to most crops, would be less affected by any natural SO2 concentrations from ASS, owing to their co- evolution over long periods of time. Nevertheless, further monitoring of SO2 concentrations would be beneficial, as even low-level SO2 exposure can result in subtle
Chapter Seven: Combined Discussion and Conclusions 189
physiological changes to the crop, such as minor reductions in yield or sugar content, or premature senescence that might not be currently accounted for.
H2S
From an ecological perspective, in dissolved and gaseous form H2S is both directly and indirectly toxic by reducing sulfur availability to higher plants (Ponnamperuma 1972), such as sugarcane, as well as being highly toxic towards nematodes and to many fishes (Bremner and Steele 1978). This last point has particular relevance to the low-pH ASS environments, as the toxicity of H2S is pH dependent. H2S, the dominant sulfide form below pH 7, can readily diffuse across the cell membrane of animals causing maximum toxicity in acidic environments (Wang and Chapman 1999). Another issue regarding sulfide toxicity, is that it also has the potential to lock up and subsequently release trace metals, which have been shown to have substantial impacts on downstream ecosystems when oxidised (Morse 1994a; b). As is the case with SO2, low concentrations of H2S may be beneficial to crops grown in sulfur-limiting locations (Durenkamp and De Kok 2002).
Other VSCs In higher concentrations, thiols are considerably more toxic than their corresponding disulfides, with toxicity increasing with chain length (Sipma et al. 2004). Inoue et al. (1955); cited in Alexander (1974), found that ethanethiol inhibited the ripening of rice plants. Whilst this and the above sections have focussed on the negative impacts of the sulfur gases on vegetation (as weighted in the literature), it should be emphasised that atmospheric S can also have a beneficial effect, particularly for crops on sulfur deficient soils, at certain concentrations (Noggle et al. 1986). Levels of some VSCs may provide some protective benefits towards the sugarcane crop in the correct amounts. Lewis and Papavizas (1970) found that the addition of substances releasing VSCs (MSH, DMS, DMDS) controlled root rot of some crops (peas) caused by soil-borne phycomycetous pathogens. Conversely, the same compounds were also found to be highly toxic towards the germination of fungi (Lewis and Papavizas 1971), and in larger concentrations, MSH has been suggested to be phytotoxic (Alexander 1974).
Chapter Seven: Combined Discussion and Conclusions 190
7.2.3 Human Health Impacts
With around 80% of Australians living in coastal zones, of which 66% are concentrated around large urban centres on estuaries and inlets (White et al. 2006), there is an increasing possibility of coastal ASS imposing a greater level of social and economic cost on the population. One of these areas may be the impact of VSCs on human health, or alternatively there is also the potential for the developed soils to cause annoyance, as most VSCs have a range of clearly identifiable and objectionable odours. Additionally, they have a low odour threshold limit making even trace concentrations a nuisance to those close by. All VSCs impact detrimentally on human health to varying degrees, see Table 7.1 for a summary of the exposure effects and odour thresholds.
Table 7.1. VSC odour thresholds, short-term and long-term exposure effects. Sourced from (Kilburn 2003; Li and Shooter 2004; WHO 2000). Odour Low Level Acute Low Level Chronic VSC Characteristic Exposure Effects Exposure Effects† (Threshold)
Pungent Respiratory irritation, eye SO 2 (1-5 ppm) soreness
Possible permanent CNS Rotten egg Headache, dizziness, H S damage, increased coronary 2 (0.0005-0.3 ppm) respiratory irritation, nausea disease
Decaying DMS vegetables Slight respiratory irritation (<0.1 ppm)* Decaying cabbage / Headache, dizziness, ESH onion CNS damage? respiratory irritation, nausea (0.0003-0.001ppm)
Pungent / cabbage Respiratory irritation, MSH CNS damage? Lung edema (0.001 ppm) headache, nausea
* Level as yet not established / Estimate only † Not conclusively established
Whilst a great number of studies have documented the adverse health effects of atmospheric sulfur compounds, there is a distinct lack of information on the impact of long-term, low exposure to VSCs, or indeed most volatile compounds, primarily because of the near impossibility to separate confounding factors. In light of the literature available though, the atmospheric concentrations of SO2 and H2S measured at Blacks Drain are considered below.
