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The geochemistry of Fe, Mn, Ni, Cu, Zn and As in the water column, sediments and porewaters in a seasonally.. Martin, Alan 1996

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THE GEOCHEMISTRY OF Fe, Mn, Ni, Cu, Zn A N D As IN THE WATER C O L U M N , SEDIMENTS A N D POREWATERS IN A SEASONALLY ANOXIC L A K E  by  A L A N MARTIN  B.Sc, The University of British Columbia, 1991  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER V OF SCIENCE in THE FACULTY OF GRADUATE STUDIES (Department of Earth and Ocean Sciences)  We accept this thesis as conforming to the required standard  THE UNIVERSITY OF BRITISH COLUMBIA December 1996 © Alan Martin, 1996  In  presenting  degree freely  at  this  the  available  copying  of  department publication  of  in  partial  fulfilment  University  of  British  Columbia,  for  this or  thesis  reference  thesis by  this  for  his thesis  and  study.  scholarly  or . her for  of I  I further  purposes  gain  shall  requirements  agree  that  agree  may  representatives.  financial  the  be  It not  that  is be  of  £^rr#7  CH^/  Date  DE-6  (2/88)  XfM/  by  understood allowed  Oc&2tn ?>0<£<s>rr*<  The University of British Columbia Vancouver, Canada  Library  an  advanced  shall  permission for  granted  permission.  Department  the  for  the  make  extensive  head  that  without  it  of  copying  my or  my _ written  ABSTRACT  The distributions of Fe, Mn, Ni, Cu, Zn and As in the water column, interstitial waters and associated solid phases in Balmer Lake, Ontario, were determined from samples collected in July and October of 1993, and March and May of 1994, in order to assess the seasonal biogeochemical controls governing trace metal behaviour and mobility. The basin has served as a repository for tailings pond effluents since 1967, and as a result, hosts elevated levels of contaminants in the sediments and lake waters. During the ice-free periods of summer, fall and spring, the water column is characterized by fully oxygenated bottom waters and homogeneous distributions of all measured parameters. However, reducing conditions develop in deeper areas during the period of ice cover in response to the high biological demand of the organic-rich sediments coupled with restricted atmospheric exchange. Trace metal profiles collected during winter exhibit considerable lakewide variation and appear to reflect variability in the duration and extent of bottom water anoxia, and the relative influence of metal-rich lateral inputs from the tailings circuits. Solid-phase profiles indicate that the top decimetre of the sediment column has received variable contributions of both organic matter- and feldsparrich natural detritus, and carbonate-, chlorite- and metal-rich tailings inputs. High resolution profiles of porewater  N O 3 " ,  Fe, Mn and S O 4 " illustrate that the 2  sediments in Balmer Lake become anoxic within a few centimetres of the sediment-water interface during well-mixed periods.  Evidence for the  precipitation of a Cu-bearing sulphide phase in the lowermost winter bottom waters suggests that the sulphate redox-cline migrates above the sediment-water interface at some point during ice cover. The seasonal porewater distributions of  ii  dissolved Ni, C u and Zn exhibit pronounced consumption profiles at depths consistent with sulphate reduction zones, suggesting they are sequestered as metal-sulphide phases.  Enrichments of dissolved N i and Zn in the surficial  sediments at most sites reflect the dissolution of labile particulates at the sediment-water interface.  Diffusive influxes calculated for Zn, N i and C u  suggest that diffusion mechanisms contribute insignificantly to the accumulation rates of these metals. Arsenic is remobilized at deeper sediment depths and appears to be largely governed by the redox geochemistry of Mn. The data collectively demonstrate that trace metal mobility in the Balmer Lake water column/sediment system varies significantly over the course of the four seasons. With the exception of arsenic, however, dissolved metal fluxes indicate that the underlying contaminated sediments are providing a significant and permanent sink for dissolved metals.  iii  TABLE OF CONTENTS  ABSTRACT  ii  TABLE OF CONTENTS  iv  LIST OF FIGURES  .7.  viii  LIST OF TABLES  xiv  ACKNOWLEDGEMENTS  xvi  I. INTRODUCTION  1  1.1 Trace Metal Geochemistry  1  1.2 Trace Metal Cycling in Lakes  3  1.2.1 Inputs of Trace Elements to Lacustrine Systems  4  1.2.2 Major Processes Governing Trace Metal Behaviour and Mobility  5  1.2.3 Transformations Across the Sediment-Water Interface 1.3 Contamination of Aquatic Systems From Metal Mining 1.3.1 Bioavailability and Toxicity 1.4 Research Objectives  10 13 14 15  II. STUDY SITE  18  2.1 Physiography, Physical and Chemical Limnology  18  2.2 History of Mine-Related Inputs  20  2.2.1 Campbell Mine  21  iv  2.2.2 Arthur White Mine  22  2.3 Sampling Periods  23  2.4 Sampling Sites  '.  III. SAMPLING A N D METHODS  24  26  3.1 Sampling  26  3.1.1 Water Column  26  3.1.2 Physical Profiling  27  3.1.3 Porewaters  27  3.1.3.1 Peeper Description and Preparation  29  3.1.3.2 Peeper Deployment and Subsampling.  32  3.1.4 Sediment Sampling  '.  34  3.2 Instrumentation  35  3.2.1 ICP/MS  35  3.2.2 GFAAS  36  3.2.3 Ion Chromatography  36  3.2.4 X-ray Fluorescence  36  3.2.5 Oculometry  37  3.2.6 Carbon, Nitrogen and Sulphur  37  3.2.7 Spectrophotometry  37  3.2.8 Digestion of Polycarbonate Filters  38  IV. RESULTS  39  4.1 Water Column  39  4.1.1 Physical Profiling  39  4.1.2 Nitrate, Ammonium and Sulphate  42  4.1.3 Suspended Organics  53  v  4.1.4 Trace Metals  57  4.1.4.1 Dissolved Fraction  57  4.1.4.2 Particulate Fraction  64  4.2 Interstitial Waters  69  4.2.1 Peepers  69  4.2.1.1 Nutrients and Sulphate  69  4.2.1.2 Trace Metals  79  4.2.1.3 p H  95  4.2.2 Core Porewaters  97  4.2.2.1 Nutrients and Sulphate  97  4.2.2.2 Trace Metals  99  4.3 Sediments  101  V. DISCUSSION  109  5.1 Water Column  109  5.1.1 Physical Limnology  Ill  5.1.2 Sulphate  Ill  5.1.3 Ammonium and Nitrate  113  5.1.4 Suspended Particulates  117  5.1.5 Trace Metals  121  5.1.5.1 Iron  122  5.1.5.2 Manganese  124  5.1.5.3 Arsenic  126  5.1.5.4 Nickel  127  5.1.5.5 Copper  129'  5.1.5.6 Zinc  130  5.1.6 Water Column Model  vi  131  5.1.7 Comparison of Four Lake Zones  135  5.2 Sediment Geochemistry  142  5.3 Porewater Chemistry  147  5.3.1 Ammonium and Nitrate  149  5.3.2 Sulphate  153  5.3.3 Iron and Manganese  157  5.3.4 Nickel  169  5.3.5 Copper  174  5.3.6 Zinc  179  5.4 Factors Controlling the Diagenetic Behaviour of As  186  5.4.1 Experimental and Field Observations  186  5.4.2 The Behaviour of As in Balmer Lake Sediments  194  VI. SUMMARY A N D CONCLUSIONS  202  VII. BIBLIOGRAPHY  205  VIII. APPENDICES  220  Appendix A. Typical ICP/MS operating conditions  220  Appendix B. Quality Assessment/Quality Control  221  Appendix C. Porewater Data  225  Appendix D. Sediment Data  235  Appendix E. Core Logs  237  Appendix F. Diffusive Fluxes  239  Appendix G. Accumulation Rates  243  vii  LIST OF FIGURES Fig. 2.1 Location map showing the sampling stations, lake bathymetry and tailings <r' ^  circuits of the adjacent mines.  p  Figure 3.1. Schematic diagram of membrane dialysis sampler (peeper).  Pfl- - ^ °  Fig. 4.1 Summer water column profiles of temperature, dissolved oxygen and pH for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, June, 1993. The hatched line f>f • 4*°  represents the sediment surface.  Fig. 4.2 Fall water column profiles of temperature and dissolved oxygen for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, October, 1993. The hatched line represents the sediment surface. Fig. 4.3 Winter water column profiles of temperature and dissolved oxygen for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, March, 1994. The hatched line represents the sediment surface.  ^'  Fig. 4.4 Spring water column profiles of temperature, dissolved oxygen and pH for stations 1, 2 and 4 (A-C, respectively), Balmer Lake, May, 1994. The hatched line represents the sediment surface. Fig. 4.5.  Pf • ^ ¥  Winter water column profiles of nitrate for stations 1, 2, 3 and 4 (A-D,  respectively), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface. p<j .sv Fig. 4.6. Winter water column profiles of ammonium for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface.  57  Fig. 4.7. Winter water column profiles of sulphate for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface.^. £2  viii  Fig. 4.8. Winter water column profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 1, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface.  ^  Fig. 4.9. Winter water column profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 2, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface.  Pft' ° &  Fig. 4.10. Winter water column profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 3, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface. Fig. 4.11.  t'  *  p  <  1  Winter water column profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 4, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface. Fig.  P%- ^  4.12. Duplicate summer peeper profiles of dissolved NH4+, NO3"  and SCH^- for  station 1 (A-C) and station 2 (D-F), Balmer Lake, June, 1993. Replicate samples are represented by double symbols at specific single depths. Fig. 4.13. Fall peeper profiles of dissolved NH4+,  - ^°  NO3" and SO42- for stations 1  (A-C)  and 5 (D-F), Balmer Lake, October, 1993. Replicate samples are represented by double symbols at specific single depths.  pft • T' 2  Fig. 4.14. Duplicate fall peeper profiles of dissolved  NH4+, NO3"  respectively) for station 2, Balmer Lake, October, 1993. represented by double symbols at specific single depths.  and  SO4" (A-C, 2  Replicate samples are pft- ^  Fig. 4.15 Winter peeper profiles of dissolved NH4+, N03~ and SO42- for stations  3  1  (A-  C) and 2 (D-F), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  ix  Fig. 4.16. Winter peeper profiles of dissolved NH4 , NO3" and S04^ for stations 4 (A+  _  C) and 6 (D-F), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  ^'  Fig. 4.17. Duplicate spring peeper profiles of dissolved NH4+, NO3" and S04^- (A-C, respectively) for station 1, Balmer Lake, May, 1994. Replicate samples are represented by double symbols at specific single depths.  ^'  Fig. 4.18. Duplicate summer peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (AF, respectively) for station 1, Balmer Lake, June, 1993.  Replicate samples are . b°  represented by double symbols at specific single depths.  Fig. 4.19. Duplicate summer peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (AF, respectively) for station 2, Balmer Lake, June, 1993. represented by double symbols at specific single depths. Fig. 4.20.  Replicate samples are Ff- ^ 6  Fall peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 1, Balmer Lake, October, 1993. represented by double symbols at specific single depths.  Replicate samples are pq-  5  5  Fig. 4.21. Duplicate fall peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 2, Balmer Lake, October, 1993. represented by double symbols at specific single depths. Fig. 4.22.  - ^ 6  Fall peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 5, Balmer Lake, October, 1993. represented by double symbols at specific single depths. Fig. 4.23.  Replicate samples are  Replicate samples are f*§ -  ^  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 1, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. Fig. 4.24.  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 2, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  x  p%.  Fig. 4.25.  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 4, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. Fig. 4.26.  PQ'**°  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 6, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  P^-  Fig. 4.27. Duplicate spring peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 1, Balmer Lake, May, 1994. Replicate samples are represented by double symbols at specific single depths.  ' ^  Fig. 4.28. Spring profile of porewater and bottom water pH measured from peeper 2, station 1, Balmer Lake, May, 1994.  f% •  f  6  Fig. 4.29. Distributions of dissolved NO3- and SO4 " obtained from peeper (open circles) 2  and core porewaters (diamonds) at station 1 (A-B) and station 6 (C-D), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. Fig. 4.30.  Pff- * 9  Dissolved distributions of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) obtained from peeper (open circles) and core porewaters (diamonds) at station 1, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. Fig. 4.31.  Pf-  / o € >  Dissolved distributions of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) obtained from peeper (open circles) and core porewaters (diamonds) at station 6, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  Pft •  / o 2  -  Fig. 4.32. Sedimentary weight ratio profiles of Si/Al, Ti/Al, Mg/Al, K/Al, Na/Al and P/Al (A-F, respectively) for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994.  xi  f^. / " V  Fig. 4.33.  Sedimentary weight ratio profiles of Zr/Al, Sr/Al and Rb/Al (A-C,  respectively) for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994. Fig. 4.34.  Pf- '°  5  Sedimentary concentrations of organic carbon, calcium carbonate, total  nitrogen, values for organic carbon/nitrogen, total sulphur and iron/aluminum weight ratios (A-F, respectively), for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994.  p^-  / 0 < a  Fig. 4.35. Sedimentary concentrations of manganese, nickel, zinc, copper, arsenic, and lead (A-F, respectively) for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994. Fig. 5.1. Time series of progressive trace metal water column profiles (A-E) over the fall-winter transition in Balmer Lake: A) well-mixed water column, B) influence of remobilized bottom-water source, C) influence of interfacial sulphide sink, D) greater extent of sulphide removal, and E) influence of metal-rich lateral advective flow. Depth and concentration axes are in arbitrary units. Fig. 5.2.  rf •  Seasonal peeper profiles of ammonium for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths.  P% • > s2  Fig. 5.3. Seasonal peeper profiles of nitrate for stations 1 and 2 (A and B, respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single Pf-  depths.  1 5  *  Fig. 5.4. Seasonal peeper profiles of sulphate for stations 1 and 2 (A and B, respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. Fig. 5.5.  n-  /  r  6  Seasonal peeper profiles of dissolved Fe for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths.  pft • ^°  xii  Fig. 5.6. Winter peeper profiles of dissolved Fe for stations 1, 2, 4 and 6 in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. Fig. 5.7.  p^. Ibt  Seasonal peeper profiles of dissolved Mn for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths.  f^.i^  Fig. 5.8. Time series of progressive bottom water and porewater profiles (A-D) over the fall-winter transition in Balmer Lake: A) fall profile, B) instantaneous input of Mn to bottom waters, C) profile after some time of equilibration, and D) late winter profile. Depth and concentration axes are in arbitrary units. Fig. 5.9.  Pf ^ /6>  Seasonal peeper profiles of dissolved Ni for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. Fig. 5.10.  Pfi"'  Seasonal peeper profiles of dissolved Cu for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. Fig. 5.11.  P%'  Seasonal peeper profiles of dissolved Zn for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths.  Prj-  / & l  Fig. 5.12. Summer peeper profiles of dissolved Fe, Mn and As for station 2 in Balmer Lake. Replicate samples are represented by double symbols at specific single depths, pg. fif Fig. 5.13. X-Y scatter plot illustrating the dependence of dissolved As on the distribution of dissolved Mn. The values (molar concentraions) represent a depth interval spanning from ~3 cm above the sediment surface to ~20 cm below. Coefficients of determination (r ) from linear regression analyses for the respective stations are included. 2  Fig. 5.14.  p<g  / < ?  ^  Seasonal peeper profiles of dissolved As for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths.  ill  xiii  LIST OF TABLES Table 2.1. Estimated average yearly ranges of mine-related inputs from the Campbell Mine tailings system to Balmer Lake, Ontario, before and after chemical treatment. All values expressed in mg/L.  Pft- 22  Table 2.2. Estimated average yearly concentrations of mine-related inputs from the Arthur White tailings system to Balmer Lake, Ontario. All values expressed in mg/L.  ft-  3 3 >  Table 2.3. Summer, fall, winter and spring sampling periods at Balmer Lake, Ontario, indicating stations of peeper deployment, coring and water column sampling.  P?"  z  Table 4.1. Summer water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb, NH4+, N03", S04 " and P O 4 - for stations 1, 2, 3, and 4 in Balmer Lake, 3  2  Ontario, June 1993. All values are expressed in u\g/L unless specified otherwise (BDL denotes below detection limit).  F%• ^  b  Table 4.2. Fall water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb, NH4+, N O 3 - , S O 4 - and P O 4 - for stations 1, 2, 3, 4 and 5 in Balmer Lake, 2  3  Ontario, October, 1993.  A l l values are expressed in | i g / L unless specified  otherwise (BDL denotes below detection limits).  Pf- ^ ^  Table 4.3. Winter water column distributions of dissolved Mn, Fe, N i , Cu, Zn, As, Pb, N H 4 + , N O 3 - , S O 4 - and P O 4 - for stations 1, 2, 3 and 4 in Balmer Lake, 2  Ontario, March, 1993.  3  A l l values are expressed in u g / L unless specified  otherwise (ND denotes not determined).  P(f" ® ¥  Table 4.4. Spring water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb, NH4+, N03', S O 4 - and P 0 4 ' for stations 1, 2 and 4 in Balmer Lake, 2  3  Ontario, May, 1993. All values are expressed in ug/L unless specified otherwise (ND denotes not determined).  • * s  xiv  Table 4.5. Concentrations of total suspended solids (TSS), particulate organic carbon (SPOC), particulate organic nitrogen (SPON), Corg^itrogen ratio (C/N) and total particulate organic matter (POM) in the water column of Balmer Lake, Ontario, for the summer, fall, winter and spring field sessions. TSS values were determined from the particulate mass on pre-weighed 0.45 um Nuclepore filters, while all organics were determined from glass fibre filters.  P%' ~ ss  Table 4.6. Particulate trace metal concentrations of Fe, Mn, Ni, Cu, Zn, As and Pb for the summer, fall, winter and spring water columns in Balmer Lake, Ontario. All values are expressed in ug-L~l.  Table 4.7.  Pf -  Concentrations of trace metals (mg-kg"l), organic carbon (%) and  organic nitrogen (%) in suspended particulates of the summer, fall, winter and spring water columns in Balmer Lake, Ontario. Trace metal values were determined from the particulate mass on pre-weighed 0.45 um Nuclepore filters, while all SPOC and SPON values were determined from glass fibre filters, p*.  xv  (af-  ACKNOWLEDGEMENTS This project was jointly jointly funded by Placer Dome Inc. and Gold Goldcorp Inc. Special thanks goes to Richard Janicki, Dave Hiller and Elias Dibb of Campbell Mine, Red Lake, who helped ensure the success of the fieldwork. We also thank Jim Robertson (PDI, Vancouver) for his interest and continued support. I wish to thank Drs. Tom Pedersen, Steve Calvert, George Poling and Bill Cullen of my supervisory committee for their various degrees of input into my thesis. In particular, I deeply express my thanks to Tom Pedersen who is not the uncultured, useless, dirty, stinking heathen everyone says he is. I will always have fond memories of sitting in his office, Tom reclining back in his chair, hands clasped and brought to his mouth, his eyebrows progressively lowering over his piercing stare as I presented him with my latest views. "You have done a good job" he would say, as he passed you your crimsoned document, the pages weighted and thickened by the rigorous onslaught. I also extend my gratitude to Steve Calvert, whose insight and depth would intimidate the likes of Carl Sagan and Mr. Spock. As for Bill Cullen, I appreciate the fact that he left the country as I heard he can be quite intimidating at oral examinations. Additional thanks goes to my fellow research group members, Remy Chretien, Darcy McDonald, Jay McNee, Raja Ganeshram, Robert Mugo, Brad Mckelvey, Maureen Soon and Bert "high sticking" Mueller, for making my graduate years (all four of them) a lot of fun. I will also never forget John "Leatherman" Ridley, whose mighty power aided in both the peeper deployments and his ability in snow wrestling. The boys (you know who you are) should also be praised for their efforts in getting me out of the house. I dedicate my thesis to my partner in crime, Debbie, who not only proofread this document, but ensured I received regular doses of mountain air throughout my degree.  xvi  I. INTRODUCTION 1.1 Trace Metal Geochemistry The study of trace metal geochemistry is driven partly by curiosity, and as a result of the inherent toxicity of many metal species, partly by an obligation to the maintenance of human health and the well-being of the environment. The last decades have seen a large increase in the burden of trace metals in the environment, and as a consequence, enrichments now commonly exist in terrestrial and aquatic reservoirs (Ferguson and Gavis, 1972; Nriagu, 1989). Considerable attention has been given to the fates and effects of toxic metals. In recent times, the major research objectives on metal-polluted systems have progressed from initial surveys of sources and pathways toward more detailed investigations of the mechanisms controlling the distribution, mobility and bioavailability of different metal species. Knowledge of the distribution and behaviour of trace elements in aquatic and marine systems has grown significantly over the past two decades, largely in part to advancements in sampling techniques and analytical methods. The development of clean collection techniques (Boyle and Edmond, 1977; Bruland et al, 1979; Schaule and Patterson, 1981; Nriagu et al, 1993) and sensitive analytical instrumentation has facilitated the reliable measurement of trace constituents; the use of clean methods has also exposed inaccuracies in previously reported data.  A s a result, well-defined vertical and horizontal variations in  concentrations of dissolved components for the world's oceans now exist for the majority of trace elements. Prior to the implementation of clean techniques, the state of trace metal biogeochemistry of lacustrine systems suffered similar misrepresentation. Contamination artifacts introduced during the collection,  1  handling and analyses of samples severely compromised published data acquired in the absence of rigorous methods. The assumption that stringent, ultraclean laboratory methods are not required in "polluted" waters warrants close scrutiny following recent studies of the Great Lakes which report trace element concentrations comparable to, or even lower than, the levels commonly observed in the pelagic ocean (Nriagu, 1996; Flegal, 1988; Coale, 1989). In the open ocean, dissolved trace element distributions are generally considered to be in steady state (Murray, 1987). Integrated, systematic studies have been used to determine the biogeochemical and physical processes governing the vertical and horizontal distributions of many trace elements (Bruland, 1980; McKelvey, 1994).  Application of one-dimensional, vertical  advection diffusion models fitted to oceanic profiles allow the extraction of chemical parameters such as production and scavenging rates (Morel and Hudson, 1985). In addition, oceanographic investigations have revealed the usefulness of many trace elements as effective markers of various marine processes. The science of limnology has evolved greatly since the original pioneering works of Forel (1892, 1895, 1904), Mortimer (1941, 1942, 1951) and Hutchinson (1957). In spite of significant progress, chemical limnology has not yet reached the same state of sophistication as its marine counterpart. Essentially the same mechanisms act to control trace element distributions in lacustrine settings; however, trace metal distributions are driven by time-dependent processes which vary from days to seasons (Balistrieri and Murray, 1992). The variability induced by the transient nature of lake processes can make steady state assumptions invalid, and as a result, may inhibit the isolation of the variables controlling trace metal cycles (Murray, 1987). Furthermore, the dynamic nature of lakes hinders comparisons of different systems due to variations in factors  2  such as the surrounding geology and trophic status. Defined patterns of trace metal distributions in lakes are further hindered by the greater availability of competing scavenging phases (e.g., detrital particles, phytoplankton and authigenic precipitates) and compressed depth scales. Although the transient nature of lake processes adds extra complexities, time-dependent variations (e.g., oxic/anoxic cycles) can be taken advantage of to establish the dynamics of biogeochemical cycles (Balistrieri and Murray, 1992).  1.2 Trace Metal Cycling in Lakes  The intrinsic properties of a particular metal (e.g. crystal field stabilization energy, electronegativity, ionic radius, etc.) and the environmental conditions prevalent (e.g., types and concentrations of ligands present, available adsorptive surface areas, pH, pe) are ultimately responsible for the distribution of particular species and their corresponding redox and coordination chemistries (Stumm and Morgan, 1981). Systematic studies of trace elements in lakes provide information on the physical, chemical and biological processes controlling their respective mobility and distribution.  These processes include diffusion, advective  transport, nutrient-like biological cycling, sorption by sediments or suspended particles, precipitation, atmospheric deposition, redox cycling, and fluxes across the sediment-water or air-water interface (Murray, 1987).  The relative  importance of these mechanisms depends on the prevalent abiotic and biotic environment as well as the element under consideration. Spatial and temporal changes in the chemical environment of lacustrine systems can strongly influence biogeochemical pathways of various metal species. For example, lakes may possess chemical gradients with depth (e.g.,  3  dissolved O S 0 y  4  , etc.) and exhibit significant seasonal differences in such major  variables as p H and pe (Jenne, 1986). Furthermore, seasonal variations in productivity can foster corresponding shifts in pe and concentrations of organic products.  Increases in autochthonous particles for instance, may provide  additional complexation sites for trace metals, resulting in lower concentrations of dissolved metal species (Francois, 1990). These multiple influences are briefly elaborated upon in the general review of trace element cycling in lake systems that is presented in the following sections.  A brief introduction to the sources of trace metal inputs will be  followed by more detailed descriptions delineating the processes governing metal behaviour and mobility in both the water column and underlying sediments and porewaters.  1.2.1 Inputs of Trace Elements to Lacustrine Systems  Trace metals are incorporated into lake systems primarily via tributary inflow, ground water inputs, coastal runoff, industrial and domestic effluent discharges and atmospheric fallout. Natural sources are primarily water or wind-borne soil particles, seasalt aerosols, volcanism, forest fires and biogenic particulate matter (Nriagu, 1989) . Nriagu (1989) estimated that aerosols of biogenic origin dominate the natural fraction and can account for 30-50% of the global baseline emissions of trace metals in non-urban areas. Hydrothermal activity has also been shown to provide significant quantities of trace metals to specific fresh waters (Agett and O'Brien, 1985). Background metal levels in inland waters due to weathering processes are more difficult to establish than in the marine environment as a result of varying geology and fluctuations in water discharge (Baccini, 1984). In contrast, anthropogenically enriched levels of  4  metals are derived from domestic wastewater effluents, mining-related inputs, gasoline combustion, coal burning power plants, non-ferrous metal smelters, iron and steel plants and the dumping and incineration of sewage sludge (Nriagu, 1989).  1.2.2 Major Processes Governing Trace Metal Behaviour and Mobility  The distribution of trace metals between solution and particulate forms has widely been recognized as a major factor controlling the geochemical behaviour, transport and biological effects of metals in natural waters. Many trace metals (e.g. Cu, Pb, Hg, C d , Zn, Ag) have been shown to be primarily associated with suspended and sedimentary particulates as opposed to dissolved phases (Davis and Leckie, 1978b; Nriagu et al, 1981; Laxen, 1985; Tessier et al., 1985). Particulates in natural waters consist predominantly of detrital particulate and colloidal organic matter, inorganic solids such as metal oxides and hydroxides (e.g., Si02, MnC>2, FeOOH, A I 2 O 3 ) , algal skeletal remains, carbonates and detrital aluminosilicates (e.g., clay minerals, feldspars) (Tessier et al, 1985). In general , particulate organic matter (POM) often presents the major component of suspended material, accounting for up to 70% of the total particulate fraction (Nriagu et al, 1982). Settling particles, especially organic aggregates, can play a dominant role in the binding and transfer of heavy metals to lake sediments, thereby regulating the concentrations of dissolved species (Nriagu et al, 1981; Sigg, 1985; Jackson and Bistricki, 1995). In general, inverse relationships between the residence times of particulate metals and the standing crops of suspended matter in the water column are evident for most lakes (Sigg, 1985). Variations in the residence times of various metals primarily reflect differences in their leachability from source  5  particulates and in the rates of biotal assimilation (Nriagu et al., 1982). Sigg (1985) and more recently Morel and Hudson (1985) have extended the Redfield ratio concept (Redfield, 1958)  to include trace elements, suggesting a  stoichiometric formula of C106N16P1 (Fe, Zn, Mn)n.oi (Cu, Cd, N i , Co) o.ooi- Sigg (1985) suggested that algal-metal associations could result in metal uptake with constant stoichiometric proportions. Although the concept of Redfield ratios may be applicable to the metal content of algae, the dynamic nature of lakes limits its usefulness because only under severely restricted conditions will the stoichiometry be reflected in dissolved metal concentrations (Reynolds and Hamilton-Taylor, 1992). Cyclic variability in the composition of the total suspended matter is evident in many temperate lakes.  In winter and early spring, isothermal  conditions typically predominate and storm events can resuspend quantities of fine sedimentary matter (Baccini, 1984). This particulate fraction may have undergone considerable diagenetic alteration and be characterized by relatively low organic carbon and high mineral contents (e.g., clays). With the onset of large-scale vernal primary production, phytoplankton, zooplankton and the associated detrital material may dominate the particle pool. During periods of thermal stratification, the epilimnion is largely decoupled from the underlying colder waters of the hypolimnion resulting in unidirectional, downward transport of particulate matter. During the latter stages of stratification in the late summer and fall, the p H may be driven high enough by the production of organic matter to promote the precipitation of CaC03 in surface waters. Inorganically precipitated calcite has been shown to account for 20-90% of the particle pool in the Great Lakes during such periods (Eadie and Robbins, 1987). The influence of biologically- and chemically-mediated redox reactions on the behaviour of trace elements has received considerable attention owing to the  6  pronounced effect such transformations have on the spatial and temporal distributions and on the speciation of such elements in solution, suspended particulates and sediments.  The predominant mechanisms driving the redox  cycling of elements in lakes are plankton growth in surface waters and the bacterially-mediated degradation of organic matter in sub-surface waters and sediments (Hamilton-Taylor and Davison, 1994). Redox-driven biogeochemical pathways of elements may involve direct redox transformations (e.g., As(V) - » As(III)) in which the metal species itself undergoes a redox reaction (De Vitre et al, 1991). Alternatively, metal species may be subject to indirect involvement, in which the distribution of the element is affected by redox-driven cycling, but apparently does not itself (e.g., Zn) undergo a redox transformation (Reynolds and Hamilton-Taylor, 1992). The redox behaviour of iron and manganese in lakes has received considerable attention as a result of the central role that these abundant metals play in the geochemical cycling of trace elements (Sholkovitz, 1985; Leckie, 1986; Davison, 1993; Hamilton-Taylor and Davison, 1994). Both elements are common to natural waters in discrete mineral phases and in surface coatings on particulate material, usually occurring in oxic conditions as insoluble oxides at their higher oxidation states. Their reduced states, Fe2+ and Mn2+, are soluble and relatively uncomplexed (Davison et al,  1982).  Sedimentary iron and  manganese oxyhydroxides exhibit high affinities for metal ions (e.g., As, Ni, Cu, Cd, Zn, and Pb) and have been repeatedly inferred to be sinks for trace metals (Crecelius, 1975; Nriagu et al, 1981; Davison and Woof, 1984; Tessier et al, 1985; Tessier et al, 1996). The sorptive affinity of trace metals for oxide surfaces is strongly influenced by p H (Vuceta and Morgan, 1978), ionic strength (Stumm and Sulzberger, 1992), adsorbed ligands (Davis and Leckie, 1978a; Vuceta and Morgan, 1978) and humic substances (Tipping and Cooke, 1982). Theoretical and  7  experimental works on solid-solution interactions of trace metals with amorphous iron oxyhydroxide demonstrate the existence of a multiplicity of discrete binding sites with wide-ranging adsorptive characteristics (Davis and Leckie, 1978b; Benjamin and Leckie, 1981). The preferred adsorption sites for such trace elements as C d , Zn, C u , and Pb are apparently distinct from one another, which explains why competitive effects observed are weaker than one might otherwise expect. This heterogeneity of complexation sites has been explained in terms of the number and identity of nearest neighbours, proximity to various crystal defects, microcrystallinity, local electrical field strength, and physical constraints such as the size of pores leading to internal sites (Benjamin and Leckie, 1981). Seasonal variations in the redox chemistry of iron and manganese in lakes can result in pronounced inter-annual variability in the distribution, behaviour and mobility of many trace elements. Non-steady state redox environments induced by productivity pulses and changes in local hydrography can significantly affect the distribution of their respective soluble and particulate phases (Tipping et al, 1981). Meromictic lakes offer a range of redox conditions which vary spatially and temporally, and consequently, such basins are particular well suited to studying nuances of redox processes (Davison and Woof, 1984). Temperate lakes tend to exhibit seasonal stratification of the water column in the summer months which is induced by surface heating of a usually shallow mixed layer (Baccini, 1984).  The redox boundary in a seasonally  stratified lake may migrate from a horizon within the sediments under wellmixed conditions upwards into the water column during episodes of maximum stratification (Laxen, 1985). During unstratified conditions, oxygen may be present at the sediment-water interface, favouring the formation of M n and Fe oxyhydroxides in the surficial sediments (Davison et al., 1982). Upon burial of  8  the sediment, diagenetic respiration processes deplete oxygen and promote a reducing environment. The reductive dissolution of oxyhydroxides in reducing sediments promotes the mobilization of Fe(II) and Mn(II) and the concomitant mobilization of sorbed metals.  In the absence of further chemical reactions,  Fe(II), Mn(II) and the associated trace metals may diffuse upward along concentration gradients towards the sediment-water interface, where the predominant reactions are reprecipitation under oxic conditions, and release into the hypolimnion when anoxic conditions prevail (Agett and O'Brien, 1985; Agett and Kriegman, 1988).  In general, as the oxic zone of the sediments is  compressed, retention of metals in the sediments primarily associated with metal oxides is decreased.  Those elements with lower ionization potentials (i.e.  decreased capability to take up electrons and thus a reduced tendency to form complexes) are more readily mobilized and thus enter solution at higher pe values (Stumm and Morgan, 1981). During the later stages of stratification, the hypolimnion may become sufficiently devoid of oxygen to enable the dissolved species to diffuse across the sediment-water interface where they may accumulate until overturn (Davison and Woof, 1984; Agett and Kriegman, 1988). The duration of stratification and the volume of the hypolimnion strongly influence the fluctuations in the concentrations of dissolved trace metal species (Davison 1982).  The misinterpretation of diagenetic metal enrichments as  anthropogenically related features is, however, a potential problem in mildly reducing, Mn- and Fe-rich lake sediments.  9  1.2.3 Transformations Across the Sediment-Water Interface  Bottom sediments of lakes represent complex mixtures of diverse components  that include: residues  of weathering and erosion (e.g.,  aluminosilicates), inorganic and organic products of biological activity, and diagenetic products such as iron and manganese oxyhydroxides and sulphides (Tessier and Campbell, 1988). Trace metals may be scavenged from the water column and incorporated into the bottom sediments via several mechanisms including: sorption onto hydrous metal oxides, ion exchange with clay minerals, sorption on organically coated particulates, adsorption to or incorporation with detrital organic matter, and via diffusive influxes along concentration gradients. The particular physical and biogeochemical characters that dictate sediment geochemistry ultimately determine the nature of the various diagenetic (i.e., post depositional) reactions and resulting authigenic fractions. Authigenic phases encompass a wide spectrum of inorganically-derived precipitates which form in the water column or in the sediments after deposition.  Organic matter  constitutes a very important sedimentary component, as its diagenesis during decomposition provides the primary impetus behind the post-depositional reactions responsible for the formation of numerous hydrogenous components (Suess, 1979). Organic matter degradation has additional ramifications with respect to nutrient regeneration and benthic community structure (Emerson, 1976; Froelich et al, 1979). Studies of the microbially-mediated decomposition of organic matter have led to the development of a well-established model in which the distributions of oxidants, reductants, and metabolites with depth follow a predictable and universal sequence, given an elastic depth scale. Detailed accounts have been described by several authors (Sholkovitz, 1973; Froelich et al,  10  1979;  Klinkhammer, 1980; Van der Weijden, 1992). Briefly, this degradation follows a thermodynamic order of sequential  oxidation reactions  which yield  progressively lower free energy changes per mole of organic carbon oxidized. In practice, the reaction sequence is subject to oxidant availability and kinetic constraints, including the dominance of particular microbial populations. As previously stated, thermodynamics dictates that relative pore water distributions of metabolites will be consistent from site to site with the only variable being the depth scale over which these reactions occur.  Organic-rich sediments,  characterized by high diagenetic intensity and thus steep thermodynamic gradients, may, however, exhibit significant overlap in the sequence of oxidation reactions (Froelich et al, 1979). A recent evaluation of the redox zonation during the degradation of organic matter assessed the controls responsible for observations of the simultaneous reduction of Fe(III) and sulphate in several field studies (Postma and Jakobsen, 1996). They suggest that the sequential occurrence of these processes can be better explained by a partial equilibrium approach where initial fermentation determines the overall rate (i.e., the reactivity and amount of organic matter) while equilibrium is approached by the terminal electron accepting processes. They concluded that depending on the stability of the iron oxide and porewater pH, sulphate may be reduced simultaneously with, or even prior to Fe(III). The flux of organic matter to the sediments exerts the primary control on diagenetic intensity and has a corresponding influence on metabolite distribution. In environments of rapid sediment accumulation and those with high input rates of organic matter, interstitial metabolite concentrations (e.g., H C O 3 - , H P O 4 " , HS") may exceed the solubilities of certain inorganic equilibria 2  11  resulting in the formation of authigenic minerals and thus preferential removal of one or several of the dissolved constituents (Suess, 1979; Berner, 1984). Thermodynamic calculations on the bulk solution of most freshwater and seawater typically show undersaturation with respect to known metal solid phases (Tessier et al. 1985). Even for interstitial waters of oxic sediments, in which trace metals have longer residence times than in the bulk solution and are thus more apt to approach a state of equilibrium with solid phases in the sediments, metals are usually undersaturated with respect to their least soluble compounds (Schindler, 1981). Trace metal adsorption to solid surfaces has been invoked to account for observed distributions.  The trace metal-particulate  association may include: sorption of metal ions at oxide surface sites, ion exchange within clay minerals, binding by organically-coated particulate matter or organic colloidal material, or adsorption of a metal-ligand complex.  In  particular, the important roles played by iron and manganese oxyhydroxides and humic materials in the adsorption of trace metals in soils and sediments have been strongly emphasized due to their high sorptive properties and ubiquitous presence in natural systems . Thermodynamic modeling of interstitial waters is a useful technique for identifying possible controls on authigenic mineral formation. Experimentally derived ionic strengths, activity coefficients, and activities of various interstitial metabolites are often used to calculate the controlling mechanisms responsible for the distribution of various aqueous and solid fractions. Production and consumption rates of interstitial metabolites can be used to constrain further thermodynamic gradients and redox pathways. The availability of powerful mathematical tools which incorporate variables such as saturation indices, distribution coefficients  and diffusive fluxes, aid in the development of  conceptual models of sediment geochemistry. In more complex natural settings,  12  however, calculated values often deviate from the observed as complications arise from mixed mineral phases, adsorption and/or ion exchange equilibria (Matisoffef al, 1980).  1.3 Contairrination of Aquatic Systems From Metal Mining  While estimates of trace element emission rates to the environment differ (Lantzy and Mackenzie, 1979; Jaworowski et al, 1981; Zoller, 1984; Nriagu, 1989), it is agreed that for most toxic metals emissions from industrial activities are larger than natural fluxes. Human operations have thus become the key agent in the global atmospheric cycle of trace metals and metalloids. For example, a comparison of world-wide median emission values from natural and anthropogenic sources suggests that industrial emissions of Pb, C d and Zn exceed the flux from natural sources by factors of 18, 5 and 3, respectively (Nriagu 1989). Further, contributions from anthropogenic sources of As, Hg, N i , Sb and V may exceed those from natural sources by 100-200% (Nriagu 1989). The metal mining industry contributes significantly to the anthropogenic input of many trace elements to the Earth's atmospheric, terrestrial and aquatic systems (Nriagu, 1989). Impacts resulting from such activities can be highly localized, such as an acid-generating waste-rock pile, or can extend over hundreds of kilometres via emissions from the smelting of ore concentrates. In the case of sulphide-bearing ore deposits, the potential for loadings of acidity and metals to the environment presents the primary concern. In particular, smelter emissions, acidic drainages and the erosion of waste rock deposits can foster extensive and lasting environmental impact (Aimer et al, 1978; Salomons, 1995).  13  1.3.1 Bioavailability and Toxicity  The uptake of trace elements by biological systems is a speciation-sensitive process for which chemical specificity varies from one element to another (Turner, 1984; Campbell et al., 1988). Bioavailable forms include free metal ions, organometallic species including organic complexes, and inorganic complexes including hydrolysis products (Turner 1984). Though much remains to be learned about the mechanisms of trace element uptake and nutrition in phytoplankton, the assimilation of trace metals appears to be in most cases a two stage process involving the binding of a relatively large pool of metal to the cell surface and transport through the cell membrane, presumably via metal-porter molecules (Morel and Hudson, 1985). The exact nature of either the surface binding groups or the porter molecules remains questionable.  Seemingly, a  significant fraction of the surface ligands must be complexed by a particular metal for that species to be taken up, the extent of binding being related to the intracellular assimilation (Turner 1984). Indiscriminate uptake of non-essential metal species appears to be related to the relatively low specificity of metal transport mechanisms and thus presents an important mode of toxicity (Harrison and Morel, 1983). In addition to the major nutrient elements, many trace metals including Fe, Mn, Zn, Cu and N i have been determined to be essential for life, and it has been argued that many major and trace elements, both essential and toxic, may be simultaneously controlling biological production in the oceans (Morel and Hudson 1985). These trace metals exhibit fairly narrow "concentration windows" that separate essential and toxic levels (Florence 1982). In the open ocean where concentrations are low, toxic effects are negligible; however, in coastal regions and lakes, metal inputs from weathering and anthropogenic processes result in  14  substantially higher metal concentrations that may exceed toxic thresholds (Elinder, 1984).  1.4 Research Objectives  Many trace elements exhibit toxic effects towards plant and animal species when present in sufficiently high concentrations. Particular attention has been given to the environmental significance of several heavy metals and metalloids (e.g., As, Hg, Pb, N i , Zn, Cd, Cr) in view of their inherent toxicity and common occurrence (Elinder, 1984; Cullen and Reimer, 1989). Lake sediments represent the ultimate site of deposition for many trace elements introduced into the environment. Since an important fraction of the trace metals present in the aquatic environment is reversibly associated with the surficial sediments (Campbell et al., 1988), the study of trace metal transformations and dynamics across the sediment-water interface is prerequisite to the understanding of their behaviour in whole aquatic systems. The primary impetus behind this research was to investigate the poorly constrained biogeochemical cycling of metals within the water-sediment system in Balmer Lake. It is anticipated that the project will provide basic insight into the cycling of metals in shallow, boreal contaminated lakes as well as data which can be used to assess both the overall impact of the final discharge from Balmer Lake on the local water system, and prospects for mitigation of the contamination problem over the long term. More specifically, the research comprises three principal objectives: 1) To determine the biogeochemical processes governing the behaviour and mobility of trace metals in Balmer Lake sediments and assess the extent to which these metals are mobile; 2) To determine the direction and significance of benthic fluxes of trace metals; and 3)  15  To assess how the chemistry of the water column, underlying sediments and associated porewaters, change over the course of a full four-season cycle. A seasonal study of Balmer Lake was essential in order to resolve the nature and extent of geochemical variations in response to seasonal changes in hydrography, oxygenation and plankton productivity. In particular, it was desired to link periods of seasonal anoxia and the influence of redox-related reactions on metal mobility and diagenetic chemistry. To meet these objectives, a two component research program was implemented: first, a seasonal assessment of the water column chemistry and second, a detailed seasonal investigation of the chemistry of the sediments and associated porewaters. The latter involved high resolution sampling of the interstitial waters and bottom waters. Balmer Lake provides an ideal in situ laboratory to address the nature of metal transfer mechanisms and partitioning across the sediment-water interface for two reasons: first, there is a large signal of metal contaminants; and second, the large benthic area to volume ratio makes the chemistry of the lake suceptible to influences from interfacial fluxes. To determine accurately the nature of post-depositional reactivity, detailed analyses of both solid and dissolved phases are essential (Colley et al, 1984; Carignan and Nriagu, 1985). Undetectable changes in the sediment fraction can significantly alter the chemical composition of the associated porewaters which can thus provide a sensitive indicator of the diagenetic reactions and equilibria between solid phases and dissolved species (Matisoff et al. 1980). In addition, the residence times of trace elements in the interstitial fraction are longer than in the overlying waters; porewaters are therefore more apt to approach a state of equilibrium with solid sediment phases (Schindler, 1981). Consequently, the composition of interstitial waters can be used to identify and quantify mineral equilibria and thus infer chemical mass transport (Matisoff et al.  16  1980). Porewater studies done in conjunction with investigations of the water column can provide a link between water column transport processes and sedimentary accumulation by showing evidence for the release of trace elements associated with the degradation of biogenic particulate matter (Shaw et al, 1990).  17  II. STUDY SITE  2.1 Physiography, Physical and Chemical Limnology  Balmer Lake is a small, shallow, boreal lake situated in the Canadian Shield of northwestern Ontario (Fig. 2.1). The waterbody is 260 hectares in area, has an estimated mean depth of 2.4 m and a maximum basin depth of 4.0 m. Typical of shallow Precambrian shield lakes, Balmer Lake is productive and hosts organic-rich sediments (Conroy and Keller, 1976). The shallow water column maintains fairly well-mixed conditions during the ice-free periods of late spring, summer and fall. Periods of bottom water anoxia develop in the winter months in deeper areas (>2.5 m) of the ice-covered lake. The waterbody has served as a repository of mining-related wastes for two underground gold mines since 1965 (Fig. 2.1).  Both mines process high  grade ores rich in N i , Cu, Zn and As, and as a result, concentrations of these metals are significantly enriched in the lake waters and sediments. In addition, the lake hosts high inventories of other mining-related byproducts, including cyanide, ammonia, nitrate and sulphate. The water retention time in the lake is approximately 230 days, which aids the natural degradation and settling of cyanide and heavy metals before final discharge to the local water system. In the design of the tailings circuit, the lake was intended to receive only dissolved and colloidal loads. However, historical dam breaches of the surrounding tailings ponds have resulted in significant and episodic particulate loadings to areas of the lake. Over the period of mining activity, the lake has supported few fish and is not used recreationally. Since January of 1994, only the Arthur White Mine (Fig. 2.1) has been discharging effluents that have exceeded the recommended  18  Balmer Creek  Tailings Discharge  Fig, 2.1 Location map of Balmer Lake showing the sampling stations, lake bathymetry and tailings circuits of the adjacent mines.  19  C O A (Certificate of Approval) limits with respect to total arsenic and heavy metals. The general topography of the Balmer Lake region has been described as flat to rolling, with lowland coniferous and upland mixed forest cover. Examinations of the Quaternary geology show that the soil types are dominated by glacial deposits of bouldery to silty till one to several metres thick. The area is also characterized by scattered bedrock outcrops and glaciolacustrine deposits of clay and silt deposited in Lake Agassiz (Gulson et al., 1993). Balmer Lake receives inputs from three main drainage basins, encompassing a total watershed area of 3,290 hectares (Masala, 1995 ). The natural drainage area comprises 2,550 ha and has been influenced to a moderate degree by road construction, timber harvesting and mining exploration, contributing to breaks in the natural forest cover. The remaining two drainage basins are represented by the Campbell and Dickenson mine sites which comprise active and revegetated tailings, clear water (polishing) ponds, and parts of the mines and townsite.  Inputs from these  drainages are predominantly mine-related with a small fraction being natural runoff (Masala, 1995a).  2.2 History of Mine-Related Inputs  The Archean greenstone belts in the Red Lake area of northwestern Ontario have been mined for their gold deposits since the gold rush to the district in 1926.  Ore-grade gold mineralization in the area generally occurs in three  specific associations: (i) massive ferroan-dolomite veins that contain gold + quartz + arsenopyrite in crosscutting veins; (ii) silicified patches with disseminated gold; and (iii) sulphide disseminations (arsenopyrite, pyrite, and lesser pyrrhotite) (Gulson et al., 1993). More detailed summaries of the geology  20  of the Red Lake area can be found in Colvine et al. (1984), Corfu and Wallace (1986) and Corfu and Andrews (1987). Production mining in the area of the Balmer Lake began in 1948 upon operation of the Arthur White Mine. Balmer Lake, however, did not receive mining-related inputs until 1965, at which time, tailings were deposited along the basin of the south shore of the lake. A rock berm was subsequently constructed in 1967 to contain particulate mill effluents. Since this time, the waterbody has served as a tertiary polishing pond for mining wastes from both the Campbell (Placer Dome Inc.) and Arthur White (Goldcorp Inc.) mines.  