Chapter Seven: Combined Discussion and Conclusions 191
SO2
In the case of SO2, the vast majority of studies have been within urban situations (Dab et al. 1996; Kunzli et al. 2000; Pandey et al. 2005; Pope et al. 1995), primarily because of its link with anthropogenic sources and synergistic effects with other pollutants (Samet et al. 2000; WHO 2000). The limited research regarding health monitoring of environmental SO2 emissions have focussed around geothermal / volcanic areas (Bates et al. 1998; Hansell and Oppenheimer 2004). Regardless of the source, there has been a clear link established between long term exposure to relatively low-levels of SO2 and an increased incidence of asthma and bronchitis (Stjernberg et al. 1985), and potentially to specific types of cancer (Amaral et al. 2006). The current (2006) Australian guidelines as identified by the National Environment Protection Council allow for a yearly average
SO2 exposure concentration of 20 ppb. Fortunately though, with maximum levels of
SO2 measured at Blacks Drain being < 3 ppb during both 2003 and 2005, the impact on human health, even considering long-term exposure, would be insignificant. The levels measured would also indicate that SO2 emissions would not even be an odour nuisance within the measured ASS. Nevertheless, the sulfur gas research initiative at the Tweed to pursue van Breemen’s hypothesis of SO2 emission from ASS, was in partial response to believing it was smelled in ASS areas after rain events.
H2S & Other VSCs
Again, as is the case with SO2, the majority of the literature is focussed on the potential health impacts of H2S on urban areas from wastewater (Burgess et al. 2001; Devai and DeLaune 1999; 2000; Muezzinoglu 2003) and pulp mills (Jaakkola et al. 1999). Studies from environmental sources are also limited to volcanic / geothermal areas (Bates et al. 2002; Hansell and Oppenheimer 2004). Again, regardless of the nature of the emissions, potential links between exposure to low-concentrations of H2S and neurological behaviour (Inserra et al. 2004), including an increased incidence of depression (Saadat et al. 2006), as well as respiratory and blood disorders (Legator et al. 2001), have been previously documented. Even low-dose chronic exposure has been shown to result in impaired balance, slowed reaction time and verbal recall (Kilburn 1997). See
Beauchamp et al. (1984) for a comprehensive review of the toxicity of H2S.
Chapter Seven: Combined Discussion and Conclusions 192
The data from Blacks Drain showed that the maximum concentration of H2S was in the order of 25 ppb in 2003 and 1 ppb in 2005. As opposed to the concentrations of SO2 measured, these numbers are more noteworthy. One study from a site in Finland (discussed in more detail below), showed that this concentration falls into what could be considered moderate exposure levels, with the study demonstrating that daily levels between 6.7 ppb and 20.1 ppb (two-thirds of which was H2S) resulted in a significantly greater incidence of eye and respiratory symptoms, as well as nausea (Marttila et al.
1995). One of the few examples of long-term H2S exposure comes from Rotorua in New Zealand, where the data suggests a higher incidence of nervous system and eye disorders from the elevated levels of geothermally-derived H2S (Bates et al. 1998). This particular study relied only on hospital discharge data and made no measurements of the sulfur gas concentrations, but previous research has suggested that levels at Rotorua are far greater than the measurements at the Blacks Drain site.
This last point highlights the issue that research into this area is by no means conclusive in terms of the levels of long-term, low-concentration environmental exposure to H2S and their health impacts, particularly with establishing confounding factors (Kilburn and
Warshaw 1995). Another problem is that the atmospheric lifetime of H2S is relatively short, implying that exposure would need to take place close to the source. However, without attempting to be alarmist in any way, there is clear scope to extend the monitoring of H2S in coastal ASS to establish whether the concentrations are continually exceeding levels which have previously been shown to cause adverse human effects. This would particularly concern those who spend large portions of the day working with ASS such as the sugarcane farmers.
Studies on the health effects of emissions of other VSCs aside from SO2 and H2S are relatively scarce. The site mentioned previously in Finland (South Karelia), has extensively examined impacts on the local population of emissions of total reduced sulfides (H2S, MSH, DMS and DMDS) from a paper mill. One study within this area found that average concentrations of approximately 2 ppb (max. 103.9 ppb), resulted in an increased incidence of various symptoms, including coughing, respiratory infections and headaches in adults and, although not significantly, in children (Marttila et al. 1994; Partti-Pellinen et al. 1996). The measurements made at Blacks Drain do not include
Chapter Seven: Combined Discussion and Conclusions 193
atmospheric concentrations of these gases (DMS and ESH), so the potential health effects cannot be referenced.
Regardless of the potential health impacts posed by VSC emissions from ASS, there is a case for also looking at the potential nuisance factor created, with peak emissions of
H2S easily exceeding its odour threshold. As these findings are novel to the area of ASS research, the implication of H2S as an atmospheric nuisance is an unexplored idea with potentially large ramifications for the expanding urban development of Australia’s coastal locations. There have been a number of complaints to Council from Tweed Shire ratepayers concerned with the smell of rotten egg gas after certain rainfall / flood events near some Tweed River areas (C. Easton, TSC Entomologist, pers. comm.). Botany residents have also complained to the local council about this odour from the drain entering Botany Bay to the East of the Sydney Airport 3rd runway, and following the construction of that facility. It seems with the latter problem, that excessive organic inputs from weed control in freshwater wetlands upstream on the drain have combined with sulfate from Botany Bay in the poorly flushed drain environment to give greatly increased H2S production (M. Melville, UNSW, pers. comm.).
Along a similar line of thought is the possibility that acidic sulfur gases generated from ASS, could contribute to the corrosion of urban infrastructure. ASS have been demonstrated to result in corrosive impacts primarily through the action of the acidity generated by oxidation (Dent 1986). However, as suggested by Devai and DeLaune (1999; 2000) within the wastewater and sewage treatment industry, gaseous sulfur compounds may also contribute to corrosive effects, particularly on concrete. Although this would not be considered to occur to the same extent, it could be a potential impact worth exploring in the development of ASS for urban expansion.
7.2.4 Atmospheric / Climatic Impacts
Aerosols (including sulfate aerosols) affect the global climatic balance by changing the characteristics of clouds in many ways, including: acting as cloud condensation and ice nuclei thus altering cloud albedo; and inhibiting freezing and potentially influencing the hydrological cycle (Andreae and Crutzen 1997; Charlson et al. 1987; Lohmann and
Chapter Seven: Combined Discussion and Conclusions 194
Feichter 2005). Berresheim et al. (1995) indicate that the role of clouds in the atmospheric radiation budget is one of the largest uncertainties in the prediction of future climate change, and the role that S gases play is believed to be important because of its strong association with cloud formation via the production of cloud condensation nuclei’s (CCN) (Bates et al. 1987). Indeed, this point has yet to be fully resolved, with the magnitude of the impacts highly dependent on the model used (Myhre et al. 1998; Penner 2001) and the season in which the measurements are made (Ayers and Gras 1991). Only a minority suggest that the effects of DMS and other sulfur gases, including anthropogenic SO2 emissions, have no impact on cloud albedo or mean surface temperatures (Schwartz 1988).