2.2.1 Campbell Mine  The Campbell Mine is an underground gold mine which uses sub-level longhole and cut-and-fill ore extraction methods. Ore processing proceeds via cyanidatioh with carbon-in-pulp and Merrill-Crowe Gold Collection Processes. The coarse tails fraction (~ 460 tons per day (TPD)) is combined with flyash and cement for underground backfilling while the fines (« 470 TPD) are sent to the tailings impoundment and settling ponds.  Prior to January, 1993, decanted  excess water from the main tailings pond proceeded through of a series of polishing ponds (including Balmer Lake) before final discharge to the environment at the Balmer Lake outlet (Fig. 2.1).  However, this system was  insufficient in reducing contaminant concentrations of Ni, Cu, Zn, As and CN" to within C O A limits on a consistent basis.  Subsequently, Placer Dome Inc.  commissioned the installation of a wastewater treatment plant which commenced operations in 1993.  The chemical process utilizes INCO SO2  technology (Devuyst et ah, 1982) and brought effluent parameters to within guideline specifications, thus effectively removing Balmer Lake from the  21  Campbell tailings circuit. A comparison of the effluent quality to Balmer Lake before and after chemical treatment is presented in Table 2.1.  Table 2.1. Estimated average yearly ranges of mine-related inputs from the Campbell Mine tailings system to Balmer Lake, Ontario, before and after chemical treatment. All values expressed in mg/L.  Year  Total  Total N i  Total Cu  Total Zn  Total As  Cyanide Prior to treatment 1992  0.5-30  2-6  1-7  0.3-1.5  0.2-0.5  1993  0.5-25  0.4-2.0  0.1-8  0.05-1.5  0.01-0.5  1994  0.4-1.5  0.05-0.3  0.05-0.2  0.02-0.06  0.01-0.03  1995  0.1-0.6  0.04-0.07  0.02-0.08  0.02-0.04  0.01-0.03  After treatment  2.2.2 Arthur White Mine  The Arthur White Mine began production mining in 1948 and shares a similar mining protocol to its PDI counterpart.  Ore production, rated at  approximately 800 TPD, is achieved by shrinkage, horizontal hydraulic cut and fill, and long-hole stoping mining methods. Arthur White also possesses its own internal primary tailings disposal system, which consists of series of interconnected ponds (Fig. 2.1). Clarified effluent from the primary tailings area overflows the primary dam and drains into a secondary polishing pond. The drainage then proceeds through two rockfill seepage-dams before final discharge to the tertiary polishing pond (Balmer Lake). At the present time, no effluent  22  treatment facilities exist at the mine, and as a result, inputs from this system present the most significant contributor of heavy metals, sulphate and cyanide to Balmer Lake.  Typical effluent concentrations of various parameters are  presented in table 2.2.  Table 2.2. Estimated average yearly concentrations of mine-related inputs from the Arthur White tailings system to Balmer Lake, Ontario. All values expressed in mg/L.  Year  Total  Total N i  Total Cu  Total Zn  Total As  Cyanide 1983  9  0.9  1.0  0.7  0.4  1984  15  1.1  3.2  1.1  1.3  1985  7  0.2  0.4  0.1  0.5  1986  3  0.4  1.1  0.3  0.8  1987  5  0.6  1.0  0.3  0.4  1988  2  0.6  0.9  0.1  0.8  1989  4  0.9  2.1  0.2  1.3  1990  15  1.1  3.3  1.1  1.3  1991  8  0.9  1.1  0.7  1.1  2.3 Sampling Periods  Four sampling periods encompassing a summer-fall-winter-spring transition were carried out in order to represent accurately the full seasonal spectrum of the chemical environment in Balmer Lake. The summer sampling  23  period (Jun-Jul, 1993) was chosen to coincide with a potentially stratified water column, the condition being common to many temperate lakes in the summer months.  The fall session (Oct-Nov, 1993) was planned to follow complete  breakdown of stratification developed in the summer; sampling was performed just prior to freeze-up and after convective overturn of the shallow water column.  In order to observe the maximal extent of sub-ice stratification  established during the winter months, sampling was carried out just prior to iceoff in March-April of 1994.  Gravity coring was also performed during this  period to take advantage of the stable surface offered by the ice; this permitted slow and precise corer insertion. The spring sampling period was carried out two weeks subsequent to ice-off (May 1994) and was designed to capture a postthaw episode of elevated heavy metal and cyanide levels previously observed during this time.  2.4 Sampling Sites  A survey of Balmer Lake in 1990 (PDI, 1990) established a suite of 45 permanent sampling sites, five of which were chosen as sampling stations for this project (Fig. 2.1)  During the winter period, the ice extended to the bottom  sediments at station 5 (~1 m deep), and consequently, sampling in that region of the lake required a move into deeper waters to station 6. It was originally anticipated to sample at stations 1 and 2 during every season. However, as a result of an improper insertion, peeper profiling in the winter session was restricted to station 1 only (see below). To evaluate the precision of the porewater sampling method, two sets of tandem peepers (~ 10 m apart) were first deployed at stations 1 and 2 during the summer session.  Subsequent results indicated good agreement among the  24  porewater profiles yielded by duplicate samplers.  The fall period included  deployments at stations 1 and 2, including a tandem arrangement at the latter. A fourth peeper was deployed at station 5 in order to assess the porewater chemistry in an area of the lake that has received substantial tailings inputs. This area, coined "Tailings Bay", is bordered by a tailings beach and occupies a significant portion of Balmer Lake (Fig. 2.1); the recent sediments are predominantly mine-related. Deployments at stations 1 and 2 were repeated in the winter session. In addition, peeper sampling was performed in a shallow area of the lake near the out-flow (station 3) and at another deeper site in Tailings Bay (station 6., Fig 2.1). Only three peepers were available for the spring field session, two of which were deployed in tandem at station 1. Upon retrieval of the third sampler positioned at station 2, it was concluded from sediment smears on the aluminum lander and stains on the polyethylene frit that the peeper had fallen on its side at or shortly after deployment. As a result, no data for station 2 were acquired. A summary of the peeper and water column sampling stations is presented in Table 2.3.  Table 2.3. Summer, fall, winter and spring sampling periods at Balmer Lake, Ontario, indicating stations of peeper deployment and water column sampling.  Peeper Deployment  Sampling Period  Water Column Sampling  Summer  Tandem peepers at stations 1 & 2  Stations 1,2,3,4  Fall  Stations 1,5; Tandem peepers at  Stations 1,2,3,4,5  station 2 Winter  Stations 1,2,3,4  Stations 1, 2, 3, 6 Cores at station 1 and 6  Spring  Tandem peepers at station 1  25  Stations 1,2,4,6  III. SAMPLING A N D METHODS  3.1 Sampling  3.1.1 Water Column During ice-free conditions, all sampling of the lake water column and sediments, as well as peeper deployment, was carried out from a free-floating surface platform powered by a 9.9 H p outboard.  The deck provided  approximately 12 m^ working area, a sufficiently large area to accommodate a gas generator, winch, a peristaltic pump, a nitrogen gas cylinder, and permanent attachment of a nitrogen glove bag containment structure. Gravity cores, surface grabs and water column samples were drawn through a 3 n\2 hatch in the platform centre.  Four 20 kg anchors maintained position at any given site.  Winter sampling was carried out through 30 cm holes augered through the ice. Water column samples were drawn from the lake using a "Masterflex" peristaltic pump (Cole-Palmer Instrument Co., Chicago, IL) and sampled directly in a nitrogen-filled glove-bag. Three litres were taken at each sampled depth, placed in a cooler, and taken immediately to the lab for filtration. One litre, to be analyzed for dissolved and particulate metal fractions, was filtered through an acid-cleaned, pre-weighted Nuclepore® 0.45 um polycarbonate membrane filter using nitrogen overpressure, acidified to p H 2 with nitric acid (Seastar Chemicals), and stored in acid-cleaned polyethylene bottles. The remaining two litres were filtered through a 0.4 um precombusted Glass Fibre filter using nitrogen overpressure for determinations of suspended particulate organic carbon (SPOC) and nitrogen (SPON).  26  Ammonium concentrations were  determined from the filtrate spectrophotometrically on site; the remainder was frozen for subsequent analyses of nitrate, phosphate and sulphate.  3.1.2 Physical Profiling  Hydrographic profiling of temperature, dissolved oxygen and p H was performed using a Cole Palmer water analyser (model 5150). Profiles were collected by slowly lowering the probe to desired depths and allowing the electrodes and thermistor to equilibrate before proceeding to the next depth. During periods of stratification, higher resolution profiling was attained across the depth of the pycnocline. Profiles were obtained commensurate with water column sampling in order to be temporally consistent. For both ice-free and icecovered sampling periods, all hydrographic profiling was performed at least 5 m away from the point of water column sampling to minimize potential contamination and water column disturbance in the vicinity of the sampler. The oxygen electrode was calibrated in the field on a daily basis while calibration of the p H sensor was performed in the lab.  3.1.3 Porewaters  The diagenetic cycling of metals in the sediments, and their exchange with the overlying water column was assessed by high resolution sampling of the interstitial and bottom waters using dialysis arrays. Diffusion sampling of sediment porewaters has become widely used since their initial application by Hesslein (1976) and Mayer (1976). Due to lower requirements of processing time and equipment, dialysis methods offer a suitable alternative to squeezing and  27  centrifuging techniques. Samplers typically consist of plastic sheets machined with compartments down their length which are filled with DDW and covered with a semi-permeable membrane; upon insertion into the sediments, the samplers are allowed to approach diffusive equilbrium with the adjacent porewaters. In situ dialysis has been shown to be particularly well suited to the study of trace constituents in sedimentary porewaters under field conditions (Carignan et al., 1985) and has been successfully implemented in studies of Canadian Shield Lakes (Carignan and Nriagu, 1985; Carignan and Tessier, 1985a; Gaillard et al, 1986). Potential artifacts are inherent to the spectrum of interstitial water sampling techniques. Methods such as squeezing and centrifugation followed by filtration, for example, are prone to effects resulting from sample oxidation and temperature variations (Bischoff et al, 1970). Diffusion techniques share a similar succeptibility to potential sources of error including the type of membrane used, construction material, dialyzer design, equilibration time and sampler preparation (Carignan, 1984; Carignan et al, 1985; Brandl and Hanselmann, 1991; Carignan et al, 1994). f  Knowledge of the permeability characteristics of each membrane type, its sorptive properties for chemical compounds and microorganisms, and its resistance to microbial degradation are essential to the application of diffusion samplers and determination of optimal incubation times. A review by Brandl and Hanselmann (1991) determined from a study of 13 different membranes that cellulose-based dialysis membranes were the best choice in lake sediments at low temperatures.  The cellulose membranes tested, however, exhibited marked  weight losses due to microbial degradation when incubated at elevated temperatures (~28 °C). Martens and Klump (1980) also observed some physical breakdown of cellulose-based membranes. A digestible membrane polymer  28  could easily produce false porewater compositions due to a weakening of membrane stability, changes in membrane permeability or blockage of membrane pores.  In consideration of warm water conditions (>20 °C)  characteristic of Balmer Lake in the summer and fall, such membranes were excluded from this study.  In its place, a Gelman (HT-450) polysulfone  membrane was chosen for its mechanical stability and resistance to microbial attack. The suitability of such a dialysis system has been validated in several examinations of lake porewaters (Carignan, 1984; Carignan et al, 1985; Gaillard et al, 1986; Carignan et al, 1994). In some peeper preparations, a thin Nuclepore polycarbonate membrane (0.2 urn )was used in conjunction with the Gelman membrane to provide a better seal. Most samplers, however, were prepared solely with a 0.45 u.m Gelman polysulfone sheet. For measurements of porewater constituents, Carignan et al (1985) showed that profiles obtained with a dual 0.45 urn (polysulfone) + 0.03 urn (polycarbonate) membrane system were not significantly different from those obtained with a 0.45 um membrane alone. This result suggests that the Brownian diffusion of colloidal metal species through relatively large pores of 0.45 urn is unlikely  3.1.3.1 Peeper Description and Preparation  The dialyzers implemented in this study are similar in design to that of Hesslein (1976) and consist of 120 x 20 x 2.3 cm Plexiglas plates with two rows of chambers, each 6.5 x 0.7 x 0.7 cm, machined in at 1.25 cm intervals, centre to centre (Fig. 3-1). This scheme is interrupted by a 20 cm section containing 31 chambers, each 3.5 x 0.7 x 0.7, spaced at 0.625 cm intervals. This zone is intended to coincide with the sediment-water interface upon deployment and provides  29  Peeper base plate Machined sample wells, Acrylic face plate machined to match base plate  High resolution section (5mm interval)  \  'A' *  r-  I  1  iiimi  70 fxm  polyethylene frit 0.45 fim filter"  DETAIL 'A*  Figure 3.1. Schematic diagram of membrane dialysis sampler (peeper).  30  greater resolution in the surficial sediments and bottom waters. The peeper design affords high resolution pore water and bottom water profiles of the dissolved constituents from 40 cm above the sediment-water interface to a sediment depth of 50 cm. A membrane filter arrangement positioned between the chamber and face plates was implemented to facilitate dialysis and preclude the entrance of particulates (Fig. 3.1). In addition, the exterior surface of the face plate was fitted with a polyethylene frit to provide a coarse filtering screen, facilitate easier insertion and thus minimize sediment disturbance. Three rows of nylon screws positioned at 4.5 cm spacing down the peeper securely fastened the frit, face plate and filters to the chamber plate. Peeper preparation and assembly was performed at U.B.C. prior to shipment to the study site. The peepers and their Plexiglas cases were first washed in a mild detergent (Isoclean, ISOLAB Inc, Akron Ohio) followed by several rinses in DDW.  The components were then soaked in dilute ultrapure  H N O 3 (Seastar Chemicals) followed by several 24 h soaks in DDW to remove any residual contaminants and acid. Other than serving as nitrogen flushing sleeves, the plexiglass cases served as sturdy transport containers and provided a sterile transport environment. The Nuclepore and Gelman membranes received a dilute soak in ultrapure H N O 3 and several rinses in DDW. Peeper assembly was carried out in a laminar flow hood (class 100) in a shallow plexiglass tank of deoxygenated, distilled, deionized water (DDDW). The chamber plate was first submersed, and then using a clean pipette tip, all bubbles on the acrylic surface were dislodged. configurations were utilized.  Two peeper membrane  The first consisted of a twin membrane  arrangement of a Nuclepore filter (polycarbonate, 0.2 u,m poresize) overlain by a Gelman HT-450 filter (polysulfone, 0.45 urn poresize). The second arrangement consisted of only a 0.45 |xm Gelman filter membrane. Upon placement of the  31  filter membrane(s) on the chamber plate, the acrylic face plate and polyethylene frit (70 um pore size) were placed appropriately and securely fastened with 51 nylon screws.  The assembled peepers were stored in sealed acrylic cases  containing DDDW bubbled with nitrogen until deployment. All peeper flushing cases were purged with nitrogen for approximately 48 hours before transport and for a further 24 hours prior to deployment. Carignan et al. (1994) have recently brought attention to potential effects of free oxygen liberated from acrylic samplers. The presence of such oxygen traces could introduce an artifact by reacting with ferrous iron which may diffuse into peeper cells after deployment and perhaps precipitate as Fe(OH)3. Carignan observed that acrylic (Plexiglas) in equilibrium with the atmosphere can absorb 1.6% (vol/vol) O 2 that is lost slowly (half-time -5.7 d) once the material is exposed to anoxic conditions. This translated into significant differences in the observed porewater distributions of total Fe, Fe2+, dissolved reactive P, and SO42" between those samplers conditioned by a 30-d exposure to N 2 before deployment and those that received an 18-h bubbling in N 2 (Carignan et al., 1994). This phenomenon was originally described by Carignan (1984), who observed an orange discolouration on peepers deployed in anoxic sediments. The peeper preparation in this study included flushing periods greater than 72 h and suggests the potential for post-deployment releases of oxygen is minimal.  3.1.3.2 Peeper Deployment and Subsampling Each sampling session involved the deployment of four peepers at various sites within the lake. Peeper deployment for the first sampling period was achieved via diver insertion; peepers were pushed vertically into the sediments to a specified depth marked on the acrylic surface. Each sampler was marked via  32  a 6 m slack line from the peeper to a 20 kg anchor and a line from the anchor to a surface float. Subsequent peeper deployments and retrievals during ice-free periods were performed using weighted aluminum "landers". These cubeshaped support structures  (1 m x 1 m x 0.8 m) remained in situ during the  equilibration period with the peeper attached and ensured vertical insertion to a specific sediment depth. During the winter sampling period, peepers were fixed to 5 m long teflon-coated pipes and pushed vertically into the sediments through augered holes in the ice surface. A i m long peg attached horizontally to the pipe and placed flush with the ice ensured a correct insertion depth and maintained the position of the pole over the deployment period. All peepers were allowed to equilibrate for at least 14 days. The most important factors determining the equilibration time are the diffusion coefficient of the constituent, its sorptive properties, and the temperature and porosity of the sediments (Carignan, 1984). Carignan determined that for the porosities observed in recent lake sediments, equilibration periods of 15 days for warm ( 2 0 ° - 2 5 ° C) and 20 days for cold ( 4 ° - 6 ° C) appear to provide sufficient incubation. Peeper removal and subsampling occurred on a daily basis, with one peeper being removed per day. Upon retrieval, each peeper was brought to just below the lake surface, briefly agitated to dislodge adhering particulates, transferred to the boat and immediately placed in a nitrogen-purged sleeve for transport to the lab. The former procedures took less than 2 minutes. Peepers were subsequently processed in a nitrogen-filled glove bag and involved the careful removal of nylon screws, frit and face-plate, followed by the removal of the 0.45 urn Gelman filter (when used in tandem with the 0.2 um polycarbonate membrane). Each peeper was then subsampled by direct puncturing of the filter membrane with an Eppendorf pipettor and acid-washed pipette tips. The dual  33  row chamber arrangement of the peepers allowed one row to be subsampled for metals while the other for nutrients and sulphate. Direct probe insertion into chambers at selected depths allowed for measurements of porewater p H . Samples to be analyzed for dissolved metals, nutrients and sulphate were treated as described above for the water column protocol. Sample extraction and partitioning was accomplished within 2 hours of peeper retrieval.  3.1.4 Sediment Sampling  Coring was performed during the winter survey as it was hoped that the stable ice platform would facilitate the acquisition of high quality cores. Three such cores with well-defined sediment-water interfaces were collected through augered holes in the ice by lowering a light-weight gravity corer (Pedersen et al, 1985) suspended from a metre-block and tripod. Cores were carried by hand to the lab and immediately processed. Cores were first logged and then firmly attached to the underside of an extrusion table fitted with a nitrogen-filled glove-bag. The core was extruded through a hole in the center of the table and carefully sectioned; core slices were placed in acid-washed 250 mL high-density polyethlyne (HDPE) centrifuge bottles and spun at ~1,200 gravities for 20 minutes at which point they were transferred to an additional nitrogen-filled glove bag for decantation of the interstitial waters and filtration. Supernatant water from the centrifuged sample was decanted into an acid-washed 50 mL disposal polypropylene syring fitted with an acid-washed Syfril 0.45 um mixed cellulose acetate disposable filter. A n acid-washed polyethylene piston was utilized to express the interstitial solution through the filter and into an acid-washed HDPE sample bottle. The remaining  34  filtrate was passed through a 0.1 um disposable filter and acidified to p H 2 with ultra-pure nitric acid (Seastar Chemicals). Sediment samples were bagged and frozen until analysis.  3.2 Instrumentation 3.2.1 ICP/MS A  V G "PQ Turbo  Plus"  inductively-coupled  plasma  mass  spectrophotometer (ICP/MS) (VG Elemental, Surrey, UK) was used for the determination of the concentrations of Mn, Ni, Cu, Zn, As and Pb. Data analyses were handled by a Dell 486 computer equiped with PQ vision 4.1.1 software. Pure (99.998 %) ICP-grade argon (Medigas, Vancouver, B.C.) was used as the plasma and carrier gas.  A Rheodyne 6 port flow injection system (Mandel  Scientific, Guelph, Ont.), positioned between the peristaltic pump and the nebulizer, was used to facilitate small sample volumes. Operating conditions of the ICP/MS were optimized daily to obtain maximum sensitivity.  Resolution and peak shape were optimized using a  solution containing 10 ppb Co, In, Pb and Bi. Sample analyses were performed using pneumatic nebulization (de Gallen v-grove glass nebulizer).  The  instrument was run in a multichannel peak jump mode with a 10 milli-second dwell time, one point per peak and a 20 second aquisition time.  The  concentrations of the respective metals were calculated using a linear calibration curve derived from certified metal standards. A 10 ppb In internal standard was applied to all samples, standards and blanks to correct for sensivity variations during analyses. Typical operating conditions are outlined in Appendix A .  35  3.2.2 GFAAS  Graphite furnace atomic adsorption spectroscopy  (GFAAS)  was  performed using a Varian Spectra AA300 with a Zeeman background correction to determine Fe concentrations in the water column and porewaters.  All  determinations were conducted using L'Vov platforms. Concentrations were caculated from linear calibrations derived from certified metal standards using instumental parameters suggested by the manufacturer.  3.2.3 Ion Chromatography  Determinations of porewater and water column concentrations of nitrate and sulphate were performed using a DIONEX DX-100 ion chromatograph. The system provided efficient isocratic separation of the respective inorganic anions using a 1.8mM N a 2 C 0 3 / 1.7 m M NaHC03 eluent in conjunction with an IonPac® AS12A anion exchange column and a ASRS-1 anion self-regenerating suppressor.  3.2.4 X-ray Fluorescence  A n aliquot of each freeze-dried sediment sample was ground in a tungsten carbide disc mill prior to preparation for X-ray fluorescence, C-N-S and coulometric analyses. Metal concentrations were determined in the solid phase by X-ray fluorescence spectrometry using pressed powders and glass discs for minor and major elements, respectively. Calibration procedures and precision estimates for most elements were very similar to those reported by Calvert et al. (1985) with the exception that the calibration range for As was extended by  36  making artificial standards spiked with known weights of A S 2 O 3 . A linear calibration was obtained for As up to a maximum concentration of 5000 ppm. Accuracy of the elemental analyses was monitored by including aliquots of the standard sediment powders MESS-1, BCSS-1, PACS-1 with each batch of samples.  3.2.5 Coulometry  Carbonate carbon was determined photometrically in a Coulometrics Inc. 5010 coulometer.  Precision, as the relative standard deviation, was ± 3.7 %.  Reagent-grade CaC03 was used to assess routinely the accuracy of the coulometer.  3.2.6 Carbon, Nitrogen and Sulphur  Total C , N and S of sediments and suspended particulates were determined by combustion /gas chromatography using a Carlo-Erba (NA-1500) elemental analyser. Precisions were ± 0.6, 3.5 and 3.5 % (1 a rsd), respectively. Organic carbon was obtained from the difference between the total and carbonate carbon values (combined 1 a rsd = 1.9 %).  3.2.7 Spectrophotometry  Water column and porewater ammonium concentrations were determined spectrophotometrically similar to the alternative method outlined by Parsons et al. (1984). Ammonia reacts with phenol in an alkaline citrate medium with  37  sodium nirtoprusside added as a catalyst and forms blue indophenol. The absorbance is then measured at 640 run.  3.2.8 Digestion of Polycarbonate Filters  Polycarbonate membranes were placed in pre-weighed 5 mL teflon vials and oxidized at 80°C with 1 mL of ammonia (Seastar Chemicals). After being dried down (80°C) 0.5 mL each of HC1, HF and H N O 3 (Seastar Chemicals) were /  added and microwave-digested for 45 minutes at 60 PSI. After being dried down once again, the digest was brought up in 250 uL H N 0  3  and diluted to  approximately 4 mL. ICP-MS was used to determine the concentrations of Mn, Ni, Cu, Zn, As and Pb. Total Fe concentrations were determined via GFAAS.  38  IV. RESULTS  Results for the water column, interstitial waters and sediment components of the Balmer Lake survey are presented in this chapter. The quality assurance and quality control ( Q A / Q C ) for all measured parameters was excellent (Appendix B). Values obtained for field blanks, field replicates, laboratory blanks, laboratory replicates and reference materials all met acceptable criteria.  