Due to the dominance of oceanic DMS emissions (> 95 %) compared to terrestrial sources, based on figures from Andreae and Jaeschke (1992), and Kettle and Andreae (2000), the impact of any significant global changes in the creation of CCN due to DMS emissions from ASS is expected to be non-existent. Even more so in the case of CCN formation from SO2, the levels emitted from anthropogenic and even volcanic sources far exceed any other terrestrial source such as emissions from ASS.
Nonetheless, the contribution of ASS to the creation of sulfate aerosols, although limited in a global sense, still warrants examination of budgeting from a local and possibly even a regional scale. Because of the complexities of interrelationships between cloud formation and CCN (Ayers and Gillett 2000), the relationships between DMS and climatic changes are still yet to be fully resolved. Studies which enhance our understanding of the sources of CCN precursors are nonetheless valuable in quantifying global risk assessment.
7.2.5 Can the emissions from ASS be managed / decreased?
From an agricultural perspective, since the amounts of sulfur gases being emitted from sugarcane under ASS are negligible compared to other gases (nitrogen and carbon, see Denmead et al. (2005; 2006)), there is little need for a particular S gas management plan. Indeed, focus on the rate of fertiliser application at this study site is currently being investigated (Denmead et al. 2005), as the emission of nitrogen oxides are an
Chapter Seven: Combined Discussion and Conclusions 195
important contributor to the regional greenhouse gas budgets. From the perspective of the cane farmers, the application of N fertilisers is their largest chemical input, and therefore is an expensive process. As such, farmers are keen to reduce any over-usage.
The results linking decreases in SO2 concentrations to saturation levels and watertable heights within the profile are certainly not an excuse to reflood the landscape in an attempt to reduce emissions. It is certainly true that iron sulfides are stable when kept beneath the watertable. This makes the notion of reflooding an attractive option when not in conflict with economic, and moreover, associated social costs. The rehabilitation of wetlands by reflooding is therefore an often discussed management technique. However, as clearly outlined by White et al. (2006; 1997), there are several misconceptions to the science underlying this theory, making the potential for reflooding applicable to only a small minority of cases. The changes in VSCs that would most likely result from reflooding support the already identified limitations of the technique. Whilst a reduction in SO2 emissions may be observed by reflooding, owing to its high solubility, there would most certainly be at least a concomitant increase in reduced sulfur compounds being formed and emitted. The results from this work show that DMS is being produced in large quantities beneath the watertable at both study sites. An increased watertable elevation would stop any microbial oxidation activity, switching the upper profile to anaerobic conditions. This would mean concentrations of DMS would be generated much closer to the air-surface boundary, allowing for the substantial gaseous emissions of the compound. The impact on releases of other volatile compounds also needs considering. The initial shift to anaerobic conditions would likely result in the emission of large quantities of N2O in a manner similar to that observed during other flooding events (Christensen and Tiedje 1990; Hawkins and
Freeman 1994). The impact on CH4 emissions is less definite, with initial rewetting decreasing atmospheric emissions (Freeman et al. 1993), likely due to sulfate-reduction preferences (Freeman et al. 1994), against the fact that wetlands contribute one-fifth of global methane emissions (Cicerone and Oremland 1988). These principles may also be applied to coastal ASS under pressure from global warming; where it has previously been suggested (Gorham 1991) that sea-level rises may have a greater influence on gas emissions than the direct effect of increased temperatures.
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7.3 FURTHER RESEARCH
As with most environmental-based scientific research studies, nearly as many questions are raised as answers provided. Additionally, because of the novel findings of the research here, there are a number of investigative avenues worth pursuing.
It still remains to be determined how the levels of measured H2S are being emitted to the atmosphere when the zone of reduction is so far from the sediment surface. The use 34 32 of S/ S measurements would allow an accurate determination of the SO2 source within ASS, in a manner similar to that used by Hitchcock and Black (1984). This could be further enhanced by using the radioactive isotope of sulfur (S35), which would enable determination of the solid phase transformations and modelling of the sulfur cycling.