4.1 Water Column  4.1.1 Physical Profiling  Hydrographic data of temperature and dissolved oxygen suggest that the water column of Balmer Lake maintains fairly well-mixed conditions during icefree periods. The first sampling period in the summer of 1993 was chosen to coincide with a stratified water column, since stratification is common to many temperate lakes in the summer months. Basic hydrographic data, however, revealed fully oxygenated bottom waters and invariant temperature and p H profiles (Fig. 4.1). Temperatures ranged from 16-18 °C while dissolved oxygen concentrations ranged from 7-9 mg/L. Profiling of p H revealed a slightly acidic to slightly basic water column with values ranging from 6.2 to 7.8 p H units. The overall homogeneity was probably a result of the common windy conditions and the resultant mixing of the shallow water column (~4 m deep). Fall sampling took place subsequent to convective overturn and was also characterized by a well-mixed water column. Cooler water temperatures (~2-3°C) and higher dissolved oxygen concentrations (10-12 mg/L) were evident (Fig. 4.2). A p H  39  Summer Station 1 Temperature, deg. C 14  15  16  17  18  Summer Station 2 Temperature, deg. C  B  19  20  14  15  "  "  :  o i  1 "  16 ' ' 1 "  17 "  18  19  20  1 •'  f  :  / /  -  / "  ?  :  *  -  6  1 * 1 1  1  *  \ \  CL  &: To  i  i  ;  <D  \  pH-*-  1  A  5  1 1  i<—Temp.  1 1  o . . . .  5  6  7  8  Dissolved O  14  15  " ' "  16  17  18  1 . . . .  20  15  14  ;  t  i  :  P  1  H — » » ;  CL <D  1 f  1  •  \  •_  5  / -  '  +\ 9  18  19  y  A •  T  I I  CO  _ :  17  20  1  ? A  -  Tjjmp.-*^  3  5  16  I I \\ V I [ P H — H ,I II I ?\ I  r-°-.  /  CL  1 . . . .  ppm; pH  M I  -  ,  z >  *  -  i  4  1 . . . .  f  :  ;Temp.  1 . . . .  -  Summer Station 4 Temperature, deg. C  D  19  1 . . . .  Dissolved O  ppm; pH  Summer Station 3 Temperature, deg. C  i  10  10  9  -  1  1  [ -  ?f iI j  Vs/s/s/s/s/s/s/s/s/sSs/s/s/sM  i  6^*^  . . . . i . . . .  i....  i . . . .  i . . . .  i...  . . . .  .'  10  10  Dissolved 0 , ppm; pH  Dissolved 0 , ppm; pH  2  2  Fig. 4.1  0 1 • • • • 1 • • • • 1  Summer water column profiles of temperature, dissolved oxygen and pH for  stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, June, 1993. The hatched line represents the sediment surface. 40  B  Fall Station 1 Temperature, deg. C 0 0  1  2  3  4  5  6  Fall Station 2 Temperature, deg. C  7  1  2  3  4  1 1 1 1 1 1 1 1 1 • 11  •  o  5  6  7  11,1111111111111  r  •  H  :  M  Q_  5.  M  o  O  1  5  o '• Temp.—»>',  ' —  6  8  10  12  8  14  2  3  4  5  D  6  '  10  —  11  1  i— i — ' 13 14 15  —  12  3  2  • 111111111111  \  1 "  1  6  7  1111111111111  1  I  >  CL  Q_ <D  5  4 ' 1  \  0  -  0  a  Q  I  '  Fall Station 4 Temperature, deg. C )  7  I IT 1 | 1 1 1 1 | 1 1 I 1 | 1 1i i1 i 1| i| i i i | i i i i O  !  —  j  Dissolved O ppm  Fall Station 3 Temperature, deg. C 1  i  9  Dissolved O ppm  0  -  %  :  T5 2 -  2  5  : Temp.—•!  '  O  j  Temp.—»  -  l  o  -  •  -  'f s s / / sssssssss. s 6  8  10  12  '  8  14  •  1  10  Dissolved 0  Dissolved O ppm z  . . .  12 2  1 .  14  ppm  Fig. 4.2 Fall water column profiles of temperature and dissolved oxygen for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, October, 1993. The hatched line represents the sediment surface. 41  meter was unavailable for the fall and winter field sessions, and as a result, such data are absent. Winter profiling near the end of the ice-over period indicated a moderately stratified water column at the stations sampled (Fig. 4.3). Temperatures increased gradually from 0 °C immediately below the ice (~1 m thick) to 3-4 °C in the bottom waters.  Conversely, dissolved oxygen levels  decreased quickly in the lowermost half metre from approximately 5-7 m g / L to below detection limits 10-30 cm above the sediment surface (Fig. 4.3). However, at the shallowest station (station 3, 2.6 m depth) only a slight basal oxygen depletion was observed during this period (Fig. 4.3).  The early-spring sampling,  conducted less than two weeks after ice-off, revealed the reoccurrence of a wellmixed water column (Fig. 4.4). Constant profiles of temperature (9.5-10.5 °C) and dissolved oxygen (8-9 mg/L) were evident. Profiling of p H revealed a neutral to slightly basic water column with values ranging from 7.0-7.7 p H units.  4.1.2 Nitrate, Ammonium and Sulphate  Variations in lake-wide average concentrations of nitrate, ammonium and sulphate closely parallel one another throughout the seasonal sampling periods. For all three ions, the seasonal hierarchy of mean concentrations follows the order winter > spring > summer > fall. The well-mixed water column and welloxygenated bottom waters characteristic of the summer and fall field sessions, are consistent with the uniform profiles of dissolved constituents measured during these periods. Stations sampled during the summer period exhibited constant depth profiles of nitrate and sulphate with lake-wide values averaging  42  Winter Station 1 Temperature, deg. C 0  1  2  3  4  5  1111.  6  i i i  Winter Station 2 Temperature, deg. C  B 7  0  1  2  0  1  2  3  4  5  6  7  3  4  5  6  7  | < . 1 i  Ice  .A  S  J  J  J  J  J  J  J  J  J  J  J  J  J  J  '  0  1  0  2  3  4  5  6  [O ], ppm  Winter Station 3 Temperature, deg. C  Winter Station 4 Temperature, deg. C  1  2  3  ;:  4  Ice  5  6  1'"  • ! » • • •  a.  \  CD Q  0  1  2  3  4  5  3  4  5  :  SI a. o>  6  -  \  Ice  7  \ . •  ! hr* Temp.  i  Q  r  6  6  cf  '(3  -  "S*-—°  —  0  2  " I " '  L . i  f>—<v V  :  1  • I ' H • ! >< i i | , < > >| t i p< |  i  '  } / f  q»S  Temp  7  (  \  to  7  [O ], ppm  !o  5  '  I . . . .  0  7  1  1  2  3  4  ' • • •  11111  5  6  7  [O ], ppm  [OJ, ppm  Fig. 4.3 Winter water column profiles of temperature and dissolved oxygen for stations 1, 2, 3 and 4 (A-D, respectively), Balmer Lake, March, 1994. The hatched line represents the sediment surface. 43  Spring Station 1 Temperature, deg.C 5  6  7  8  9  Spring Station 2 Temperature, deg.C  B 10  11  5  •1 11 11 1 1 1 I|I 1ifi|I 1 11I Ii I[QiI I  f  f  t  9 :  j  j  •  f  '-  '-  E  f •  pH->J  '-  i  '-  •  6  7  8  9 <E  10  11  III I 1  '•  E  •  pH-  6  1  9  -i  f  Y°> 1 -  i  6  I 111 I I 11 111 I II  CO Q  2  CO  i  -  *  6  iemp.->i  1  i  JTemp.W?  •  *  -  6  vs///s//sssssss  SSSSss/ssssssss. • I I I  5  1  6  t t • I  7  1  I I  P  1  8  9  10  11  5  Dissolved O , ppm; pH  6  7  8  9  10  11  Dissolved 0 , ppm; pH 2  C  Spring Station 4 Temperature, deg.C 5 0  6  7  i iiiIiii iI  8  i  ir i |  9  10  1  • i 1  11  11 i y | i 111 [ Q  rr*  f -.  t  f  4  9 ~  -  :  2  pH^  i Temper  i  3  •  ?  •  6  .//////////  1 1 1 1  5  1  1 1 1 1  6  7  -  / /./ / s.  1 8  t t  9  10  11  Dissolved O , ppm; pH z  Fig. 4.4  Spring water column profiles of temperature, dissolved oxygen and pH for  stations 1, 2 and 4 (A-C, respectively), Balmer Lake, May, 1994. represents the sediment surface. 44  The hatched line  330 umol/L (274-342) and 3.30 mmol/L (2.69-3.45), respectively (Table 4.1). Ammonium was fairly uniformly distributed in the surface waters, with values ranging from 86-123 umol/L (Table 4.1). Depletions of ammonium with depth were evident for stations 2 and 3, where bottom water concentrations were approximately 60 % of their respective surface values (Table 4.1). At all sites and depths, phosphate concentrations remained below detection limits (~1 umol/L). Significant decreases in average ammonium concentrations were evident over the summer-fall transition with average lake values falling from 94 to 50 umol/L (Table 4.2). Water column nitrate and sulphate contents during the fall period remained consistent with the summer season and averaged 312 umol/L (292-346) and 3.41 mmol/L (3.32-3.69), respectively (Table 4.2). As seen for the summer period, phosphate concentrations were not detectable throughout the water column. The major ion distributions observed during the winter period show distinct relationships to the stratification indicated by the temperature and oxygen data (Table 4.3; Figs. 4.5 -4.7). Nitrate removal in the suboxic to anoxic bottom waters, and/or in shallow sediments, results in pronounced depletions at all stations 40-60 cm above the sediment-water interface (Table 4.3; Fig. 4.5). Subice nitrate values averaging 415 umol/L (372-435) decline to minima of 16 to 80 u m o l / L in the bottom waters.  Less significant reductions in nitrate  concentrations are observed in the 'deep' waters at shallow station 3, where values decrease to -175 umol/L (Fig. 4.5C). A large increase in the inventory of ammonium was evident during the winter sampling period (Fig. 4.6); contents in bottom waters at all locales except station 4, ranged from 260 to over 500 umol/L. Ammonium values were not obtained below 2.6 m at station 4, and as a result, the presence or absence of a bottom water increase was not determined.  45  Table 4.1. Summer water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb, N H 4 , NO3*, SO42- and P O 4 " for stations 1, 2, 3, and 4 in Balmer Lake, Ontario, June +  3  1993. All values are expressed in ug/L unless specified otherwise (BDL denotes below detection limits).  Station  Depth, m  Mn  Fe  0.5  271  35  2  287  41  2  285  3.5  Ni  Cu  Zn  As  Pb  NH4 mM  NO3 mM  SO4 PO4 mM mM  387  164  46  282  0.4  128  327  3.32  BDL  405  170  49  288  0.4  128  329  3.19  BDL  43  409  168  50  286  0.3  130  326  3.28  BDL  282  39  399  171  59  286  0.3  128  327  3.26  BDL  0.1  281  38  406  172  45  276  0.4  86  362  3.38  BDL  1  277  43  403  173  48  277  0.6  70  331  3.39  BDL  2  274  46  409  172  48  286  0.6  77  274  2.69  BDL  2  275  51  411  175  50  283  0.4  75  294  2.87  BDL  3.44  BDL  3  300  50  437  181  56  293  0.7  77  337  3.6  289  39  420  179  60  290  0.4  55  313  3.15  BDL  0.5  282  47  423  172  46  285  0.4  99  335  3.35  BDL  2  280  35  400  175  48  287  0.3  68  341  3.38  BDL  2  282  35  405  179  48  289  0.4  70  342  3.42  BDL  2.8  292  58  411  169  52  288  0.4  59  337  3.25  BDL  0.5  287  35  407  175  45  290  0.5  103  338  3.45  BDL  0.5  289  32  410  175  46  288  0.5  105  342  3.43  BDL  2  289  39  411  169  48  292  1  103  332  3.32  BDL  3.5  286  44  418  173  52  289  0.4  3.34  BDL  284  39  421  175  54  290  0.4  91 91  337  3.5  332  3.31  BDL  46  Table 4.2. Fall water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb, NH4+ N03 , SC»4 and PO43- for stations 1, 2, 3,4 and 5 in Balmer Lake, Ontario, -  2_  October, 1993. All values are expressed in ug/L unless specified otherwise (BDL denotes below detection limits).  Mn  Fe  Ni  Cu  Zn  As  PbNH4NO3 uM uM  SO4 PO4 mM uM  0.5 2 2 3  161 165 170 171  95 92 88  242 262 260 265  114 112 110 114  63 59 60 55  203 207 208 210  0.4 0.5 0.4 0.3  56  302 309 299 307  3.33 3.38 3.32 3.42  BDL BDL BDL BDL  05 1.5 3  171 158 166  89 93 101  278 273 269  121 121 119  48 63 47  214 208 211  0.3 0.3 0.4  51 49 48  346 310 311  3.44 3.41 3.36  BDL BDL BDL  05 0.5 15 1.8  186 180 185 196  79 90 84 99  281 283 283 288  121 124 123 126  63 58 59 55  203 203 207 210  0.5 05 0.6 0.5  55  317 306 313 292  3.47 3.46 3.42 3.42  BDL BDL BDL BDL  0.5 0.5 1.5 1.5 3  181 182 183 182 180  94 89 97 90 88  279 282 284 280 275  124 118 122 125 112  40 45 65 62 56  214 219 217 215 210  0.3 0.2 0.4 0.4 0.4  46 48 45  307 316 316 338 307  3.38 3.52 3.45 3.40  BDL BDL BDL BDL BDL  0.1 0.1  227 203  98 88  312 300  130 101  124 59  242 235  0.3 0.3  300 304  3.62 3.69  BDL BDL  Station Depth, m  47  45 51  55 56  Table 4.3. Winter water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb,  NH4+ NO3-, SO4- and PO4- for stations 1, 2,3 and 4 in Balmer Lake, Ontario, 2  3  March, 1993. All values are expressed in \ig/L unless specified otherwise (ND denotes not determined). Note that replicate samples were collected at a number of depths at all stations. F e N i  Cu  Z n A s  Ft!  404 396 285 299  i'A) 455  294 285 274 331  83 84 68 63  337 343 384 358  0.6 0.6 0.5 0.2  306  158 147 151 141  438 432 473 474  301  67  338  0.1  504 926  73 76  502 532  177 52  130 149  199 260  0.3 0.1  972 1043  48 52  554 578  46 46  158 161  283 289  0.2 0.4  502  312 311  131 139 133 129 134  523 512  450 444  68 63  386 365  0.1 0.1  115 73 119 133  503 425 467 407 493 503 532  438 289 314 216 100 58 59  77 64 65 63 121 178 180  392 346 345 270 209 282 298  0.4 0.3 0.3 0.2 0.2 0.6 0.4  347 325 348 334 358  177 170 175 173 167  468 445 471 457 487  314 302 307 287 310  67 66 74 61 67  353 337 354 324 349  1.1 1.1 0.3 0.9 0.9  362 369 342 667  165  525 539 524 451  328 320 388 382  65 64 67 135  349 345 381 410  0.4 0.3 0.3 0.4  Station Depth, m Mn 1.0 1.0 2.0 2.3 2.3 2.5 2.8 2.8 3.0 3.0 3.2 3.2 1.0 1.8 1.8 2.3 2.5 2.5 2.7 2.8 2.9 2.9 1.0 1.5 1.8 1.8 2.0 2.0 2.2 2.2 2.4 2.6  307 275 300 312 662 978 1015  160 152  48  NH4 NO3 SO4 PO4 uM uM mM uM 372  4.94  NIL)  121 133  409 415  5.19 5.17  ND ND  125 245 251 446  418  5.31  ND  288 174 173 16  8.09 13.79  ND ND  12.04  ND  211 208 211 213 190  395 390  5.02 4.97  ND ND  388 379  5.00 5.09  ND ND  214 248 406  282 185 82  4.92 8.66 11.63  ND ND ND  160 162 158  411 414 420  5.16 5.26 5.32  ND ND ND  160 162 170  385  5.20  ND  407  5.18  ND  198 263  394 175  5.13 4.93  ND ND  Station Depth, m 4  1.0 1.0 1.5 15 2.0 2.0 2.5 2.6 2.7 2.7 2.8  Mn  Fe  Ni  Cu  Zn  As  Pb  322 291 295  154  466 427 416  235 221 208  73 71 75  331 319 317  0.1 0.3 0.3  315 326 362 341 674  150 158  140 148 156 129  461 494 528 500 429  240 320 346 310 149  73 87 87 79 132  49  324 357 367 333 310  0.2 0.4 0.2 0.2 0.4  NH4 N 0  3  SO4  PO4  94  379  4.95  ND ND  89 89 91 118  425 436  5.27 5.38  ND ND  48  6.55  ND  Winter Station 2  B  Winter Station 1  • •I :  Q .  "O i_  c>  2 !>  To  5  « • « « 1 • • • • 1 • •  1  1  • • '  1  1  • * • •  0 100 200 300 400 500  0 100 200 300 400 500  Dissolved N O ', p m o l L"  1  Dissolved N O " , p m o l L"  1  Winter Station 3  1  Winter Station 4  D  Ice 1  t  Q .  "D  2  cs  \ \ \ \ \ \ \ \ \ \ \ \ \ 3 h  3 r I I •  1 1  I I  1  i i i i 1 i i i i 1 i • • • I • • • • I <  I I I I • • « • 1 I  0 100 200 300 400 500  0 100 200 300 400 500  Dissolved N O '  Fig. 4.5.  pmolL"  Dissolved N O ", u.mo.L"  1  1  Winter water column profiles of nitrate for stations 1, 2, 3 and 4 (A-D,  respectively), Balmer Lake, March, 1994.  Replicate samples are represented by double  symbols at specific single depths. The hatched line represents the sediment surface. 50  Winter Station 2  B  Winter Station 1  1t  Q .  CD X)  k. CD  9  CO  5  • I I  I  • • • I ' ' ' • 1 ' • • • I • • • '  0 100 200 300 400 500  Dissolved N H , +  4  umolL"  0  100 200 300 400  500  Dissolved N H , ixmolL" +  1  1  4  Winter Station 4  Winter Station 3  Ice 3  cL CD  T3  I  i  • • • ' 1 ' ' • ' I ' ' ' ' 1 ' ' ' '  0  1  ' • • •  100 200 300 400  0  500  Dissolved N H , nmol L ' +  500  Dissolved N H , |xmolL" +  1  W i n t e r water c o l u m n profiles o f a m m o n i u m for stations 1, 2, 3 and 4  respectively), B a l m e r L a k e , M a r c h , 1994.  1  4  4  F i g . 4.6.  100 200 300 400  (A-D,  R e p l i c a t e samples are represented b y double  s y m b o l s at specific single depths. T h e hatched l i n e represents the sediment surface.  51  Winter Station 2  B  Winter Station 1  A  o  Ice >  Q .  Q)  •o  o  •O *  2  CD CO  CO  <:  k \ \ V* V '  0  "  5  10  0  15  Dissolved S0 ", m m o l L " 2  1  Winter Station 3  Dissolved  Fig. 4.7.  5  10  "  I''  *  5  10  0  15  Dissolved  1  4  2" 4  15  , m m o l L "1 mmr\l-l  Winter Station 4  D  S0 ", mmol L" 2  "  Dissolved S 0  4  0  "  5  10  15  S0 ", m m o l L " 2  1  4  Winter water column profiles of sulphate for stations 1, 2, 3 and 4 (A-D,  respectively), Balmer Lake, March, 1994.  Replicate samples are represented by double  symbols at specific single depths. The hatched line represents the sediment surface. 52  The winter water column was also characterized by a significantly greater burden of sulphate, with values averaging 5.2 mmol/L immediately below the ice surface (Table 4.3; Fig. 4.7) . Pronounced bottom water enrichments of sulphate were observed at or below depths of 2.8 m throughout the lake (Fig. 4.7). In the deeper lake areas (stations 1 and 2), sulphate values increase to near 14 mmol/L at water depths greater than 2.9 m (Fig 4.7A-B). During spring sampling, the concentrations of the major ions exhibited little lake-wide variation throughout the water column with values averaging 145 umol/L for N H 4 + , 268 umol/L for N O 3 - and 3.47 mmol/L for S O 4 " (Table 4.4). 2  Due to analytical difficulties, phosphate concentrations were not determined for the winter and fall periods.  4.1.3 Suspended Organics  The seasonal water column concentrations of total suspended material (TSM), organic-C (SPOC), organic-N (SPON) and total particulate organic matter (POM) are shown in Table 4.5. A crude estimate of the particulate organic matter (POM) content was obtained from the relationship, [POM] = 1.8 x [POC] (Nriagu et al, 1981). The TSM content of the water column exhibited negligible variability over the summer-fall-winter transition with values averaging 3.2-4.4 m g / L (Table 4.5). However, the winter water column was characterized by bottom water enrichments of total suspended solids at stations 1 and 4. Significant increases in the concentrations of suspended material were observed in the spring, during which values averaged 11 m g / L (Table 4.5).  The fraction of  suspended particulates represented by POM over the summer, fall, winter and spring varied considerably, with contributions averaging 47, 24, 26 and 14 %, respectively (Table 4.5). During the well-mixed periods of the summer, fall and  53  Table 4.4. Spring water column distributions of dissolved Mn, Fe, Ni, Cu, Zn, As, Pb, N H 4 , NO3-, SO42- and P04 " for stations 1,2 and 4 in Balmer Lake, Ontario, May, +  3  1993. All values are expressed in u,g/L unless specified otherwise (ND denotes not determined).  Station Depth,  Mn  Fe  Ni  Cu  Zn  A"s  PS  m  NTT4 uM  NO3 uM  3O4 mM  PO4 uM  1  1 2 3  301 284 299  86 86 85  428 403 428  220 210 223  59 87 57  224 223 225  1.1 0.9 1.3  136 141 142  252 268 275  3.25 3.47 3.46  ND ND ND  2  1 2 2 3 3  250 257 257 244  82 91 81 92  350 380 384 358  190 202 201 191  57 57 55 60 60  216 217 219 229 226  1.1 1.7 1.7 1.4 1.6  139 144 144 144  259 265 270 273  3.34 3.36 3.42 3.45  ND ND ND ND ND  4  1 1 2 3.2  267 262 265 251  80 78 76 75  377 361 363 343  184 180 180 171  65 65 64 59  230 244 250 230  0.9 1.3 1.6 1.7  148 152 152 156  269  3.63  269 275  3.62 3.71  ND ND ND ND  54  Table 4.5. Concentrations of total suspended solids (TSS), particulate organic carbon (SPOC), particulate organic nitrogen (SPON), C rg:N weight ratio (C/N) and total 0  particulate organic matter (POM) in the water column of Balmer Lake, Ontario, for the summer, fall, winter and spring field sessions. TSS values were determined from the particulate mass on pre-weighed 0.45 urn Nuclepore filters, while all organics were determined from glass fibre filters. ND deontes not determined. Season  Station  Depth, m  TSS mg-LT  SPOC Hg-L"  SPON ug-Lf  Wt. Ratio  POM ug-L-  0.5 2.0 2.0 3.5  3.7 3.3 4.0 3.3  988 876  135 119  7.32 7.36  1778 1577  920  122  7.54  1656  0.1 1.0 1.0 2.0 3.0 3.6  3.7 4.2 4.2 3.9 4.7 3.6  772 883  103 116  7.50 7.61  1390 1589  893 762 1191  118 no 171  7.57 6.93 6.96  1607 1372 2144  0.5 2.0 2.8  3.2 2.1 4.0  1017  128  7.95  1831  1308  165  7.93  2354  1  Summer  Summer  Summer  1  2  3  1  1  Summer  4  0.5 2.0 3.5 3.5  3.8 2.4 5.5 4.7  923 856 1033 ND  114 107 130 ND  8.10 8.00 7.95 ND  1661 1541 1859 ND  Fall  1  0.5 2.0 3.0  3.9 4.4 4.2  540 521 552  81 77 82  6.67 6.77 6.73  972 938 994  Fall  2  0.5 1.5 1.5 3.0  4.5 5.5 4.4 4.6  ND ND ND ND  ND ND ND ND  ND ND ND ND  ND ND ND ND  Fall  3  0.5 1.5 1.8  4.0 4.0 4.1  ND ND ND  ND ND ND  ND ND ND  ND ND ND  Fall  4  0.5 1.5 2.8  3.3 4.0 6.2  515 508 ND  76 74 ND  6.78 6.86 ND  927 914 ND  55  1  Season  Station  Depth, m  TSS  SPOC  SPON  C/N  POM  Winter  1  1.0 2.0 2.25 2.5 2.75 3.0 3.15  3.0 2.4 2.6 2.3 2.6 4.1 6.3  376 336 ND ND 332 416 754  82 79 ND ND 77 69 138  4.59 4.25 ND ND 4.31 6.03 5.46  677 605 ND ND 598 749 1357  Winter  2  1.0 1.75 1.75 2.25 2.5 2.65 2.8 2.9  4.0 3.1 2.7 3.2 3.1 3.5 2.7 4.4  577 498 498 ND 530 ND 376 ND  137 131 131 ND 149 ND 112 ND  4.21 3.80 3.80 ND 3.56 ND 3.36 ND  1039 896 8% ND 954 ND 677 ND  Winter  3  1.0 1.5 1.75 2.0 2.2 2.4 2.55  3.0 3.2 2.2 2.7 3.1 ND  360 ND ND 372 ND ND 945  80 ND ND 84 ND ND 157  4.50 ND ND 4.43 ND ND 6.02  648 ND ND 670 ND ND 1701  Winter  4  1.0 1.5 2.0 2.5 2.6 2.7 2.8  2.6 3.2 2.5 2.6 2.0 3.5 5.6  300 ND 333 310 404 ND 722  63 ND 67 65 83 ND 110  4.76 ND 4.97 4.77 4.87 ND 6.56  540 ND 599 558 727 ND 1300  Spring  1  1.0 2.0 3.0  8.2 7.9 9.3  798 731 743  142 135 139  5.62 5.41 5.35  1436 1316 1337  Spring  2  1.0 2.0 2.0 3.0  ND ND ND ND  749 706 700 739  137 135 134 137  5.47 5.23 5.22 5.39  ND ND ND ND  Spring  4  Spring  5  1.0 1.0 2.0 3.2 0.5 0.5  10.8 11.9 12.1 12.7 11.2 10.9  816 ND 834 823 913 867  138 ND 144 142 151 148  5.91 ND 5.79 5.80 6.05 5.86  1469 ND 1501 1481 1643 1561  56  spring, concentrations of suspended POC and P O N varied only slightly with depth. Values for SPOC over these seasons averaged 960 u g C / L (762-1308), 530 HgC/L (508-552) and 815 u g C / L (731-913), respectively (Table 4.5). Furthermore, concentrations of SPON over the summer, fall and spring periods averaged respectively 125 u g N / L (103-171), 78 UgN/L (74-82) and 142 u g N / L (135-151) (Table 4.5).  The discrepancy in the ranking of magnitudes in the SPOC and  SPON values over these periods can be explained by seasonal differences in their respective particulate C / N ratios (Table 4.5). With the exception of station 2, all sampling sites during the winter period exhibited bottom water enrichments of SPOC and SPON (Table 4.5). The bottom waters of stations 1, 3, and 4 were characterized by higher C / N ratios and SPOC/SPON concentrations over two times greater than their respective near-surface values.  4.1.4 Trace Metals  4.1.4.1 Dissolved Fraction  The well-mixed conditions of the summer, fall and spring water columns, revealed by physical profiling of temperature and dissolved oxygen, are in good agreement with the constant vertical distributions of dissolved trace metals observed during such periods. Concentrations of dissolved species during the summer period averaged 286 |ig/L for Mn, 40 u g / L for Fe, 411 u g / L for N i , 173 u g / L for Cu, 51 ug/L for Zn and 287 ug/L for As (Table 4.1). During all seasons and at all depths, concentrations of dissolved Pb hovered at or below its detection limit of approximately 0.5 Ug/L. The fall sampling period was characterized by generally lower dissolved metal inventories with values  57  averaging 175 u g / L for Mn, 91 u g / L for Fe, 275 u g / L for Ni, 119 u g / L for Cu, 56 u g / L for Zn and 210 u g / L for As (Table 4.2). The winter period exhibits marked differences in water column structure with respect to dissolved trace metals. In general, concentration profiles relate to the observed stratification (Table 4.3, Figs. 4.8-4.11). At depths shallower than the oxycline (approximately 2.5 m), dissolved metal concentrations are significantly greater than those observed during other seasons with values averaging 323 u g / L for Mn, 152 u g / L for Fe, 468 u g / L for Ni, 307 u g / L for Cu, 70 u g / L for Zn and 348 ug/L for As (Table 4.3; Figs. 4.8-4.11). Dissolved trace metal profiles in the sub- or anoxic bottom waters of Balmer Lake contrast greatly with those seen during other periods.  At all  stations, dissolved Mn and Zn are characterized by bottom water enrichments up to 3.2 and 2.7 times their respective near-surface concentrations (Table 4.3; Figs. 4.8-4.11). Conversely, dissolved C u profiles suggest removal for this element in the bottom waters at all locales except at the shallowest site, station 3 (Table 4.3). Depletions of dissolved copper were more pronounced in the deeper sites (stations 1 and 2; Figs. 4-8 and 4-9), where bottom concentrations represent 16 and 13 % of their respective near-surface levels. The bottom water gradients evident for M n and Zn occur across the corresponding interval for dissolved oxygen depletion (see Fig. 4.3); similarly, zones of decreasing Cu concentrations are consistent with that of dissolved oxygen. The water column distributions for As and N i are more complex. At the two deepest stations (stations 1 and 2, Figs. 4.8 and 4.9, respectively) concentrations of dissolved As begin to decrease between 2-2.5 m, reaching minima at approximately 2.7-2.8 m depth. At slightly greater depths, arsenic levels increase sharply to values in bottom waters  approximately 100 u g / L  greater than their respective minima (Figs 4.8 and 4.9). At the shallowest site  58  Winter Station 1  Winter Station 1  Winter Station 1  c. a •c i_ a  CD T3  k_ CD OS  5 3 L » » • • I ' • • •  200  400  Dissolved [Ni], p p b  60(  0  100  1  • • ' '  200  1  •  300  1  ' ' I  1  1  400  Dissolved [Cu], ppb  0  100  Dissolved [Zn], p p b  Fig. 4.8. Winter water column profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 1, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface. 59  200  Winter Station 2  Winter Station 2  Winter Station 2  B  ICE 1  V  ->—»  CL  0)  Q .  CD TJ  o  k_  CD  o  •o '  2 f  CD  CD CO  I  5  i i i I i  0  100  200  200  400  600  0  800 1000  Dissolved [Mn], ppb  D i s s o l v e d [Fe], p p b  200  300  400  Dissolved [As], ppb  Winter Station 2  Winter Station 2  Winter Station 2  100  ICE  E < Q .  CD  •o  a  o  CD "O >_ CD  CD CO  «  .,  0  200  400  D i s s o l v e d [Ni], ppb  F i g . 4.9.  60)  0  100  200  300  400  Dissolved [Cu], ppb  0  . i . . .  . i  Dissolved [Zn], ppb  W i n t e r w a t e r c o l u m n profiles o f d i s s o l v e d F e , M n , A s , N i , C u and Z n ( A - F ,  respectively) f o r station 2 , B a l m e r L a k e , M a r c h , 1994. by d o u b l e s y m b o l s at s p e c i f i c single depths. surface.  60  111..  100  Replicate s a m p l e s are represented  T h e hatched line represents the sediment  200  Winter Station 3  CD  •S  2  Winter Station 3  Winter Station 3  r  52 CO  5  0  100  Dissolved [Mn], ppb  D i s s o l v e d [Fe], p p b  Winter Station 3  0  200  400  D i s s o l v e d [Ni], ppb  0  0 200 400 600 8001000  200  Winter Station 3  c fc  60i'  0  100  200  300  100  Winter Station 3  c  400  300 400  Dissolved [As], ppb  r  Dissolved [Cu], ppb  200  0  100  Dissolved [Zn], ppb  Fig. 4.10. Winter water column profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 3, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface. 61  200  Winter Station 4  Winter Station 4  Winter Station 4  B  Ice 3L  Q. CD  ,  i  i  CD  A  W  •  0  100  200  0  200  400  Dissolved [Ni], ppb  W  1  • •  1  W  1  1  *  W  1  1  1  1  W  • '  0  0 200 400 600 8001000  p ^  60)  Winter Station 4  0  100  200  300  100  c '  400  Dissolved [Cu], ppb  200  300  Winter Station 4  0  100  Dissolved [Zn], ppb  Fig. 4.11. Winter water column profiles of dissolved Fe, Mn, As, N i , Cu and Zn (A-F, respectively) for station 4, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. The hatched line represents the sediment surface.  62  400  Dissolved [As], ppb  Dissolved [Mn], ppb  Dissolved [Fe], ppb  Winter Station 4  W  200  (station 3, Fig. 4.10), profiles of dissolved As lack a defined minimum, although values increase slightly below 2.2 m. Conversely, the water column at station 4 is characterized by a near-bottom maximum of dissolved As at ~2.7 m, below which concentrations decrease (Fig. 4.11). Bottom water maxima are also evident for dissolved Fe, Ni, Cu and to a lesser degree Zn, at the same horizon at this site (Fig. 4.11). Dissolved N i distributions during the winter period differ markedly between sampling sites. Stations 1, 3 and 4 all exhibit some increase with depth (Figs. 4.8, 4.10 and 4.11); values at stations 3 and 4, however, fall back to nearsurface levels below deep maxima (Figs. 4.10 and 4.11). Concentrations of dissolved N i remain fairly constant at station 2, except for a minimum evident between 2.5 and 2.8 m (Fig. 4.9). Distributions of Fe share similar inconsistencies to those seen for dissolved nickel. The steady drop in dissolved Fe below 2.5 m at station 1 occurs across the same interval observed for decreases in dissolved As and Cu (Fig. 4.8). The iron profile is characterized by a bottom water minimum at station 2 coincident with the As minimum observed at this site (Fig. 4.9). In contrast, stations 3 and 4 exhibit fairly uniform profiles of dissolved Fe (Figs. 4.10 and 4.11). During the well-mixed conditions of the spring sampling period, concentrations of dissolved metal species varied little throughout the water column with values averaging 265 u g / L for Mn, 80 u g / L for Fe, 372 u g / L for N i , . 