Regarding other VSC emissions, the question needs to be resolved as to whether DMS and ESH are actually being released to the atmosphere. An obvious first step would be to use the current methodology in a field-based program with the use of chambers. Unfortunately, the GC setup is not readily transportable, requiring significant power and gas resources. Therefore, the use of tedlar gas sample bags or preferably fused-silica- lined stainless steel canisters for the in-situ sampling of chambers followed by their transportation back to the laboratory may be preferable.
A second question posed by the thesis results is whether ASS emit larger molecular weight S compounds that are most likely being created with the diagenesis of organic matter in ASS, subsequent to the formation of H2S. They are of importance within undisturbed systems because of their close relationship with the formation and degradation of living biomass (Aneja and Cooper 1989). Included in this list would be larger thiols, disulfides (the oxidation products of thiols), and also CS2 which as mentioned previously has been widely measured in anaerobic and waterlogged soils.
CS2 emission studies from ASS would be particularly interesting as they appear to be influenced by not only the organic S content of the soil (e.g. cysteine) but also the inorganic sulfur species, thiosulfate and tetrathionate (Minami and Fukushi 1981a), which are integral products formed during the oxidation of iron sulfides. A possible alternative to the field-trapping of compounds before analysis has been described by Toda et al. (2004), who describes a field-based, portable method using a membrane-
Chapter Seven: Combined Discussion and Conclusions 197
based collection and conductometric and fluorescence analysis for the measurement of sub-ppb SO2, H2S and MSH.
The interaction between the high levels of DMS measured in the agricultural ASS, and sugarcane S assimilation is also an interesting question. One possibility is that the DMS concentrations are linked to the DMSP production within sugarcane, either as an osmolyte or a potential storage compound for excess sulfur. As it is expected that the rhizosphere bacterial community will influence the plant nutritional uptake of sulfur (and indeed other nutrients), further research into the rhizosphere-sediment interface and the possibility of bacterial symbionts would help in the understanding of this issue. This idea could be explored with the employment of peepers, with their high spatial resolution, to measure porewater VSCs. However, this would require instantaneous analysis of the reduced compounds, preferably using gas chromatography.
Using SPME in conjunction with peepers to extract VSCs is a novel approach to examining volatiles with ASS porewaters. This approach would be more suitable being used in conjunction with GCMS, and there would be the further complication of sample preservation between extraction and analysis. However, initially it could be used to look at the binding of thiols and heavy metals, and their impacts on the geochemistry; particularly the sedimentary sulfide phases of ASS. Shea and MacCrehan (1988) found that thiols can enhance CuS solubility in porewaters, whilst having no effect on the solubilities of CdS or PbS.
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LIST OF APPENDICES
All appendices are either Microsoft Excel (2003 .xls) or Microsoft Word (2003 .doc) files, located on the CD-ROM contained at the back of the thesis, under the following directories;
APPENDIX A – Chapter 3 Data /Blacks Drain 2003; Peeper Data.xls /Blacks Drain 2003; Soil Data.xls /Blacks Drain 2004; Soil Data.xls /Blacks Drain 2005; Soil Data.xls /Cudgen 2002; Peeper Data.xls /Cudgen 2002; Soil Data.xls /Cudgen 2004; Soil Data.xls
APPENDIX B – Chapter 4 Data /Ferm Tube Data 2002.xls /Ferm Tube Data 2003.xls
APPENDIX C – Chapter 5 Data /Blacks Drain 2003; Gas Measurements.xls /Blacks Drain 2005; Gas Measurements.xls /Blacks Drain 2005; Micromet Data.xls
APPENDIX D – Chapter 6 Data /Calibration Chromatographs; DMS & ESH.doc /Calibration Chromatographs; DMS.doc /Calibration Chromatographs; ESH.doc /Calibration Chromatographs; SO2 & H2S.doc /Gas Chromatography Data.xls /Soil Sample Chromatographs.doc
List of Appendices 239