196 u g / L for Cu, 61 u g / L for Zn and 224 ug/L for As (Table 4.4). In general, the magnitudes fall within the ranges observed for the summer session.  63  4.1.4.2 Particulate Fraction  The seasonal concentrations of particulate trace metals in the water column of Balmer Lake have been expressed in terms of both mass of metal per unit volume (Table 4.6) and mass of metal per unit mass of total particulate matter (Table 4.7).  The well-mixed periods of summer, fall and spring, are  characterized by considerably more intra-site variability in comparison to associated dissolved trace metal distributions. In general, the particulate fraction for most metals (except Fe) contributes a small percentage of the total pool of metals in the water column during all seasons. The water column concentrations of particulate Fe, Mn, Ni, Cu, Zn and As during the summer period averaged 107, 3, 5, 13, 27 and 7 ug-L , respectively, -1  and represented 71,1,1, 7, 30 and 2% of the respective totals (Table 4.6). These lake-wide averages translate into mass ratios of 29,600 mg-kg for Fe, 850 mg-kg" _1  1  for Mn, 1,330 mg-kg- for N i , 3,250 mg-kg- for Cu, 7900 mg-kg" for Zn and 1  1  1  1760 mg-kg for Pb (Table 4.7). The autumn particulate data contrast little from -1  the summer values; concentrations of particulate Fe, Mn, N i , C u , Z n and As averaged 205, 4, 5, 15, 16 and 9 ug-L" , respectively (Table 4.6). 1  The fall  proportions of the dissolved and solid-phase fractions show additional interseason similarity with the particulates averaging 69, 2, 2, 11, 21 and 4 % of the respective totals of Fe, Mn, Ni, Cu, Zn and As. The influence of ice cover on the chemical and physical environment during the winter months can be further realized upon observation of the vertical distributions of particulate metal species (Tables 4.6 and 4.7). Slight decreases in dissolved Fe with depth are mirrored by significant increases in concentrations of Fe-bearing particulates which surpass 400 ug-L" at some sites (Table 4.6). 1  Conversely, Mn, N i , Cu and Zn particulates exhibit minimal variation with  64  Table 4.6. Particulate trace metal concentrations of Fe, Mn, Ni, Cu, Zn, As and Pb for the summer, fall, winter and spring water columns in Balmer Lake, Ontario. All values are expressed in ug-L~l.  Season  Station  Water depth, m  Fe  Summer  1  0.5 2.0  Summer  2  Mn  Ni  Cu  Zn  As  Pb  105.2  1.5  3.1  6.8  56.4  3.8  0.1  108.1  2.3  3.8  9.1  26.3  3.9  0.2  2.0  126.6  2.8  4.5  11.9  3.0  5.1  0.3  3.5  102.2  2.0  3.2  8.9  28.1  4.0  0.3  0.1  99.3  2.5  5.3  11.6  93.9  6.3  0.2  1.0  118.0  1.0  95.9  2.8  4.7  11.1  9.0  6.1  0.1  8.0  0.2  2.0  138.0  Summer  Fall  Fall  Fall  Fall  3  4  1  2  3  4  6.2  15.8  16.3  22.2  50.1  43.9  35.8  1.8  3.6  122.9  3.5  4.9  11.8  4.5  6.6  0.2  3.0  Summer  3.8  5.1  0.5  88.7  1.5  2.5  7.1  28.2  3.2  0.2  2.0  68.1  1.4  2.3  6.0  25.9  2.5  0.2  0.5  101.5  2.0  3.7  9.9  33.0  4.1  0.4  2.0  70.5  1.6  2.6  6.7  17.1  2.8  0.2  3.5  147.0  2.9  4.3  11.3  7.3  4.5  0.3  0.5  182.0  3.2  4.2  12.5  14.6  7.3  0.9  2.0  212.6  3.6  4.8  14.5  43.1  8.8  1.1  3.0  203.7  3.4  4.5  12.8  20.8  8.0  0.9  0.5  182.3  3.3  4  6.9  7.6  0.9  214.3  3.9  -7 5.5  14.8  1.5  17.7  10.8  9.5  1.3  1.5  195.2  3.4  4.9  15.5  8.7  8.1  1.0  3.0  219.8  3.7  5.3  16.9  7.4  9.2  1.1  0.5  213.6  4.1  5.7  17.5  9.0  10.2  1.2  1.5  193.4  3.4  4.8  14.6  7.3  8.4  1.1  1.8  192.0  3.4  4.8  14.5  6.6  8.4  1.0  0.5  171.0  2.8  1.5  189.1  3.3  2.8  283.2  4.5  65  '  3.8  11.7  16.5  7.2  0.9  4.5  13.2  14.9  8.3  1.0  5.3  14.8  38.4  9.4  1.1  Season  Station  Winter  1  Winter  Winter  Winter  Spring  Spring  2  3  4  1  4  Depth  Fe  Mn  Ni  Cu  Zn  As  Pb  1.0  130.8  1.6  5.6  32.1  6.3  11.6  0.8  2.0  156.3  2.0  7.7  36.0  13.5  9.5  0.9  2.25  157.0  2.6  9.6  50.8  8.6  11.4  0.7  2.5  160.5  2.6  8.9  42.6  7.0  11.7  0.7  2.75  149.4  2.4  6.5  51.2  5.6  11.7  0.5  3.0  388.3  3.0  6.4  22.6  6.9  121.4  0.6  3.15  366.2  6.2  14.8  38.0  16.3  42.3  0.7  2.0  11.7  109.6  15.5  6.4  0.2  1.0  205.5  1.75  148.7  2.25  193.6  2.5  13.4  115.2  14.9  16.0  1.3  2.5  223.9  2.6  15.4  97.4  18.2  14.0  1.3  2.65  222.3  2.7  15.6  81.6  18.8  14.0  1.3  2.8  175.3  2.4  12.6  59.5  10.7  0.8  2.9  424.4  3.4  13.1  53.3  19.1 19.4  110.9  1.4  1.5  161.5  2.2  7.8  45.5  31.7  12.3  0.9  50.1  12.9  0.8 0.8  1.75  153.3  2.1  7.7  45.4  1.0  163.9  2.6  7.2  29.1  8.7  11.4  1.5  176.0  2.8  7.6  31.3  11.8  12.2  1.0  2.0  151.1  2.2  6.1  25.2  3.9  10.0  0.7  2.5  143.3  2.0  6.0  27.6  16.3  8.8  0.8  2.6  130.1  1.8  5.5  32.8  11.3  8.4  0.7  2.7  190.2  2.5  8.6  54.5  16.7  13.5  1.1  2.8  479.9  4.5  11.2  49.7  73.9  116.1  2.3  1.0  445.5  5.7  7.5  70.7  22.6  49.2  0.8  2.0  566.0  7.2  9.4  88.5  29.9  70.8  1.2  3.0  569.6  7.5  10.2  90.8  27.5  69.0  1.0  1.0  733.0  10.2  11.2  101.1  30.6  73.4  1.4  1.0  739.3  11.3  12.5  105.9  31.3  75.8  1.4  2.0  744.5  11.1  12.1  105.2  33.7  78.3  4.7  3.2  742.3  11.4  11.6  96.8  27.3  69.1  1.3  66  Table 4.7. Concentrations of trace metals (mgkg'l), organic carbon (%) and organic nitrogen (%) in suspended particulates of the summer, fall, winter and spring water columns in Balmer Lake, Ontario. Trace metal values were determined from the particulate mass on pre-weighed 0.45 um Nuclepore filters, while all SPOC and SPON values were determined from glass fibre filters. Season  Station  Summer  1  Summer  2  Water depth, m  Fe  Mn  Summer  Fall  Fall  Fall  Fall  4  1  2  3  4  Zn  As  Pb  SPOC  SPON  0.5  28,510  410  840  1850  15380  1030  27  26.8  3.7  32,560  690  1130  2730  8010  1160  70  26.4  3.6  2.0  31,320  700  1100  2940  830  1270  65  3.5  31,290  620  970  2730  8700  1230  77  28.2  3.7  0.1  26,770  690  1420  3120  25410  1690  60  20.8  2.8  1.0  28,190  21.1  2.8  1.0  22,590  670  1100  2610  2180  1440  26  2.0  35,790  980  1610  4100  1390  2070  50  23.2  3.1  3480  4750  10710  9450  7670  380  16.3  2.4  34,020  970  1350  3270  1320  1840  52  33.0  4.7  0.5  27,780  460  790  2230  8930  1020  66  31.8  4.0  2.0  33,060  680  1140  2900  12720  1220  110 32.9  4.1  3.6 3  Cu  2.0  3.0  Summer  Ni  0.5  26,940  520  990  2620  8840  1100  95  24.5  3.0  2.0  29,300  670  1100  2790  7210  1150  67  35.6  4.4  3.5  26,600  520  780  2050  1380  810  54  18.7  2.4  0.5  46,590  810  1060  3190  3870  1870  220  13.8  2.1  2.0  47,940  810  1080  3270  9830  1970  260  11.7  1.7  3010  5030  1890  210  13.0  1.9  3.0  48,040  790  1060  0.5  40,760  740  1060  3310  1650  1710  210  1.5  38,750  710  990  3200  2040  1710  230  1.5  44,110  770  1100  3510  2080  1830  230  3.0  47,540  810  1140  3660  1710  1980  240  0.5  53.940  1030  1440  4410  2400  2590  310  1.5  48,730  850  1210  3680  1970  2120  270  1.8  46,650  810  1180  3530  1720  2040  240  0.5  52,450  870  1170  3580  5220  2220  260  15.8  2.3  1.5  47,790  840  1130  3350  3890  2100  260  12.8  1.9  2.8  45,620  720  860  2380  6260  1520  170  67  Season  Station  Winter  1  Winter  2  Depth  Fe  Mn  Ni  Cu  Zn  As  Pb  SPOC  1.0  42,960  520  1850  10540  3340  3800  260  12.4  2.7  2.0  64,660  810  3170  14880  7200  3950  360  13.9  3.3  2.25  59,280  970  3620  19190  4730  4300  260  2.5  71,220  1140  3970  18920  4840  5190  300  2.75  56,490  900  2480  19370  3600  4410  190  12.6  2.9  3.0  95,010  740  1580  5540  2650  29700  140  10.2  1.7  3.15  58,520  990  2360  6080  3220  6760  120  12.0  2.2  1.0  51,750  510  2960  27600  4890  1600  40  14.5  3.4  1.75  54,380  16.0  4.2  18.2  4.8  16.8  4.7  14.2  4.2  17.3  3.9  11.8  2.5  1.75  Winter  3  2.25  60,960  800  4210  36290  5920  5030  400  2.5  71,120  830  4890  30920  7020  4460  410  2.65  64,240  780  4510  23570  6560  4050  380  2.8  66,020  920  4760  22410  8650  4020  320  2.9  96,040  770  2960  12070  5270  25100  310  1.5  54,730  730  2640  15400  12080  4170  300  1.75  47,430  660  2380  14050  16720  4000  250  2 Winter  Spring  Spring  4  1  4  SPOIS  1.0  64,220  1020  2840  11400  4920  4480  330  1.5  55,520  890  2390  9880  4940  3840  300  2.0  61,120  910  2480  10180  3150  4030  290  13.5  2.7  2.5  54,790  780  2280  10570  7730  3350  320  11.9  2.5  2.6  65,370  890  2740  16490  7630  4220  370  20.3  4.2  2.7  54,690  720  2480  15680  5940  3870  330  2.8  85,830  800  2010  8880  13920  20770  420  12.9  2.0  1.0  54,650  700  930  8670  3000  6030  100  9.8  1.7  2.0  71,280  900  1180  11150  4000  8910  150  9.2  1.7  3.0  61,080  800  1090  9740  3160  7400  110  8.0  1.5  1.0  67,680  940  1040  9340  3000  6780  130  7.5  1.3  1.0  61,990  950  1050  8880  2780  6360  120  2.0  61,500  920  1000  8690  2940  6470  390  6.9  1.2  3.2  58,450  890  910  7620  2300  5440  100  6.5  1.1  68  depth (Table 4.6). Profiles of particulate As species are characterized by large bottom water spikes which range from 111-120 ug-L" (Table 4.6). 1  The spring period exhibited the greatest concentrations of metal-bearing particulate phases, with values for Fe, Mn, N i , Cu, Zn and As averaging 649, 9, 11, 94, 29 and 69 |igL"l. In addition, the total burdens of these metals in the water column received relatively greater contributions from their respective particulate phases; percentages for Fe, M n , N i , C u , Z n and As averaged, respectively, 89,3,3,33,31 and 23 % (Tables 4.6 and 4.7).  4.2 Interstitial Waters  4.2.1 Peepers  4.2.1.1 Nutrients and Sulphate  Two pairs of tandem peepers, set approximately 15 m apart in Balmer Lake, were emplaced at stations 1 and 2 for the first deployment during the summer session. The use of two peepers at each site was designed to assess the reproducibility of the dialysis arrays, as well as the homogeneity of the sediments over short lateral distances. A coarse examination of the porewater profiles indicated that the samples collected by the arrays provided consistent representations of the porewater constituents (Fig. 4.12). The porewater results of all parameters are presented in Appendix C. Examination of the duplicate profiles acquired 15 m apart at each site during the summer period reveals marked similarity in the distributions of nitrate, ammonium and sulphate. Values for these constituents in the peepersampled bottom waters remain fairly uniform above the sediment-water  69  Fig. 4.12. Duplicate summer peeper profiles of dissolved  NFL4+, NO3"  and SO^- for  station 1 ( A - C ) and station 2 (D-F), Balmer Lake, June, 1993. Replicate samples are represented by double symbols at specific single depths. 70  interface (Fig. 4.12) and closely match those concentrations measured from pumped samples drawn independently during the peeper deployment period (compare Table 4.1 with Appendix C). At both stations, steep porewater ammonium gradients are evident immediately below the sediment-water interface where they approach maxima of 600-700 umol'L" at ~ 10 cm (Fig. 4.12). 1  Nitrate profiles show sharp spikes at, or just above the interface, below which values decrease precipitously to low levels at shallow sediment depths (Fig. 4.12). Larger interfacial increases and steeper porewater gradients characterize the distribution at station 1. As for nitrate, sulphate concentrations decrease with depth, although less abruptly; levels fall gradually throughout the sampled sediment column, the profiles being concave-downward (Fig. 4.12). The peeper samples collected in the fall yield nitrate, ammonium and sulphate profiles similar to those observed during the summer period. The duplicate peeper arrangement repeated at station 2 again demonstrates the good consistency between profiles 15 m apart (Fig. 4.14). At all three sites sampled (stations 1, 2 and 5), porewater nitrate values plummet dramatically at shallow sediment depths, reaching very low concentrations at ~5 cm (Figs. 4.13 and 4.14). Sulphate distributions for stations 1 and 2 exhibit good agreement; concavedown profiles decrease gradually throughout the sampled sediment column from ~3.6 to ~0.5 m m o l L ' l . Peeper data for the shallow site in "Tailings Bay" (station 5), however, show evidence for a bottom water sulphate enrichment before sharply decreasing at, or slightly above, the sediment-water interface to a minimum of ~2 mmolL"!. The ammonium profile obtained for this site relates closely to that observed for sulphate; concentrations are enriched in the lowermost bottom waters and are depleted at, or below, the sediment surface (Fig. 4.13D).  Concentrations of ammonium do, however, increase in the  porewaters below ~ 10 cm depth at this site. Stations 1 and 2 show similar  71  Dissolved [ N H / ] , u m o l L  Dissolved [NH ], jimolL+  4  1  1  Dissolved [ N 0 ] , a m o l L  1  Dissolved [NO,"], umolL-  Dissolved [ S 0 ] , mmolL-  1  Dissolved [ S O / ] , m m o l L  1  2  3  4  1  Fig. 4.13. Fall peeper profiles of dissolved NH4+, NO3" and SO^-' for stations 1 (A-C) and 5 (D-F), Balmer Lake, October, 1993. Replicate samples are represented by double symbols at specific single depths.  72  Fig. 4.14.  Duplicate fall peeper profiles of dissolved NH4+, NO3" and S04^- (A-C,  respectively) for station 2, Balmer Lake, October, 1993. Replicate samples are represented by double symbols at specific single depths. 73  consistency with respect to ammonium distributions.  Steep porewater  ammonium gradients are evident immediately below the sediment-water interface where they approach maxima of 400-600 umolL"! (Figs. 4.13A and 4.14A). Peeper values for nitrate, sulphate and ammonium in the fall bottom waters closely match those concentrations measured from pumped samples drawn independently during the peeper deployment period (compare Table 4.2 with Appendix C). The bottom water distributions for nitrate, sulphate and ammonium determined from peeper sampling during the winter period (Figs. 4.15 and 4.16) accentuate nicely the values measured from independently drawn water column samples.  At the deeper locales (stations 1, 2 and 4), nitrate profiles exhibit  negligible inter-site variation, with values decreasing sharply in the lower bottom waters to minima 5-10 cm above the sediment surface (Figs. 4.15 and 4.16). At the shallow site in "Tailings Bay" (station 6, 2 m depth), nitrate levels decrease immediately below the interface; concentrations of the ion approach zero below 20 cm depth (Figs. 4.15 and 4.16). The deeper stations (stations 1, 2 and 4) exhibit marked similarity with respect to winter distribution of sulphate. Stations 1 and 2 show pronounced bottom water maxima of ~ 11-14 mmolL"! 5-10 cm above the sediment surface, below which their concave-down profiles decrease progressively to < 0.5 mmol-L'l in the lowest sampled horizons (Fig. 4.15). Station 4 is marked by a less intense bottom water enrichment and occurs in vicinity of the interface. Sulphate concentrations in the lower three decimetres of the shallow water column at station 6 vary insignificantly. Levels initially decrease at the sedimentwater interface and continue to decline to a depth of 20 cm where values approach a minimum of ~2 mmol-L-1 (Fig. 4.16).  74  Winter Station 1  Dissolved [NH ],nmolL+  W  t  e  r  p  ^  t  2  1  4  Dissolved [N0 ], u m o l L  p '  1  3  Fig. 4.15 Winter peeper profiles of dissolved NH4+, NO3" and  Winter Station 2  Dissolved [ S 0 ] , mmolL" 2  4  for stations 1 (A-  C ) and 2 (D-F), Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. 75  1  4  Winter Station 2  p  Winter Station 1  Dissolved [ S 0 1 , mmolL"  1  3  Winter Station 2  +  n  Dissolved [N0 ], u m o l L  1  4  Dissolved [NH ], u m o l L  i  Station 1  b  1  Winter Station 6  Winter Station 6  Winter Station 6  100 200 300 400 500 600  Dissolved [NH *], nmol L  1  Dissolved [NO, ], j i m o l L  1  F i g . 4 . 1 6 . W i n t e r peeper p r o f i l e s o f d i s s o l v e d N H 4 + , NO3" and S04^~ C ) and 6 ( D - F ) , B a l m e r L a k e , M a r c h , 1994.  Dissolved [S0 ], mmolL" z  4  f o r stations 4 ( A -  R e p l i c a t e samples are represented b y double  s y m b o l s at s p e c i f i c single depths.  76  The winter bottom water and porewater ammonium profiles at stations 1, 2 and 4 are quite similar.  Concentrations at all three stations increase  substantially in the lower bottom waters to values ranging from ~ 300-580 mmolL"! (Figs. 4.15 and 4.16). Stations 1 and 2 are characterized by bottom water maxima -10 cm above the sediment-water interface.  Ammonium  concentrations at station 1 steadily decrease below this depth, while those at station 2 are variable near the interface; below the interface at station 2 levels increase gradually through the top decimetre (Figs. 4.15 and 4.16). Below the bottom water maximum at station 4, the ammonium content is invariant to ~15 cm in the sediments, and steadily declines to the base of the profile. Conversely, the ammonium distribution at station 6 is characterized by a pronounced bottom water depletion in the top decimetre above the sediment surface. Concentrations decrease gradually to shallow sediment depths, below which levels increase in a concave-down profile (Fig. 4.16). The duplicate peepers deployed at station 1 during the spring period yielded good precision with respect to dissolved nutrients and sulphate. The nitrate profiles are similar to those observed at this locale during the summer and fall; uniform bottom water values decline precipitously in the top 10 cm to minima of < 10 umolL"! (Fig. 4.17). Ammonium distributions are also similar to those seen during ice-free periods (Fig. 4.17). However, sulphate profiles differ greatly from those seen during other seasons. Concentrations increase sharply from uniform bottom water values of -3.5 mmolL"! to subsurface maxima of ~6 mmol-L"l at 5-10 cm depth, and decrease rapidly below this horizon to values that approach 1 mmol-L"l (Fig. 4.17).  77  A  Spring Station 1  Spring Station 1  B  Fig. 4.17. Duplicate spring peeper profiles of dissolved NH4" ", NO3" and S04^- (A-C, 1  respectively) for station 1, Balmer Lake, May, 1994. Replicate samples are represented by double symbols at specific single depths. 78  4.2.1.2 Trace Metals  Porewaters were analyzed for a suite of dissolved trace metals including Mn, Fe, N i , C u , Z n , As and Pb. Throughout the seasons and at all depths sampled, dissolved Pb concentrations generally hovered at or below its detection limit of ~ 0.5 ug'L" . Consequently, plots of porewater Pb distributions are not 1  presented; measured values, however, are reported in Appendix C. Examination of the respective trace metal profiles suggest that some discrete samples may have been contaminated. Contamination could have resulted from the addition of small amounts of tailings to the porewater samples during peeper subsampling; the subsequent addition of acid would then have dissolved such metal-rich particles, yielding high dissolved metal concentrations. Given that diffusion acts to reduce concentration differences relatively quickly over short spatial scales, the principal criterion used here to define suspected contamination of samples is an obvious lack of consistency of the high concentrations with lower values measured in adjacent horizons. Such single-sample "spikes" are rejected from subsequent interpretation. Trace metal distributions observed during the summer period at stations 1 and 2 are very similar (Figs. 4.18 and 4.19); reasonable agreement also exists between the intra-site measurements.  Profiles are generally characterized by  uniform distributions above the sediment-water interface and sharp sub-surface concentration gradients in the shallow porewaters. Dissolved Mn concentrations increase sharply below the sediment-water interface to concentrations as high as -800-1100 ug'L" between 10-15 cm depth (Figs. 4.18 and 4.19). Below depths of 1  10 cm, M n levels remain fairly constant. Dissolved As closely tracks Mn at both stations. Concomitant increases in As levels result in large sub-surface maxima of 2-5 mg'L" within 10 cm of the sediment-water interface (Figs. 4.18 and 4.19). 1  79  Fig. 4.18. Duplicate summer peeper profiles of dissolved Fe, Mn, As, Ni, Cu anci Zn (AF, respectively) for station 1, Balmer Lake, June, 1993. Replicate samples are represented by double symbols at specific single depths.  80  Fig. 4.19. Duplicate summer peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (AF, respectively) for station 2, Balmer Lake, June, 1993. Replicate samples are represented by double symbols at specific single depths. 81  As profiles also show evidence of depletion at greater depths. Note that at station 1 (Fig. 4.18), the magnitudes of the subsurface maxima differ significantly. Significant enrichments of dissolved Fe appear below those observed for manganese. At both stations, concentrations increase rapidly between 5-10 cm, below which gradual increases result in peeper bottom values up to 4 mg'L"  1  (Figs. 4.18 and 4.19).  Dissolved Zn profiles exhibit variable sub-surface  enrichments in the top 5 cm ranging from 60-150 ug'L" , which immediately fall 1  off to concentrations less than half of the observed water column values. Dissolved C u and N i profiles exhibit marked similarity at the two sampled stations (Figs. 4.18 and 4.19). With two exceptions, the concentrations of both elements decrease sharply immediately below the sediment-water interface and remain generally constant below minima of ~5 ug'L" for C u and -60 ug'L" for 1  1  Ni. The exceptions are two dissolved N i profiles, one at each site, which show evidence of slight subsurface enrichments in the first few centimetres below the interface (Figs. 4.18 and 4.19).  Results from the summer period exhibited  excellent agreement between bottom water values obtained from peepers and those from concurrently drawn pumped samples (compare Figs. 4.18 and 4.19 with Table 4-1). The trace metal porewater data for the fall period are comparable to the summer results (Figs. 4.20-4.22).  The tandem peepers at station 2 yielded  consistent profiles for all constituents (Fig. 4.21).  Dissolved M n and As  porewater distributions are nearly parallel. Stations 1 and 2 are characterized by initially steep M n gradients in the top 5 cm, below which concentrations gradually increase to values ranging from -700-1000 ug'L" in the deepest 1  horizons (Figs 4.20 and 4.21). M n levels in the porewaters of "Tailings Bay" (station 5) steadily increase from -200 ug'L" at the interface to 1800 ug'L" 1  1  between 20 and 30 cm. Below this maximum, concentrations drop to -700 ug'L"  1  82  0  200 400 600 800 1000  Dissolved [Fe], ppb  Fig. 4.20.  0  200  400  600  Dissolved [Mn], ppb  800  0  200  400  600  800 1000  Dissolved [As], ppb  Fall peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 1, Balmer Lake, October, 1993. Replicate samples are represented by double symbols at specific single depths.  83  Fig. 4.21. Duplicate fall peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 2, Balmer Lake, October, 1993. Replicate samples are represented by double symbols at specific single depths.  84  Dissolved [Fe], ppb  Fig. 4.22.  Dissolved [Mn], ppb  Dissolved [As], ppb  Fall peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 5, Balmer Lake, October, 1993. Replicate samples are represented by double symbols at specific single depths.  85  in the deepest samples (Fig. 4.22). At all stations, dissolved As concentrations begin to rise at shallow sediment depths, reaching maxima between ~900 and 3000 ug'L" (Figs. 4.20-4.22). 1  Dissolved Fe increases in concentration with depth at all sites during the fall period; however, the depths and magnitudes of the enrichments vary. Levels at station 1 remain essentially uniform to a depth of 20 cm whereupon concentrations increase gradually to a maximum of ~1000 Ug'L" at the base of the 1  profile (Fig. 4.20). Iron distributions seen at stations 2 and 5 are characterized by variable increases starting at 5-10 cm depth (Fig. 4.21 and 4.22), with concentrations rising to values as high as -3-4 mg'L" in the deepest samples at 1  station 2, but only to 400-500 ug'L" at station 5. 1  Dissolved Zn profiles from the fall at stations 1 and 2 exhibit pronounced, shallow sub-surface enrichments ranging from -130-300 Ug'L" (Figs. 4.20 and 1  4.21) . Below these maxima, concentrations decline dramatically to values less than their respective bottom water levels. In contrast to the other sites, dissolved Zn concentrations at station 5 generally decrease from the uppermost sampled bottom waters to shallow sediment depths (Fig. 4.22). Below a few centimetres in the sediment column at this site, dissolved Zn levels decrease marginally (Fig 4.22) . The bottom water and porewater distributions of dissolved C u and N i from the autumn sampling are similar to those seen in the summer, exhibiting minimal lake-wide variability. At stations 1 and 2 (Figs. 4.20 and 4.21), C u concentrations decline precipitously at shallow sediment depths, approaching minima of <20 ug'L" -5 cm below the interface. Station 5 is characterized by a 1  similar gradient; decreases, however, begin at or slightly above the sediment surface (Fig. 4.22). Steep sub-surface gradients are also evident for dissolved N i distributions at all stations sampled. Nickel profiles at station 1, and to a lesser  86  extent at station 2, show distinct subsurface enrichments, before levels dramatically decline immediately below their respective maxima (Figs. 4.20 and 4.21). As seen for N i , dissolved C u concentrations at station 5 begin a steep decline at or above the sediment surface (Fig 4.22). The winter distributions of the various dissolved constituents demonstrate that a dramatic shift in the chemical environment of the bottom waters and porewaters occurs during periods of ice-cover (Figs. 4.23-4.26). The bottom water enrichments of dissolved Mn that were clearly depicted at all sampled stations from water column profiling are similarly represented in their respective peeper profiles.  Stations 1, 2 and 4 are characterized by increasing dissolved M n  concentrations throughout the lowermost bottom waters to the interfacial porewaters (Figs. 4.23-4.25). Profiles at stations 1 and 2 continue to increase below the sediment surface, reaching maxima between 1200-1400 ug'L" at depths 1  of ~15 and 5 cm, respectively (Figs. 4.23-4.24). At greater depths at these sites, and below the interfacial maximum at station 4, dissolved M n concentrations decrease to ~600-700 ug'L" and remain constant to peeper-bottom (Figs. 4.231  4.25). The bottom waters in "Tailings Bay" (station 6) show no evidence of a nearinterface dissolved Mn enrichment; concentrations remain essentially constant to a sediment depth of ~ 10 cm then fall steeply to levels that approach 20-30 ug'L"  1  (Fig. 4.26). Dissolved Fe distributions in the bottom waters and interstitial fractions observed during the winter period exhibit considerable lake-wide variability. Station 1 is characterized by relatively uniform levels of dissolved Fe to a sediment depth of ~5 cm, below which values increase erratically to ~2100 Ug'L"  1  in the lowest horizons (Fig. 4.23). Concentrations at station 4 remain < 200 ug'L"  1  in the bottom waters and porewaters to a sediment depth of -40 cm, but increase to -600 ug'L" at greater depths (Fig. 4.25). In contrast, the bottom waters 1  87  Winter Station 1  iuu  ^uu  JUU « u u =uu  Winter Station 1  E  ouu  0  1  Dissolved [Ni], ppb  0  2  0  3  0  4  0  5  0  Winter Station 1  F  6  0  Dissolved [Cu], ppb  7  0  Q  5  0  1  0  0  1  5  0  2  0  0  Dissolved [Zn], ppb  Fig. 4.23.. Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 1, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  88  Winter Station 2  Winter Station 2  Winter Station 2  B  E o 8  30  I 1—I -  I  — — i — i r — — i — i r — — i r I—I I |T  2 0  & 10 ».  CD  s S-iof E CO  -20  E o  4= -30 CD O  ca -40 CO  b 0  200  400  Dissolved [Fe], ppb  0  400 800 1200  100 200 300 400 500  200  400  600  800  Dissolved [As], ppb  Winter Station 2  Winter Station 2  0  20  Dissolved [Ni], ppb  Fig. 4.24.  0  Dissolved [Mn], ppb  Winter Station 2  0  -50  40  60  80 100  Dissolved [Cu], ppb  0  50 100 150 200 250 300  Dissolved [Zn], ppb  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 2, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  89  Fig. 4.25.  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 4, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  90  Winter Station 6 E u <B  o  •g  B  Winter Station 6  Winter Station 6  40  20  a> .c i—  CD  T5 c  CD  -20  E  (0  E  8 -40 <D O c  w -60  0  100  200  300  400  Dissolved [Fe], ppb  D  100  100  200  300  400  -60  o  5  10  Dissolved [Mn], ppb  Dissolved [As], ppm  Winter Station 6  Winter Station 6  Winter Station 6  200 300 400 500  Dissolved [Ni], ppb  Fig. 4.26.  0  200  Dissolved [Cu], ppb  200  Dissolved [Zn], ppb  Winter peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) for station 6, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths.  91  15  immediately above the sediment surface at station 2 are enriched in dissolved Fe. Below the interface, concentrations fluctuate significantly throughout the sediment column (Fig. 4.24). The strong similarities between the distributions of dissolved As and Mn observed for the summer and fall periods are not clearly obvious during the winter session. The As profiles at stations 1 and 2 are characterized by zones of generally increasing concentration, extending from the bottom waters, across the sediment-water interface, to depths of -20 and 10 cm, respectively (Figs 4.23 and 4.24). Dissolved As profiles at stations 4 and 6 exhibit uniform bottom water values and increased levels at shallow sediment depths (Figs. 4.25 and 4.26). A zone of relatively high As concentrations between -5 and 15 cm depth at station 4, overlies porewaters that contain variably lower concentrations (Fig. 4.25). Dissolved As at station 6 increases essentially monotonically throughout the sampled sediment column (Fig. 4.26). Dissolved zinc distributions during the winter period are similar to those observed for other seasons; profiles are characterized by variable increases near the interface at each site, below which levels decline sharply to minima in the top two decimetres (Figs. 4.23-4.26). The well-defined bottom water enrichment of dissolved Zn seen at station 1, corroborates that observed in the water-column sample set (Fig. 4.23). Similar maxima at stations 2 and 4, although evident from water column profiling, are not as clearly defined by the peeper data (Figs. 4.24 and 4.25). The winter distributions of dissolved N i demonstrate significant lake-wide consistency and resemble profiles observed during other seasons (Figs. 4.23-4.26). Gradients in bottom waters are not distinct in the peeper profiles except perhaps at station 4, where the concentration decrease accords with the respective water column profile (see Fig. 4.1 ID). At all stations, dissolved N i concentrations  92  diminish abruptly from values of 300-450 Ug'L" in bottom waters to minima of < 1  40 ugL" in the first two decimetres (Figs. 4.23-4.26). Slight sub-surface maxima 1  can be seen at stations 1,2 and 4 (Figs. 4.23-4.25). The gradients of dissolved Cu in winter bottom waters seen in the water column sample set (Figs. 4.8-4.11) are verified by corresponding peeper data (Figs. 4.23-4.26). Profiles at stations 1, 2 and 4 clearly illustrate the zone of decreasing dissolved C u concentrations above the interface (Figs. 4.23-4.25). Below this horizon, values continue to decline at stations 1 and 4 to concentrations < 10 ug'L" (Figs. 4.23 and 4.25). Station 6 is characterized by a 1  more intense sub-surface gradient where levels decrease from over 400 to < 20 ug'L" in the top 20 cm (Fig. 4.26). 1  Sampling conducted in the spring following ice break-up yielded bottom water and porewater distributions of dissolved trace metals comparable to those seen during the well-mixed periods of summer and fall. Logistic and sampling problems restricted the deployment of only two peepers which were emplaced 20 m apart at station 1. Concentrations for all constituents exhibit uniform profiles in the bottom waters and agree well with independently measured water column values (compare Fig. 4.27 with Table 4.4). Reproducibility between the profiles was fair for dissolved Cu and Ni, but limited for Fe, Zn and As. Dissolved M n contents in the spring porewaters at station 1 sharply increase at shallow sediment depths to maxima ranging from ~820-1100 ug'L"  1  (Fig. 4.27). Below these peaks, concentrations steadily decline to values of -500700 Ug'L" at -40 cm depth. The dissolved Fe profiles are characterized by 1  dissimilar sub-surface enrichments. Concentrations obtained from peeper 1 remain constant to a depth of -20 cm, at which point levels increase rapidly to -1 mg'L" in the deepest horizons (Fig. 4.27). Conversely, the profile of peeper 2 1  shows only slight enrichments up to -400 ug'L" between depths of 10 and 20 cm. 1  93  Dissolved [Fe], ppb  Dissolved [Mn], ppb  Dissolved [As], ppb  Fig. 4.27. Duplicate spring peeper profiles of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F, respectively) for station 1, Balmer Lake, May, 1994. Replicate samples are represented by double symbols at specific single depths.  94  The distributions of dissolved As and Zn are both characterized by subsurface enrichments in the shallow porewaters at station 1. Variable As maxima occur between depths of 10-20 cm; concentrations below theses enriched horizons generally decrease to peeper bottom (Fig. 4.27). Profiles of dissolved Zn also exhibit broad maxima of variable magnitude between the sediment surface and depths of -10 cm (Fig. 4.27). Concentrations in one profile steadily decline at greater depths, while in the other (peeper 2), they increase again below a minimum confined between 10 and 20 cm depth. Although bottom water concentrations of dissolved N i and Cu obtained from peeper 1 are significantly lower than those from peeper 2 (Fig. 4.27), their profiles match well at shallow sediment depths. Sharp decreases are evident for both elements in close proximity to the sediment surface, and minima are approached in the first decimetre.  Dissolved N i profiles exhibit subtle sub-  surface maxima at depths of 5 cm.  4.2.1.3 p H  In the spring sampling period, selected cells down the length of one peeper at station 1 were measured for p H by direct insertion of the probe into each chamber. The p H profile is fairly uniform, with values ranging from 7.328.02 (Fig. 4.28). A slight increase in the vicinity of the interface contrasts with a subtle decrease towards deeper horizons.  95  Fig. 4.28. Spring profile of porewater and bottom water pH measured from peeper 2, station 1, Balmer Lake, May, 1994. 96  4.2.2 Core Porewaters  4.2.2.1 Nutrients and Sulphate  Analytical results obtained from centrifuged core samples have been plotted against peeper-derived profiles in order to compare the two methods. Core porewaters were not analyzed for ammonium; consequently, peeper/core comparisons are limited to nitrate and sulphate concentrations. The interstitial nitrate and sulphate fractions from duplicate cores at station 1 demonstrate excellent inter-core coherence and agree well with peeper values (Fig. 4.29). Concentrations determined from the core supernatant, however, disagree with lowermost bottom water measurements acquired from peeper sampling. It appears the composition of the core-top waters are similar to values measured 20-30 cm above the sediment surface. Examination of the data revealed that the disparity is not due to any analytical difficulty, but instead reflects a design limitation of the corer. The upper portion of the sleeve valve of the corer has two large ports which allow upward-flowing water to escape during instrument descent. However, the flow is not free, in that the total cross-sectional area of the ports is less than that of the mouth of the core barrel. The resistance to the flow of water through the instrument ports during lowering causes some water inside the barrel to be carried down with the corer, for probably not more than 1-2 metres.  In regions of fine-scale stratification, there is the potential for the  supernatant water to consist of a mixture of waters from distinct layers. In addition, the top corer ports sit 30-50 cm above the sediment surface, a distance greater than the thickness of the compositionally distinct strata observed in the bottom waters. Since the core-top waters most likely represent an integrated sample of the lower bottom waters, data from these samples have been plotted 20  97  B  Winter Station 1  Fig. 4.29. Distributions of dissolved NO3" and  SO4 " 2  Winter Station 1  obtained from peeper (open circles)  and core porewaters (diamonds) at station 1 (A-B) and station 6 (C-D), Balmer Lake, March, 1994.  Replicate samples are represented by double symbols at specific single  depths. 98  cm above the sediment-water interface. In the following discussion, however, only peeper-derived bottom water values will be used in comparisons with shallow porewaters. The core-derived nitrate profile in the sediments in "Tailings Bay" (station 6) exhibits a steeper sub-surface gradient than the analogous peeper profile; core values approach minima at a sediment depth of 10 cm as opposed to 20 cm (Fig. 4.29). Excluding one presumably contaminated core sample, the corresponding sulphate distributions parallel one another reasonably well.  4.2.2.2 Trace Metals  The duplicate trace metal distributions obtained from core samples at station 1 are generally similar in form but exhibit variable congruence with the respective peeper profiles. Dissolved Fe concentrations at station 1 increase at depth, with core-derived profiles exhibiting steeper gradients (Fig. 4.30). The magnitudes of sub-surface dissolved M n maxima differ markedly in the two cores; core l b is comparable to the maximum in the peeper profile, albeit shallower, while the corresponding peak in core l a is -400 ug'L" less. 1  Fundamental differences distributions.  are evident in comparisons of dissolved As  Core-derived duplicate profiles show dramatic sub-surface  enrichments in excess of 1200 ugL" at 5 cm depth; such increases are not 1  observed in the peeper profile (Fig. 4.30). Distributions for dissolved Ni, Cu and Zn at station 1 exhibit better agreement between their respective core and peeper profiles; decreases in their concentrations occur abruptly at shallow sediment depths. The core-derived sub-surface gradients for all three elements, however, are steeper than those seen in the peeper profiles (Fig. 4.30).  99  Winter Station 1 40  •' • • I  • • •' I' • • • 1 • ' ' l • • • • I • • • •! • 1  —e- -Peeper 0 Core 1  •  Core 2  Winter Station 1  Winter Station 1  B 40  t t i » t t t i i ) i  11  i i i i i i | i t i t  n  -e— Peeper o Core 1 • Core 2  E 0 <D u 20  E o  40  1 T[ | T T I| T T T | T I I | I I I | 1 1 I [ I I I  •g 20 0?  •g S c  *  1 •  c  CD  E  •  c  CD  E 20 <o E  00  E o  e  CD -40 O  8- 40  CO • CO  1000  2000  3000  Dissolved [Fe], ppb  Fig. 4.30.  b  —e--Peeper o Core 1 • Core 2  c ca  c  0  0  CO  i •, • • * t • •»»i  -60 0  500  <»*  1000  Dissolved [Mn], ppb  1500  b  -60 0  400  800  1200  Dissolved [As], ppb  Dissolved distributions of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) obtained from peeper (open circles) and core porewaters (diamonds) at station 1, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. 100  The differences in peeper- and core-derived trace metal profiles at station 6 are similar to those at station 1 (Fig. 4.31). For most elements, gradients in the core porewaters are steeper. For example, decreases in dissolved Mn, Ni, Cu and Zn in core 2 occur over a sediment depth one-half that of decreases of similar magnitude in the peeper profile. Similarly, porewater As increases more rapidly with depth in the core porewaters (Fig. 4.31)  4.3 Sediments  Three high quality cores were obtained using the lightweight gravity corer described by Pedersen et al. (1985). Duplicate cores l a and lb were collected approximately 3 m apart in the more "natural" sediments at station 1, while core 2 was taken from a predominantly tailings region in "Tailings Bay" (station 6). Core logs are presented in Appendix D. Duplicate cores la and lb were visually and compositionally identical. Both cores were veneered by a 2 to 3 mm thick cap of brown noncohesive very fine-grained sediment. The top two centimetres were additionally characterized by alternating light/dark millimetre-scale laminations. These surface horizons were underlain by ~2 cm of gray, homogenous, less-cohesive ooze which graded into a black layer occupying a depth from 4 to 8 cm. Small bubbles, presumably methane, were evident in the latter. Below the 10 cm horizon, a homogeneous gray/chocolate brown ooze comprised the remainder of the core.  Core 2,  collected from station 6 in "Tailings Bay" was very different from the duplicate cores obtained at station 1. Most of the core comprised homogeneous, very fine grained, light-gray, silty tailings material; intermittent thin amber bands (~1 mm thick) occurred in the top 3 cm.  101  Fig. 4.31.  Dissolved distributions of dissolved Fe, Mn, As, Ni, Cu and Zn (A-F,  respectively) obtained from peeper (open circles) and core porewaters (diamonds) at station 6, Balmer Lake, March, 1994. Replicate samples are represented by double symbols at specific single depths. 102  The "natural" sediments at station 1 and the tailings deposits at station 6 are compositionally distinct, particularly below 10 cm depth, as shown by the solid-phase ratio data (Figs. 4.32-4.33).  Below the 10 cm horizon, the  distributions of most parameters remain essentially constant at both sites. The upper decimetre at station 1 has clearly received a significant input of anthropogenic material. XRD analyses indicate that the "natural" Balmer Lake sediments and the tailings material represent two distinct mineralogical assemblages; the major and minor elemental compositions of the three cores are reported in Appendix E. The natural deposits contain variable mixtures of quartz and feldspar (mainly plagioclase).  These contrast with the tailings deposits, which are relatively  feldspar-poor and replete with quartz, abundant clay minerals and halite. The clay fraction is composed primarily of chlorite, with minor contributions from montmorillonite and micaceous minerals (illite and biotite).  Both natural  sediment and tailings signatures are evident in the top horizons. Cores la and b consist of essentially tailings-free, organic-rich sediments below -10 cm depth, as shown by the Corg/ nitrogen, P, N i , C u , Zn and As distributions (Figs. 4.34-4.35). Extraordinarily high metal concentrations in the top several centimetres of both cores, reaching up to -0.5 wt. % N i , 1.8 wt. % Cu, 0.6 wt. % Zn and 0.5 wt. % As (Fig. 4.35), demonstrate that the upper stratum is composed of a mixture of tailings material and natural deposits, which is capped by a 0.5 to 1 cm thick surface veneer of more natural sediments. The opposing C a C 0 3 and organic carbon distributions in the upper 7 cm reflect dilution of natural sediments by relatively carbonate-enriched tailings (Fig. 4.34). The data in Figures 4.34E and 4.35A imply that S and Mn are also enriched in the tailings component at station 1.  103  B 0 I i i i i 1.1 i i i | i i i Clfai \ » i | i i i i  5r E u  £ O  u  io o  1 0  E o  E  2  15  O  u  c Q. CO Q  CD  O O  15  - Core 1a  •o- - Core 1b  30  _c  £20  CD Q  a  35  imliiillnn  3  llllllllllllllllllllll 4 5  CD  25 t •O—Core 1a 30  -Core 2  25  •e -Core 1b 30  —<•—Core 2  35 0.02  Si/AI Wt. Ratio  15  e  £ 20  25  I0r  • • • . i •. • • i .  0.04  35  0.06  Ti/AI Wt. Ratio  r HJJJIIIIIU  0.1  0.3  0.5 0.7  Mg/AI Wt. Ratio  D • 11111 i m i ' 1111 .'.ici.y '  m^ft-t | t t . 1  111111111  1  1  < O."0  y " .°-. (3  E o  1 0  -  2O 15 o  10  6b  •  :  15  *  I 20  <•  CD  o  25  r-  \  * ^  a' c  -  —O -Core 1a  30 •  •O— Core 1a  Core 1b  30  [ — • — Core 2 35  • •  1  • • • •  0.05  •Q—Core 1b —»—Core 2  *  . i —  0.1  0.15  1  25  i . . . .  0.2  K/AI Wt. Ratio  0 25  0.02 0.06  0.1 0.14  Na/AI Wt. Ratio  0.18  35  1  0  * *  1  • * • *  0.01  1  • • • • ' • •  0.02  P/AI Wt. Ratio  Fig. 4.32. Sedimentary weight ratio profiles of Si/AI, Ti/AI, Mg/AI, K / A l , Na/AI and P/Al (A-F, respectively) for station 1 (duplicate cores l a and l b , open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994. 104  1  •  '  B i n |ii  U !C»  I»I  I•" I'  -1 1 1 1  1  1 1 1 1  -  X '"'•'•'° *.... £  E o  9 "**.*H  1 0  2  E u 2 O o c  9% '  o 15 u  c ct  f  CD Q  15  •  >  V  20  Q  25  25  —o- - Con 1a 30  •v u  1 0  -a-  •fa •  —El--Core 1b  30  -Con 2 2  :  di  :  —•—Core 2  11111111111111111  0  6o  —O—Cora 1a  - Con 1b  35  6a  8  . . . . i . . . 1 1 .1. i  35  12  0  Zr/AIWt. Ratio (*10 ) 4  10  •  20  Sr/AIWt. Ratio (*10 ) 4  c rO-^EJ:..;...... •|3"--.0 • rjjr'' *  f. '• oh '.  K  9 W  /  9  —O— Core 1a  —a—Core 1b  d  :  ! —•—Core 2 .—i— •— • —  0  2  4  6  8  Rb/AI Wt. Ratio (*10 ) 4  Fig. 4.33. Sedimentary weight ratio profiles of Zr/Al, Sr/Al and Rb/AI (A-C, respectively) for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994. 105  Total Sulphur, wt.%  Fig. 4.34.  Fe/AI Wt. Ratio  Sedimentary concentrations of organic carbon, calcium carbonate, total  nitrogen, values for organic carbon/nitrogen, total sulphur and iron/aluminum weight ratios (A-F, respectively), for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994. 106  0  I i • i i  I  i i i  •I  OB1.L  5  9" E o E o o  10  f  r  20  25  10  o  6  15  CD Q  i T ;  I {  :  h i i  . . . < j — C o r e 1a  j 30  35  : t  •••O  o I I  « I  I  -Core  — o — Core l a  1b  1  I I  I  I I  1 0 0 0  1 1 I  I  - C o r e 1b  — Q  3 0 ft  —«—Core 2  —•—Core  • ' • I III 2000  Manganese, | i g g '  D  1  K  1a  •a -- C o r e  1b  — » - -Core 2  2  0  * • * •  10002000  1  1  • • • *  3000  * * • '  1  4000  35  5000  I I I I I I 1 I I I I t I I I I  2000  ,-1 Nickel, ngg"  4000  Zinc, ugg"  E  Copper, jig-g"  30  —o- - C o r e  » 1  0  Fig. 4.35.  r  6000 1  F  Arsenic, jigg-  1  Lead, jig-g"  1  Sedimentary concentrations of manganese, nickel, zinc, copper, arsenic, and  lead (A-F, respectively) for station 1 (duplicate cores la and lb, open symbols) and station 6 (core 2, closed symbols), Balmer Lake, March, 1994. 107  In contrast to the "natural" sediments site, "Tailings Bay" sediments (core 2) are characterized by higher carbonate contents, low overall organic C, total S, and total N and P contents, and higher F e / A l weight ratios (Fig. 4.34). Paradoxically, the concentrations of total sulphur and trace elements (e.g., N i , Cu, Zn and As) are significantly depressed at shallow sediment depths relative to maximum values observed at station 1 (Fig. 4.34 and 4.35). Profiles for Mn, N i , Zn, Cu, As, total S, and F e / A l are essentially invariant throughout the sampled sediment column, with the exception of slight surficial enrichments in C u and As. The compositional contrasts between the two lake sites can be further assessed by means of element/Al ratios.  Variations in such ratios can be  interpreted in terms of the texture and mineralogy of sediments (Calvert, 1976). The ratio profiles in Figs. 4.32 and 4.33 show that there is some compositional similarity in the top horizons, but at depths greater than a few centimetres, there are major differences in mineralogy between the two locations.  108  V. DISCUSSION  5.1 Water Column  Balmer Lake is a highly dynamic system which exhibits pronounced seasonal variability, both physically and chemically. The non-steady state nature largely results from the geographic position of this inland, boreal waterbody, and from variable loadings and dynamics of mining-related inputs. The annual freeze-thaw cycle of the terrestrial and aquatic regimes fosters significant seasonal signatures. For example, the formation of ice in early winter essentially cuts off atmospheric interactions for several months; subsequent reactions in such shallow "closed systems" can greatly affect water column chemistry. The freeze-up period is also characterized by reduced circulation, allowing the development of fine-scale stratification. In addition, the associated seasonal temperature swings of up to 20°C undoubtedly influence the kinetics of abiotic arid microbially mediated reactions in the water column and sediments. Pronounced seasonal pulses of fresh water and associated allochthonous inputs to Balmer Lake are evident in the spring. Spring-melt at these latitudes is rapid, and the combined inputs of terrestrial run-off and the melting of 1 m of surface ice, may lead to a 30 % increase in liquid-water volume. Furthermore, vigorous stream activity can result in increased particulate loads in spring. The nature of the tailings circuit design, in combination with the annual freeze/thaw cycle in Balmer Lake, results in a pronounced seasonal signal with respect to heavy metals and cyanide. Dramatic increases in the concentrations of these parameters are consistently observed at the Balmer Lake outlet (Fig. 2.1) in the spring, and have been shown to be strongly correlated with ice deterioration  109  (EIG, 1986). Moreover, the total inventory of heavy metals also increases at this time. Four independent controls are believed to account for the observed annual cyclicity. First, the absence of wind turbulence under the ice allows the denser, metal-rich effluents to collect in the bottom waters over the winter months. Prior to whole-lake turnover in the spring, the outfall drains the less contaminated surface waters. However, upon ice break-up and subsequent wind-mixing, the accumulated inventory of mine-effluents is dispersed throughout the water column; this can contribute to higher metal levels at the outfall. The total increased metal burden in Balmer Lake at ice-out can also be attributed to freeze/thaw processes.  As will be mentioned in section 5.1.4 below, the  resuspension of a non-cohesive flocculant layer (built up over the winter months) has been suggested to account for lake-wide increases in springtime metal levels (EAG, 1986). Greater dissolved loads at this time may also stem from the dissolution of cyanide compounds.  Metal-cyanide complexes, such as  ferricyanide [Fe(CN)6]" , can form insoluble double-metal cyanide precipitates 4  such as copper ferrocyanide [Cu Fe(CN) ] (Higgs, 1979). Such complexes are 2  6  stable in the absence of U V radiation and may potentially accumulate on the lake bottom over the winter months. Upon ice break-up, however, their dissolution may foster significant releases of CN" and associated metals into the water column. Finally, the processes described above also occur in the tailings ponds; however, similar increases in heavy metal concentrations and cyanide precede those in Balmer Lake due to the earlier thaw. Thus, seasonally elevated inputs from the tailings circuits also contribute episodically to the lake burden.  110  5.1.1 Physical Lirnnology  The shallow water column of Balmer Lake is fairly well-mixed for the approximately 7-month ice-free period. During this time, the solubility of dissolved oxygen is temperature dependent. The generally windy conditions during the summer field session obviate detectable thermal stratification (Fig. 4.1). The high biological oxygen demand of the organic-rich sediments in the lake, (6-10 % organic carbon) in combination with restricted atmospheric exchange, fosters a gradual depletion of dissolved oxygen in the deeper lake areas over the duration of ice coverage (~ 5 months) (Fig. 4.3). Spring sampling, conducted less than 2 weeks after ice-off, revealed no traces of the previous stratification (Fig. 4.4).  The rapid breakdown of chemical gradients and the  pronounced warming of the water column over a short time period in the spring, suggest that wind and solar influences play a significant seasonal role.  5.1.2 Sulphate  The oxidation of sulphide-bearing ore in the milling circuit liberates large quantities of SO4 " which are delivered with the tailings to the clearwater ponds. 2  Dissolved SO4 " loads entering Balmer Lake average over 20 mmol/L. This has 2  resulted in a lake-wide yearly average dissolved sulphate concentration of > 3 mmol/L, greatly exceeding natural concentrations in non-perturbed Canadian Shield lakes (Nriagu et al, 1982). Due to the scale of this input, month-to-month variations in the sulphate loadings from the two mines should ultimately determine any seasonal variability.  Ill  2  While SO4 " distributions remain essentially constant through the spring, summer and fall, winter profiling revealed large bottom water enrichments in the deeper lake areas (Fig 4.7).  Peeper profiles of bottom water and porewater  sulphate, however, indicate that the sediments were not releasing sulphate during this period; interfacial gradients in fact suggest pronounced sedimentary SO4 " consumption (Figs. 4.15-4.16; porewater sulphate profiles will be discussed in greater detail in section 5.3.2). The bottom water sulphate increases (which are seen only at depths > 2.5 m) must therefore represent a lateral advective flow of dense, sulphate-rich water moving out across the deeper lake basin. Indeed, profiling conducted to 2.6 m at the shallowest water column site (station 3), and at the shallow peeper site (station 6) did not reveal any evidence of such enrichments (Fig. 4.7C and 4.16F, respectively); only a marginal signal was observed at the 3 m depth at station 4 (4.7D). Although the exact source and timing of such events cannot be determined accurately, the respective inputs from the Dickenson and Campbell tailings ponds represent two obvious potential sources of the high density discharge. During the winter sampling period in March, 1994, the Campbell mine was not discharging to Balmer Lake; complete cessation of discharge was in effect from Dec. 10/'93 to May 2/'94 (PDI, pers. comm.). However, waste inputs from Dickenson persisted throughout the winter months.  The effluents  potentially contributing to the observed chemical stratification averaged approximately 25 mmol/L SO4 ", 1.5 mmol/L N r V , 10.2 mg'L" C N , 2.5 mg'L" 1  2  1  Cu, 1.5 mg'L" Ni, 1.0 mg'L" Zn, and 2.0 mg'L" As (Goldcorp, pers. comm.). The 1  1  1  high dissolved salts content of the tailings effluents, reflecting largely the high sulphate concentrations, would impart considerably higher densities than associated lake waters. A n episode of slumping and oxidation of sulphide-rich  112  tailings material from somewhere within the lake represents a less likely, indeed implausible, explanation for the bottom water SO4 " increases. 2  During the fall sampling period, a bottom water sulphate enrichment was also evident at the site most proximal to the Campbell Mine effluent discharge (Fig. 4.13F). In light of the typical windy conditions, it is reasonable to assume that during ice-free periods such profiles would only be manifested within a short distance from the input point. The winter water column distributions of dissolved Fe, As and Cu suggest that sulphide precipitation occurs in the seasonally reducing bottom waters of Balmer Lake (Figs. 4.8-4.11). Due to the large excess of S0 ", however, direct 2  4  evidence for sulphate reduction in the water column profiles is probably obscured. Bottom water SO4 "distributions seen in peeper profiles at stations 1 and 2, and to a lesser extent at station 4, suggest a mechanism for sulphate removal near the sediment-water interface at these sites (Figs. 4.15 and 4.16). Considering the trace metal profiles, the abundance of reducible sulphate and the reducing conditions prevalent in the winter, the seasonal precipitation of authigenic mineral sulphides in the water column is a tenable hypothesis that will be further assessed in the trace metal sections below.  5.1.3 Ammonium and Nitrate  The burdens of ammonia and nitrate in the water column of Balmer Lake greatly exceed concentrations observed in non-contaminated fresh water bodies. The observed enrichments can be partially attributed to mining-related byproducts delivered to the lake via drainages from lake-side clearwater polishing ponds. More specifically, excess levels of N H * and N 0 4  from the breakdown of cyanide compounds.  113  3  appear to be derived  The addition of sodium cyanide in the tailings circuit facilitates the removal of gold, and other metals, in carbon-in-pulp milling operations. The long-term use of such concentration techniques has resulted in significant additions of cyanide to the waters of Balmer Lake, with levels in the past sometimes exceeding 5 m g / L . Historically, mitigation of cyanide inputs has entailed degradation that occurs naturally in tailings ponds. In such systems, the volatilization of H C N to the atmosphere has been shown to be primarily responsible for cyanide removal (IEC, 1979). More sophisticated treatments emerged with stricter environmental standards in the late 1970's, at which time hydrogen peroxide (Vickell et al, 1989) and Inco-S02 (Devuyst et al, 1982) methods became more widespread. The large ammonium burdens in Balmer Lake are partly attributable to the hydrolysis of cyanate (CNO") and cyanide (CN") compounds. Cyanate is generated in the milling processes due to reaction of free cyanide and cupric ions. It may additionally stem from the treatment of cyanide-bearing wastes upon oxidation with hydrogen peroxide or hypochlorite (Higgs, 1993). Cyanate can hydrolyze in basic mill effluents to produce ammonia via:  CNO" + OH" + H 2 O -> C O 3 - + N H 3  (1)  2  At pH's < 9.3 (i.e., Balmer Lake), however, N H 4 is the thermodynamically +  favoured product of cyanate hydrolysis (Stumm and Morgan, 1981). Cyanide will also hydrolyze in neutral waters to form ammonium via:  CN" + 3H20 -> HCOO- + N H 4  <>  +  2  114  The large bottom water ammonium enrichments in the winter water column in Balmer Lake are believed to be derived from both sediment sources and from mining-related lateral advective flows. Ammonium is generated by heterotrophic bacteria as a primary end-product of organic matter decomposition, either directly from proteins or from other nitrogenous organic compounds (Van der Weijden, 1992). The remineralization of organic matter in Balmer Lake sediments, and the concomitant release of N H  + 4  to the porewaters,  results in steep gradients at shallow sediment depths. Ammonium released during aerobic respiration into oxic horizons will be oxidized by nitrifying bacteria (nitrification) to nitrate (Sholkovitz, 1973). Similarly, the same fate holds true for upward diffusing N H  4  that migrates into oxic sediments. However, the  complete consumption of oxygen, and the reductive dissolution of M n oxyhydroxide phases in the reducing winter bottom waters, effectively remove these potential oxidants from interfacial horizons. remineralized N H  + 4  In this environment,  is allowed to diffuse from permanently reducing horizons  and accumulate in the anoxic bottom waters. Such effluxes have been described elsewhere: Stadelmann (1971) and Wetzel (1975), for example, report the presence of >500 umol'L" N H / in temperate, productive lakes in Michigan and 1  Switzerland, respectively . This topic will be discussed in greater detail in section 5.3.1. The large N H  4  concentrations reported in the effluents discharged to  Balmer Lake indicate that the winter bottom water enrichments can also be partially attributed to a lateral transport mechanism, as suggested for sulphate. Support for this argument can be seen in the fall bottom water and porewater NH  + 4  distribution (Fig. 4.13D) in "Tailings Bay" (station 5). The ammonium  profile is characterized by a well-defined bottom water enrichment, comparable to that observed for SO4 ". Like sulphate at this site, ammonium concentrations 2  115  decrease dramatically at shallow sediment depths demonstrating that the sediments are not releasing N H  + 4  , but are serving as a diffusive sink.  Ice formation in Balmer Lake may also indirectly contribute to the large winter N H  + 4  inventories. Free cyanides exist in solution as cyanide (CN") and  hydrocyanic acid (HCN) in the equilibrium (Stumm and Morgan, 1981):  HCN  <->  H+  +  CN" (pKa = 9.3)  (3)  At the neutral p H of Balmer Lake, 95% of free cyanide exists as molecular H C N . As a result of its relatively high vapour pressure, H C N tends to volatilize to the atmosphere, even in stagnant solutions (Higgs, 1993); the presence of an ice cap during the winter months, however, prevents H C N release. Thus, the increased residence time of cyanide compounds during the winter months, and their +  subsequent breakdown, may contribute to the increased winter N H  4  concentrations in Balmer. Lake. During the well-mixed periods of the spring, summer and fall, nitrate concentrations remain fairly constant (Tables 4.1, 4.2 and 4.4). Pronounced decreases beginning 10-20 cm below the top of the oxycline at all sites in the winter, however, suggest that denitrification is occurring in the bottom waters during this period below depths of 2.4-2.7 m (Fig. 4.5). Denitrification becomes important at low O2 concentrations (<6 umol/L) and involves the bacteriallymediated reduction of one or both ionic nitrogen oxides (nitrate or nitrite) to their gaseous oxides and eventually to N (Froelich et al., 1979). Peeper bottom2  water profiles at stations 1,2 and 4 suggest that nitrate consumption (i.e., defined where d [N03"]/dz > 0) begins in winter at depths approximately 10,15 and 5 2  2  cm above the sediment-water interface, respectively (Figs. 4.15B, E and 4.16B); peeper data were not obtained from station 3. Similar profiles demonstrate that  116  denitrification does not occur in the shallow waters (~1 m) at station 6 in winter (Fig. 4.16E). Nitrate is produced in most lake systems primarily by the fixation of molecular N 2 by cyanobacteria and via the bacterial-mediated oxidation of ammonium diffusing upwards into oxygenated horizons in the sediments or water column (Wetzel, 1975). The abundance of oxidizable N H nitrification is responsible for a high proportion of the N 0  3  + 4  suggests that  burden in Balmer  Lake. In general, the ratio of N 0 - N to N H - N is a highly variable function of +  3  4  lake trophic status, hydrogeology, and natural and pollution-derived sources (Wetzel, 1975). Well-buffered, oligotrophic lakes receiving drainages from calcareous landforms, for example, can have N 0 -N: N H - N ratios of greater +  3  4  than 25:1. Conversely, in permanently anoxic, eutrophic lakes, ratios of 1:10 are common  (Wetzel, 1975).  The values of N 0 - N : N H - N in Balmer Lake +  3  4  progressively increase over the winter, spring, summer and fall, with ratios averaging <1, 1.8, 3.5 and 6.2, respectively. decrease in N H  + 4  This hierarchy largely reflects a  concentrations from the maximum observed during the winter  period. The relative removal of N H  + 4  can be attributed to biological nitrification,  phytoplankton uptake and sorption to clay minerals.  5.1.4 Suspended Particulates  The considerable intra-station variability of particulate trace metal concentrations in the well-mixed water column of Balmer Lake warrants consideration of potential sampling and analytical errors (Tables 4.6 and 4.7). Inaccuracies in the data set may have arisen from two sources. First, the overpressure filtration units, on occasion, leaked. In such circumstances, the total volume of water filtered could not be accurately determined, and as a result,  117  errors would be manifested where the total mass of particulate metal per unit volume is computed. Such error should not be introduced, however, into the determinations of total mass of metal per gram of suspended material. Yet, the coarsely paralleled variability of mass: mass calculations with those of mass: volume, suggest that some degree of particulate heterogeneity is evident in the water column. Additional error may have been introduced into the mass:mass values since the absolute particulate fraction, determined by subtracting the prefiltered from post-filtered weights, was very small. A third source of error, with respect to Zn concentrations, appears to have originated from contamination either during cleaning of the polycarbonate filters or during the filtration procedure. Values determined from filtration blanks ranged from 0.3-4.8 ppb (Appendix B). Such levels exceed some of the low-end water column particulate values.  In order to obtain the most representative  estimates, averages of the respective filter blanks for each sampling session were subtracted from the total. It should be noted that filters taken out of their wrappers and processed (i.e., not acid-cleaned) exhibited negligible levels of all metals (Appendix B). Zn enrichments most likely arose from water contact with zinc-rich rubber fittings in the filter apparatus, despite the O-rings and stoppers being wrapped in teflon tape prior to use. Concentrations of total suspended solids in the water column of Balmer Lake are high, and are fairly typical of levels encountered in productive, nearshore lake zones characterized by higher rates of sediment resuspension (Nriagu et al, 1981). The particulate trace metal levels greatly exceed values reported for other non-perturbed inland waters (Table 4.6). In general, values in Table 4.6 are high in comparison to suspended metal concentrations observed in lacustrine and estuarine waters receiving pollutant metal inputs (Kubota et al, 1974; Jackson, 1978; Sholkovitz, 1979; Nriagu et al, 1982; Jackson and Bistricki,  118  1995). Moreover, the lake-wide averages for the observed concentrations in suspended particulates of approximately 5 wt.% Fe, 850 mg'kg" Mn, 1500 mg'kg" 1  1  Ni, 7000 mg'kg" Cu, 5000 mg'kg" Zn and 4000 mg'kg" As also surpass reported 1  1  1  values of other polluted systems (see Nriagu et al, 1982). The relatively large magnitudes of the particulate values demonstrate that the settling of suspended material may present an important transport mechanism of metals to bottom sediments. Aquatic microorganisms exert controls on the geochemical cycles of many trace elements via active and indiscriminant associations (Morel and Hudson, 1985). The interactions of organic particles with metal ions and other reactive elements can occur via two mechanisms: metal ions can become coordinated to particulate surfaces of cells and/or fecal material in accordance with surface coordination or sorption equilibria, or alternatively, they can be incorporated into living cells and physiologically assimilated (Sigg, 1985). The high affinity of metal ions for biological surfaces suggests that biogenic particles may play a prominent role in the binding and transfer of metals to bottom sediments (Nriagu et al, 1981; Sigg, 1985; Morfett et al, 1988; Reynolds and HamiltonTaylor, 1992). Particulate organic matter is abundant in the water column of Balmer Lake and constitutes a large percentage of the total suspended solids (Tables 4.5 and 4.7); concentrations are similar to levels measured in other Ontario lakes (e.g., Nriagu et al, 1981, Nriagu et al, 1982). Several studies have indeed shown that concentrations of particulate trace metals are strongly influenced by the production of POC (Sigg, 1985; Nriagu et al, 1981 and 1982). They attributed more efficient scavenging of trace elements to larger productivities and higher particle sedimentation rates. Except for Zn, however, particulate metal levels in Balmer Lake do not exhibit a clear correlative relationship with concentrations of POM.  119  Other than associations with organic matter, variable fractions of the particulate trace metal burden are likely to be associated with metal oxyhydroxide precipitates (e.g., Fe, M n and Al) and clay minerals in the lake waters.  The relative distribution of trace metals between these fractions is  unknown.  However, some inferences can be made based on seasonal  comparisons of POM, TSS, metal per volume and metal per gram particulate. For example, the highest particulate metal concentrations evident in the spring water column coincided with the highest levels of total suspended solids (Tables 4.5 and 4.6). The relatively lower organic content of these particulates, however, implies that large fractions of the trace metals are associated with inorganic matter. Furthermore, the winter water column was characterized by less than half the average P O M content of the summer, but contained comparable or greater concentrations of all trace elements (Tables 4.5 and 4.6). From the above discussion, it is evident that P O M does not exert the principal control on particulate metal levels in the water column of Balmer Lake. However, given the abundance of particulate iron in lake waters, control by Feoxide phases may be important.  Indeed, existing reports on trace metal  scavenging by Fe and M n oxyhydroxides in lakes polluted with mine and smelter wastes illustrate the importance of such metal-particulate associations (Jackson and Bistricki, 1995). Seasonal concentrations of particulate Fe species in Balmer Lake range from 107-649 |ig'L-l which correspond to 3 to 6 wt.% of suspended material (Tables 4.6 and 4.7). The average seasonal concentrations of particulate Ni, Cu and As track mean suspended iron concentrations; for all four elements (including Fe) the seasonal hierarchy follows the order: spring > winter > summer and fall (Table 4.6). Variations in iron-trace metal relationships may be influenced by the relative proportion of less reactive Fe species, such as those in the lattices of resuspended clay minerals. Inorganic particulate fractions may  120  also be present as CaC03, quartz, chlorite, illite and feldspars; these minerals are also capable of sorbing both organic substances and metal ions, but they have relatively small specific surface areas and thus relatively little capacity for metal ion binding (Sigg, 1985). A basic mass balance analysis conducted during a period of elevated metal concentrations in Balmer Lake in the spring of 1986, suggests that an increase in the discharges from the adjacent tailings ponds could not alone account for the total increased metal levels (Masala, 1995). The latter point, in conjunction with observations of greater particulate metal fractions during this period, leads to the conclusion that wind-suspension of fine-particulate material were contributing to such loadings. More specifically, it was proposed that in the quiet waters under the ice in winter, finer particulates settle out gradually and form a non-cohesive flocculant layer susceptible to resusupension upon ice break-up (PDI, 1986).  This argument is consistent with the high TSS and  particulate metal levels observed in Balmer Lake observed during the study period.  5.1.5 Trace Metals  The dynamic nature of Balmer Lake is marked by considerable variability in the seasonal distributions of the measured trace elements.  Two principal  mechanisms are invoked to account for the observed variability: seasonal shifts in the redox geochemistry of the sediments and bottom waters, and the effect of mining-related discharges. It will be demonstrated below that an understanding of the timing and relative influence of such mechanisms is critical in the evaluation of the spatial variability observed in the winter water column.  121  To constrain fully the mechanisms controlling the behaviour of trace metals in the water column of Balmer Lake, both interfacial and porewater distributions must be examined.  A detailed assessment of the seasonal  porewater chemistry is presented in section 5.3, and an interpretation of trace metals in the water column is offered immediately below. The latter is divided into three sections: 1) descriptions of the respective trace element geochemistries; 2) presentation of a water column model; and 3) application of the model to four zones in Balmer Lake.  5.1.5.1 Iron  The transport of Fe and Mn in water bodies and sediments has received a great deal of attention because of the central role that these abundant metals play in the geochemical cycling of other elements. The surface properties of Fe oxyhydroxides and associated trace metal sorption mechanisms, in particular, have been well studied (Davis and Leckie, 1978a; Davis and Leckie, 1978b; Vuceta and Morgan, 1978; Benjamin and Leckie, 1982; Tessier et al, 1996). The redox geochemistry of Fe has been repeatedly stressed as a principal mechanism governing the distribution of many trace elements in lacustrine systems (Laxen, 1985; Tessier et al, 1985; Belzile and Tessier, 1990; Davison, 1993; HamiltonTaylor and Davison, 1994; Jackson and Bistricki, 1995). Iron (II) and iron (III), represent the two oxidation states of importance to the aquatic geochemistry of Fe. Ferrous iron is stable in anoxic water, exists predominantly as a simple hydrated aquo ion, and is soluble with respect to most inorganic ions; however, in some environments, the solubility products of siderite (FeC.03), amorphous iron sulphide (FeS) and vivianite [Fe3(PC>4) 2] may be exceeded (Davison, 1993). In well oxygenated waters, Fe(III) is the stable  122  oxidation state. This acidic cation occurs as completely hydrolyzed oxides at neutral p H , with its solubility being controlled by amorphous iron hydroxide (Sturnm and Morgan, 1981). Numerous field and experimental works have been carried out to characterize iron and manganese oxide phases in lacustrine waters and sediments (Davis and Leckie, 1978b; Laxen and Chandler, 1983; Davison et al., 1992). For the most part, the nature of iron oxide particles in natural waters is not completely understood. Although variations in particle size, composition and structure have been observed, general similarities exist. In oxygenated lake waters, ferric iron has been predominantly found as amorphous particles and colloids, which are negatively charged due to adsorption of humic substances (Tipping, 1981). The spherical or ellipsoidal particles are presumed to be poorly crystalline ferrihydrite with mean diameters typically in the range of 0.05-0.5 urn. The general presence of these particles as amorphous phases makes them readily susceptible to dissolution under anoxic conditions. Tipping et al. (1981) used electron microscopy to examine iron-rich particles in Esthwaite Water surface sediments. The particles contained 30-40 % Fe (by weight), with P, N , Mn, Si, S, Ca and Mg accounting for up to 8 wt. %. Organic matter constituted up to 36 wt. % with one third of this being represented by humic substances. The data in Table 4.6 clearly illustrate that particulate Fe fractions dominate the total iron inventory in the waters of Balmer Lake. The ratio of particulate to filterable species exhibits considerable seasonal consistency with particulates comprising 60-89 % of the total Fe inventory. The Fe-particulates are believed to comprise variable mixtures of iron oxides and clay mineral fractions derived from resuspended sources. It should be noted that dissolved Fe species become more important during the winter months.  123  5.1.5.2 Manganese  As for iron, the biogeochemical cycling of Mn has been shown to govern the behaviour and mobility of redox sensitive elements in freshwater environments (Davison and Woof, 1984; Agett and O'Brien, 1985; Sigg et al, 1991; Young and Harvey, 1992; Davison, 1993). Of the three naturally occurring oxidation states, Mn(II) is stable in anoxic waters, where the predominant species are simple hydrated aquo ions. Manganese (II) is very soluble with respect to most inorganic ions, although in certain environments, such as interstitial waters, it is possible to exceed the solubility product of rhodochrosite (MnCOa) (Carignan and Nriagu, 1985). Conversely, in well-oxygenated waters, Mn02 (Mn IV) is the thermodynamically stable form. In practice, a range of metastable oxidation products exist in natural waters due to kinetic limitations (Stumm and Morgan, 1981); Mn(III) is only stable in a lattice of mixed oxidation state (Davison et al, 1993). Numerous works have attempted to describe manganese oxidation products and characterize them using electron microscopy and X-ray diffraction (De Vitre and Davison, 1993). However, the basic chemistry of both the Mn(IV) and Mn(III)-bearing phases are not well understood.  Natural manganese  particles are difficult to identify because of their extreme microcrystallinity which renders them X-ray amorphous. In addition, due to their tendency to coexist and coat many other minerals and organic material, sometimes only electron microscopy can provide identification (Chiswell and Mokhtar, 1986). The most common forms of M n oxides identifiable in oxic lake sediments are y-MnOOH and birnessite (Mn02). The structure of the latter can be complex. Birnessite is thought to contain Mn(III) and possibly Mn(II) or other divalent metal cations (e.g., Pb, Cu, Zn) in addition to predominant tetravalent Mn species. The humic  124  content of these phases is low (~1 %), but bacteria and/ or algae may contribute 20-30 % by weight (Davison,1993). In well-oxygenated, neutral waters (pH 6-8), the half-life of Mn(II) is typically on the order of days to months; by contrast, Fe(JJ) has a typical half-life of ~4 h under the same conditions (Stumm and Morgan, 1981). The variable reaction rates of M n have been attributed to bacterial-mediated oxidation kinetics; various strands of evidence have been introduced to support the importance of microbial interactions (Davison,1993).  The differing oxidation  kinetics have corresponding consequences with respect to the distribution of soluble and particulate species. Examination of the partitioning of Mn in the water column of Balmer Lake indicates that particulate phases are insignificant relative to the dissolved fraction (Tables 4.6 and 4.7). During all sampling periods, dissolved manganese counted for at least 95 % of the total inventory. These results agree well with a recent study of M n speciation in a Precambrian Shield lake, where predominantly non-particulate forms were also observed (LaZerte and Burling, 1990). As for Fe, the term "dissolved" must be used with caution, as colloidal phases have been shown to contribute significantly to filterable fractions (De Vitre et al, 1988; LaZerte and Burling, 1990). In natural systems, the proportion of M n in particulate form may vary considerably. For example, Laxen et al (1984) showed that in various British rivers and streams the particulate fraction varied from 0 to 100%. The wide range has been attributed to the relative contributions of M n from two distinct sources: 1) weathering of Mn-bearing minerals which yields suspended particles; and 2) inputs of soluble manganous species from oxygen-deficient sediments and ground waters (Davison, 1993). The relative influences of these sources, and therefore the particulate proportion in a particular lake or stream, depends on the  125  local hydrogeology.  It will be shown below that the dominance of filterable  species in the Balmer Lake water column can be attributed to the direct input of dissolved manganese.  5.1.5.3 Arsenic  The inherent toxicity of many As compounds and their transformations in biogeochemical cycles has stimulated considerable interest in As speciation in the environment (Ferguson and Gavis, 1972; Sanders, 1983; Reimer and Thompson, 1988; Cullen and Reimer, 1989; Anderson and Bruland, 1991; Bowell, 1994). Although total dissolved As was measured in Balmer Lake, predictions of speciation can be made using relevant thermodynamic constraints and results of previous studies. Dissolved As can occur in natural waters in both inorganic and organic forms. In oxygenated waters at neutral pH, the anionic arsenate species (H2ASO4- at 3.6<pH<7.3) is the thermodynamically stable form of As. Conversely, neutrally charged arsenite (As(OH) °) is the predicted form in 3  reducing environments (Cullen and Reimer, 1989). As(V)/As(III) ratio of 1 0 - 1 0 15  26  The large predicted  suggests that As(III) should contribute a  negligible fraction of total inventories in neutral waters.  However, the  observation of significant amounts As(III) in oxygenated waters, and the presence of As(V) in reducing regimes, suggests that the two are rarely in thermodynamic equilibrium. Arsenic is also host to a suite of organic compounds, of which monomethylarsonic acid (MMAA) and dimethylarsinic acid (DMAA) contribute significantly (-10%) to total As concentrations in freshwaters (Andreae, 1979; Anderson and Bruland, 1991). Because arsenate and phosphate are chemical analogues there is little discrimination between the two species during biological  126  uptake.  This has led to the suggestion that organisms have developed  mechanisms to isolate and detoxify As by producing organoarsenicals (Wood, 1974). However, the processes that promote biological methylation of As are still poorly understood. Factors including high A s i P O ^ ratios in ambient waters, -  and the nature of biotal communities, have been invoked to explain greater abundances of methylated compounds (Anderson and Bruland, 1991). Arsenic concentrations in the water column of Balmer Lake greatly exceed natural baseline values observed in non-perturbed Canadian Shield lakes. The abundance of arsenopyrite in the milled ore suggests that the oxidation of this phase during the milling process is responsible for the large inventory. Arsenic is largely dominated by filterable fractions throughout the sampled seasons, contributing over 95% of the total burden (Table 4.6).  Large bottom-water  particulate spikes observed during the winter season (27-32% of the total inventory) represent the only deviations from the typically low suspended particulate As values. Considering that Balmer Lake is productive and hosts excess As with respect to phosphate, there is a high probability that concentrations of methylated species are substantial. In all likelihood, however, these species probably represent an insignificant fraction of the total arsenic in the lake waters. Arsenic species in the water column of Balmer Lake are most likely dominated by arsenate and arsenite, with their respective contributions fluctuating in accordance with seasonal oxic/anoxic cycles.  127  5.1.5.4 Nickel  Nickel exhibits a nutrient-type distribution in the oceans and is characterized by both shallow and a deep water regeneration. Vertical profiles in the top 800 m appear similar to phosphate, while below this depth nickel exhibits a deep maximum similar to silicate (Bruland, 1983). Concentrations of dissolved nickel increase from ~2 nmol'L" in surface central gyre waters of the North 1  Pacific to a deep maximum of 11 nmolL-1 (Sclater et al, 1976; Bruland, 1980). Similarly shaped profiles determined for the North Atlantic exhibit maxima one half the magnitude of those observed in the deep North Pacific (Bruland and Franks, 1983). The non-steady-state nature of lacustrine systems combined with the relative dearth of appropriate data make generalizations of nickel distributions in lakes difficult. The micronutrient-like behaviour observed for N i in marine systems, for example, has not been clearly demonstrated for freshwaters. Available evidence suggests that vertical variability is largely dictated by the redox transformations of Fe, Mn and S. Concentrations of dissolved nickel in the hypolimnion of Lake Sammamish, for example, were shown to increase in response to bottom water anoxia (Balistrieri and Murray, 1992); the supply of dissolved N i to the bottom waters was attributed to the dissolution of nickelbearing Mn-oxide phases. Indeed, Fe and M n oxides have been repeatedly inferred to control the distribution of N i in lake waters, sediments and porewaters (Green et al., 1989; Tessier et al, 1996). Conversely, the removal of nickel from reducing bottom waters (Green et al, 1989) and porewaters (Carignan and Nriagu, 1985) has been largely attributed to the precipitation of authigenic Ni-sulphides.  128  Nickel concentrations in Balmer Lake greatly exceed those in nonperturbed Canadian Shield lakes. Loadings from the tailings circuit, which average over 1 mg-L"l, contribute to the high values. The nickel inventory in the water column is dominated by dissolved species throughout the sampled seasons, which is consistent with measurements of nickel partitioning in other lake systems (Nriagu et al, 1981; Balistrieri and Murray, 1992). Speciation calculations of freshwaters suggest that nickel exists largely as the free aquo ion Ni  2 +  and as carbonato complexes during well-mixed periods in Balmer Lake  (Florence, 1982; Carignan and Nriagu, 1985).  5.1.5.5 Copper  In the oceans, copper exhibits a distribution intermediate between that of nutrient-type elements and that of scavenged elements such as Pb. The vertical distribution of C u in the Pacific, for example, is characterized by an approximately linear increase with depth (Bruland, 1980). The deviation from a nutrient-type element (e.g., cadmium) is due to in situ scavenging by particulates in the intermediate and deep waters. The remineralization of Cu-bearing phases at the sediment-water interface, however, results in a major portion of the total benthic Cu-flux being recycled back into the bottom waters (Bruland, 1980). Deep water concentrations in the Atlantic (Bruland and Franks, 1983) are onehalf to two-thirds of deep Pacific waters due to the progressive addition of dissolved Cu during deep water migration. The compressed depth scales and dynamic nature of lacustrine systems prevent generalizations of the vertical distributions of copper in lakes. However, the available data suggest that redox transformations and particle scavenging represent two important controls. The dissolution of sinking Mn and Fe oxides,  129  for example, can yield peaks of dissolved Cu at oxic/anoxic boundaries (Green et al, 1989). Conversely, hypolimnetic depletions can result from the precipitation of copper-bearing sulphide phases in reducing bottom waters (Green et al., 1989; Balistrieri et al., 1992) and porewaters (e.g., Carignan and Nriagu, 1985). Direct evidence linking C u distributions with biological cycles is rare, although algal decomposition has been invoked to account for Cu enrichments at the sedimentwater interface in a British lake (Morfett et al., 1988). Filterable copper species dominate the Balmer Lake water column throughout the sampled seasons, comprising between 65 and 90% of the total pool. Speciation calculations suggest copper exists predominantly as C u O H  +  and CuC03 in aerobic fresh waters (Stumm and Morgan, 1981), with a significant portion of the dissolved fraction potentially bound to organic ligands. Inorganic and organic colloids have also been shown to contribute significantly to total copper burdens in freshwaters (Dai et.al, 1995).  5.1.5.6 Zinc  Zinc is a group lib element which has well-established associations with biological cycles in both marine and lacustrine environments. In the oceans, zinc exhibits a nutrient-type distribution similar to that of silicic acid; dissolved concentrations in the central North Pacific increase from -0.2 nmol-L" in the 1  surface mixed layer to -8 nmol-L in the intermediate and deep waters (Bruland -1  et al, 1994). Although biologically-mediated Zn-nutrient associations are often obscured by other processes in lakes, two convincing data sets clearly illustrate a biological control. In productive Lake Windemere, for example, dissolved zinc, phosphate and silicate were positively correlated throughout the water column during a rapid growth phase of a spring diatom bloom (Reynolds and Hamilton-  130  Taylor, 1992). Similarly, acid-soluble zinc concentrations were shown to correlate with silicate distributions during summer stratification in Lake Sammamish, Washington (Balistrieri et a l , 1992). Although there is little doubt that Fe and M n oxides also act as carriers for Zn in lacustrine systems, observations indicate that redox related phenomena have a relatively small effect on the vertical distribution of dissolved Zn in lakes. Conversely, numerous examinations of the early diagenesis of zinc in oxic lake sediments have demonstrated the importance of Fe and M n oxide phases with respect to zinc mobility (Carignan and Tessier, 1985b; Tessier et al, 1989; Young and Harvey, 1992; Tessier et al, 1996). In reducing sediments, dissolved zinc is commonly removed from pore-solution via the precipitation of sphalerite (ZnS) and/or mixed Fe-sulphide phases (Carignan and Tessier, 1985b; Pedersen et al, 1993). The zinc concentrations in the water column of Balmer Lake greatly exceed those in unpolluted freshwater systems. Although dissolved fractions dominate the zinc burden, a considerable portion is present in suspended particulates. The filterable zinc species are predicted to exist predominantly as Zn  2 +  and ZnC03, with inorganic and organic colloids potentially hosting a  significant portion (Florence, 1982).  5.1.6 Water Column Model  The following section presents a model to explain the water column profiles of trace elements over the fall-winter-spring transition. The scheme, presented in Fig. 5.1, represents a time series, and encompasses the broad range of profiles observed in Balmer Lake.  131  A  C  B  Fig. 5.1. Time series of progressive trace metal water column profiles (A-E) over the fallwinter transition in Balmer Lake: A) well-mixed water column, B) influence of remobilized bottom-water source, C) influence of interfacial sulphide sink, D) greater extent of sulphide removal, and E) influence of metal-rich lateral advective flow. Depth arid concentration axes are in arbitrary units.  132  As described in section 1.2.3, the degradation of organic matter is mostly an oxidative process, for which electron acceptors must be present. Remineralization follows a sequence of oxidation reactions in which the redox couple that liberates the highest free energy yield acts before the next most efficient couple. Bacterial assemblages, specialized in using the energy released in each of these redox reactions, obey this sequence (Froelich et al, 1979). The thermodynamic sequence of oxidant depletion follows the order: O2, NO3" and Mn(m,IV)-oxides, Fe (Ill)-oxides, SO4" and finally C 0 . Deviations of this 2  2  sequence are possible, however, and relate to such factors as porewater p H and the stability of the iron oxide (Postma and Jakobsen, 1996). During the well-mixed, oxygenated periods of the spring, summer and fall in Balmer Lake, the anoxic-oxic redox boundary (as defined by the depth where SO4" reduction commences) lies below the sediment-water interface. 2  The  formation of ice in early winter, however, creates the conditions necessary to promote the gradual consumption of the oxidant pool, and the evolution of bottom water anoxia.  The SO4 " redox boundary is predicted to migrate 2  upwards towards the sediment surface over the duration of ice cover, and in some areas, cross the interface into the bottom waters. Examination of the water column and interfacial metal distributions measured in late winter suggests that extensive sulphide precipitation had already occurred earlier that season in the deeper lake areas. Authigenic sulphide formation has been suggested to present a significant control on trace element distributions during anoxic periods in other dimictic lakes (Morfett et al, 1988). Discharges to Balmer Lake from the adjacent tailings circuits present another important consideration. Measurements made on tailings pond effluents suggest that significant burdens of trace metals, sulphate, cyanide and ammonia are associated with these discharges. The influence of mining-related effluents in  133  the lake appears to be related to water depth, and to the proximity to the point of discharge.  Due to its large signal in mine discharge waters, sulphate is  considered to be the most appropriate tracer of these inputs. During well-mixed periods, homogenous conditions resulting from extensive wind-mixing of a shallow water column produce uniform depth profiles of dissolved constituents as was shown in Chapter 4, which are summarized schematically in Fig. 5.1 A. The formation of ice in early winter removes any atmospheric influences and permits the development of water column stratification.  As anoxia develops over the course of ice-cover, the  predicted reducing swing in the redox environment at the sediment-water interface promotes the reductive dissolution of labile Mn-bearing solid phases. Significant input to bottom waters subsequently yields the generalized profile shown in Fig. 5.IB. After the consumption of NO3" and Mn-oxides, the oxidation of organic matter proceeds via Fe(III) reduction. The most reactive phases are those with the highest solubility, i.e., amorphous and poorly crystalline Fe-oxyhydroxides (Van der Weijden, 1992).  Dissolved Fe distributions during such periods  resemble the profile shown in Fig. 5.1B. The remobilization of Fe and M n oxyhydroxide phases is often associated with the concomitant release of sorbed trace elements which may yield similar enrichment patterns of such elements in bottom waters (e.g., Green et al, 1989; Aggett and Kriegman, 1988). As conditions become more reducing at the sediment-water interface, and reactive Fe(III)- and Mn(IV)-bearing solid phases have become exhausted, sulphate reduction is expected to become important (Froelich et al, 1979). The introduction of an interfacial metal-sulphide sink to the situation shown in Fig. 5.1B, may yield a profile similar to that shown in Fig 5.1C. More extensive sulphide precipitation, in conjunction with diffusional processes, may remove  134  any evidence of prior metal release (Fig. 5.1D). The final profile in the model sequence results if a dense, metal-rich lateral flow displaces or mixes into metal depleted bottom waters (Fig. 5.1E). This five-step order of events is used below to assess the intra-lake variability during the winter sampling period.  5.1.7 Comparison of Four Lake Zones  To reiterate, the trace metal distributions observed during the spring, summer and fall sampling periods exhibited little variability with depth; windy conditions prevented the development of any thermal stratification and, in effect, inhibited the detection of bottom water gradients associated with sediment consumption or release.  The relatively stagnant conditions during winter  sampling, however, permitted the development of chemical stratification. The fact that Balmer Lake has a large surface area to volume ratio enables benthic exchanges to be more easily manifested in the water column structure. The •following discussion attempts to resolve the winter variability of trace metal distributions with respect to four lake areas (stations 1-4, Fig. 2.1). Evidence for interfacial sulphide precipitation throughout Balmer Lake implies that the reductive dissolution of labile sedimentary Mn- and Fe-oxides, and associated metal releases, had occurred some time prior to sampling. It should be noted that at any time during the development of reducing conditions, the redox intensity at the sediment-water interface may exhibit significant lakewide variation; differential rates can be attributed to such major variables as water depth and the concentration of organic matter. This point will become relevant in the following interpretations of water column profiles. Steep porewater manganese gradients evident during the other seasons imply that the upward diffusion of remobilized Mn(II) species across the  135  interface is at least partially responsible for enrichments in the reducing winter bottom waters.  The fact that stations 3 and 4 exhibit minimal or negligible  signatures from pollutant effluents (i.e., insignificant SO4 " signal) and are still characterized by bottom water Mn maxima, supports this premise (Figs. 4.7,4.10 and 4.11). The decoupling of the soluble and particulate M n fractions seen in Balmer Lake results from the relatively slow oxidation kinetics of its reduced form; this results in greater stability and thus longer residence time of reduced M n species in oxygenated waters.  However, in aquatic systems where M n oxidation is.  substantially accelerated by bacterial processes, or where residence times are long (i.e., some lakes), the interconversion of the two forms may be an important process. Indeed, the pronounced decrease in the dissolved M n content, which occurred over the four week sampling hiatus between the winter and spring sessions, implies a removal mechanism from the water column. The oxidation of Mn  2 +  and subsequent precipitation as M n 0 provides a probable mechanism. 2  Bottom water concentrations of M n exceed 1 m g / L in some zones of Balmer Lake. The release is predicted to be predominantly present as reduced Mn(II) hydrated aquo ions, since speciation studies suggest that inorganic and organic complexes represent insignificant fractions (Chiswell and Mokhtar, 1986; LaZerte and Burling, 1990). Manganese enrichments in the anoxic hypolimnia of lakes in excess of 1 mg'L" have been reported by many authors (Davison et al, 1  1982; Sigg, 1985; Cornwell, 1986; Agett and Kriegman, 1988; Morfett et al, 1988; Anderson and Bruland, 1991). The seasonal anoxia in these studies, however, occurs in the summer, and results from the development of thermal stratification in combination with enhanced organic transport to the hypolimnion. Sulphate reduction in the bottom waters also implies that the reductive dissolution of reactive Fe phases had occurred prior to winter sampling.  136  Interfacial enrichments of dissolved Fe derived from remobilized sediment sources have been reported in several studies of seasonally anoxic lakes (e.g., Davison et al, 1982; Cornwell, 1986; Anderson and Bruland, 1991; Sigg, 1985; Agett and Kriegman, 1988). The extremely high metal levels in Balmer Lake, and thus their increased ability to compete for surface sorption sites, may suggest that Fe and M n amorphous phases in the top few centimetres host significant burdens of exchangeable metals. Therefore, the reductive dissolution of Fe and M n oxides may promote a pronounced release of sorbed constituents (Belzile and Morris, 1995). Moreover, trace metals associated with conspicuous porewater maxima (i.e., Fe, Mn, As, Zn and to a lesser extent Ni) could diffuse into the bottom waters in the absence of an interfacial oxide trap. In addition to Mn(II), Fe(II), and associated trace elements remobilized from reducing sediments, the reductive dissolution of sinking oxide particles in reducing horizons has also been invoked to support enrichments of metals in the bottom waters of stratified lakes (Green et al, 1989; Anderson and Bruland, 1991; Balistrieri and Murray, 1992). This pattern of supply to the surface waters and remobilization at depth, may lead to maxima in the concentration of soluble constituents towards the end of seasonal stratification (Davison, 1993). The abundance of metal particulates in the winter water column of Balmer Lake, suggests that trace metals remobilized from sinking particles may present an important input to the bottom waters.  Because the lower stratified zone in  Balmer Lake is < l m thick the residence times of particulates in this zone are predicted to be brief, suggesting that remobilization occurs predominantly at the sediment-water interface. In general, considerable difficulty arises when trying to distinguish between diffusion across the sediment water interface and dissolution of settling  137  oxide particles as the source of increased concentrations. In an earlier review, Sholkovitz (1985) demonstrated that enhanced redox-related effluxes were accountable for increased bottom water concentrations of Na, K, Mg, Ca, Ba, Fe, Mn, P, Si and NH3 in some seasonally anoxic lakes. However, evidence showing similar effects with other elements is scarce. Visible dissolved trace metal gradients associated with bottom water sources have also been attributed to the remineralization of organic matter (Morfett et al., 1988; Hamilton-Taylor and Willis, 1990; Reynolds and HamiltonTaylor, 1992). In the oceans, many trace elements (e.g., Cu, Zn, Ni, C d ) display vertical concentration distributions strongly correlated with those of the major nutrients; they are depleted in the eutrophic zone by incorporation onto or into planktonic material, carried out of the surface waters by settling, and partly regenerated at depth through decomposition of the organic particles (Bruland, 1980). However, distinct correlations evident between concentrations of P and Si and several trace elements that have been observed in the oceans are rare for lake systems (Balistrieri, 1992). The absence of such correlations may reflect the contrasting environments of marine and freshwater systems, but may also reflect a lack of appropriate lacustrine data (Reynolds and Hamilton-Taylor, 1992). Metal inputs to terrestrial receiving waters are typically greater than loadings to seawater, and given the relative shallowness of lakes, little vertical segregation may develop between the biological scavenging of metals in near surface waters and their partial release at depth during respiration (Sigg 1985). Sigg (1985) and Morel and Hudson (1985) have suggested that algae incorporate trace elements stoichiometrically, resulting in the approximate formula C106N16P1 (Fe, Zn, Mn)o.oi (Cu, Cd, N i , Co, etc.)o.ooi-  Although this  concept of Redfield ratios may be applicable to the metal content of algae, the dynamic nature of lakes limits its usefulness because only under severely  138  restricted conditions will the stoichiometry be reflected in dissolved metal concentrations (Reynolds and Hamilton-Taylor, 1992). Furthermore, the scheme proposed here remains hypothetical, as some elements that have shown correlations with nutrients have no clear biological role. The mechanisms described above (i.e., trace element remobilization from sedimentary and settling particulate oxide phases, and organic matter remineralization) are proposed to account for the Ni, Cu, As and Zn enrichments observed in the winter bottom waters at station 3 (Fig. 4.10). The relative partitioning of these elements will be discussed in the porewater sections below. Station 3 has the shallowest water column (2.8 m deep), and appears to lie above the depth of influence of dense mining-related effluents (Fig. 4.7C). Several of the dissolved trace metal profiles at this site show evidence of a bottom water source and generally resemble Fig. 5.1B. The slight contrasts among bottom water gradients of dissolved Mn, N i , As and Zn, may reflect kinetic variability with respect to the oxidation and precipitation of upward diffusing species (Fig. 4.10). Although corresponding peeper data were not collected from this site, sulphide precipitation appears to be a less significant control on trace metal distributions. Indeed, the less pronounced nitrate removal with depth suggests that reducing conditions are in a less advanced stage of development as compared to deeper lake areas (Fig. 4.5C). The removal of trace metals as authigenic sulphides i n anoxic lake systems has been reported for several elements, including Fe, As, Cd, Cu, Ni, Pb and Zn (Davison et al, 1982; Morfett et al, 1988; Green et al, 1989; Davison et al, 1992; Balistrieri, et al, 1992). Evidence for metal precipitation stems largely from calculated ion activity products using measured values of dissolved constituents in the water column and porewaters. Although sulphide was neither measured nor observed (by smell) in Balmer Lake, the abundant reducible sulphate in the  139  anoxic winter bottom waters, and the observed trace metal distributions, suggest that sulphide removal presents a principal control on trace element distributions in the deeper lake sites. In a more reducing area of Balmer Lake (station 4), sulphide precipitation mechanisms become more relevant. This site (3 m deep) is slightly deeper than station 3, and appears to be only marginally affected by mining-related advective flows (Fig. 4.7D). The reduced influence of such discharges is thought be a result of the shallow depth and its relatively large distance from effluent inputs (Fig. 2.1).  Profiles of dissolved Fe, As, N i and C u at this site exhibit pronounced  decreases below deep water column maxima (Fig. 4.11) and, with two exceptions (Mn and Zn), generally resemble Fig. 5.1C. The removal zones for N i and C u appear to be up to 5-10 cm above the sediment surface (Figs. 4.25D and E). The insignificant decreases in dissolved Fe over these horizons in comparison to other trace elements suggest that co-precipitation with FeS is not significant; rather, precipitation as their respective sulphides (i.e.,  AS2S3,  NiS and CuS,)  presents a more likely mechanism. Prior to the onset of sulphate reduction, the near-bottom distributions of dissolved Fe, As, N i and Cu may have been similar at this site. However, it is suggested that sulphide removal has since consumed part of the excess of remobilized dissolved trace metals to produce the observed profiles. The fact that diffusional processes have not eroded the prominent dissolved maxima indicates that a steady input of dissolved species (presumably remobilized from sinking particulates) is being supplied to these horizons. Similar water column profiles of various dissolved constituents in permanently stratified Lake Vanda, Antarctica (60 m deep), were attributable to such processes (Green et al., 1989). However, due to the extreme compression of depth scales in Balmer Lake, and the relatively shallow depth of the bottom stratified  140  layer, particulate metal distributions analogous to those in Lake Vanda are not observed. Throughout the lake during the winter period, dissolved M n concentrations exhibit no evidence for sulphide removal, as expected given the high solubility of MnS (Emerson, 1976). Indeed, the Mn distributions suggest a continual supply of dissolved species, which are presumably remobilized from sediments and settling particulates. Although dissolved Zn profiles exhibit pronounced lake-wide removal at shallow sediment depths (Figs. 4.23-4.25), authigenic sulphide precipitation above the sediment-water interface does not appear to be relevant. Strong associations with organic matter have been suggested previously to explain the decoupling of Zn from the redox cycling of Mn and Fe in a seasonally anoxic lake (Morfett et al, 1988). The apparently strong associations of P O M with Z n in Balmer Lake (section 5.1.4; Tables 4.5 and 4.7) may suggest that organic matter remineralization is contributing to the persistent bottom water enrichments. The reductive dissolution of zinc-bearing M n oxides may also serve as an input mechanism of dissolved Zn (Sigg et al, 1987). Examinations of the deeper sites (stations 1 and 2) illustrate the greater extent of sulphide reduction in the bottom waters as well as the local importance of deep lateral advection. At station 1, dissolved Fe, As and C u gradients begin at the 2.5 m depth horizon (Fig. 4.8). The general similarity of this trio of profiles suggests similar controls on the distributions of the three elements in the nearbottom waters. Peeper profiles imply that Cu removal begins 10-15 cm above the sediment-water interface (Fig. 4.23E). While Cu and Fe concentrations continue to decrease to the sediment surface (Fig. 4.8, summarized in Fig. 5.1 D), As shows an opposite distribution below 2.8 m (Fig. 4.23C). Such an increase, as well as those for Fe, As, N i at station 2, can be perhaps attributed to the influence of  141  relatively metal-rich advective flows (Figs. 4.23 and 4.24; see Fig. 5.IE). At both stations 1 and 2, bottom water SO4" increases to values over 12 mmol'L" below 2  depths of 2.7 m (Fig. 4.7).  1  Further evidence for lateral flows lies in the  observation that sulphide precipitation alone could not account for the water column minima; i.e., sulphide removal acting near the sediment-water interface could not produce a minimum 40-70 cm above the sediment surface. Furthermore, if the bottom sediments were providing a metal source, and the water column minima were the result of trace metal scavenging at an oxic front, one should see evidence of metal release from the porewaters. However, nearinterface trace metal profiles for Fe, As and N i at station 2, and As at station 1, offer little support for diffusive effluxes fostered by sub-interface release (Fig. 4.24 and 4.23). In addition, profiles of the respective dissolved and particulate metal fractions do not support the removal of upward diffusing species by incorporation into Fe- or Mn-oxide precipitates (Tables 4.6 and 4.7).  5.2 Sediment Geochemistry  The sediments in Balmer Lake represent complex mixtures of lithogenic, biogenic and authigenic phases which vary in time and space. For example, the composition of settling particulate matter varies over the course of a year due to seasonal fluctuations in primary productivity, ice coverage and terrestrial runoff. In addition, significant heterogeneity has been introduced via episodic events of lake-wide tailings deposition. Organic matter in the deposits stems both from in situ lake production of phytoplankton and macrophytes, and from external inputs via streams and wind-borne transport which vary with the extent of runoff. Prior to the onset of mining activity, allochthonous inputs from Balmer Creek (Fig. 2.1) represented the only significant lithogenous contributions to the  142  lake. However, since the commencement of mining, tailings have become the major detrital input. Profiles of several compositional parameters illustrate that below approximately 10 cm, the sediments at stations 1 (cores la and lb) and station 6 (core 2) represent two discrete assemblages and can be clearly seen in the weight ratios of M g / A I (Fig. 4.32C), N a / A I (Fig. 4.32E), R b / A l (4.33B) and Z r / A l (4.33C). Those deposits in the lower horizons at station 1 consist of pre-industrial natural deposits, while sediments in core 2 are predominantly mining-derived tailings material throughout. Above the 10 cm transition horizon, the deposits at both locations appear to have received variable contributions from both anthropogenic and natural sources. The organic carbon content in Balmer Lake sediments of 8-9 wt.% (Fig. 4.34A), agrees with previously measured values and fall within the ranges observed in other productive Canadian Shield lakes (Conroy and Keller, 1976; Semkin and Kramer, 1976; Carignan and Tessier, 1988). Biological activity in Precambrian Shield Lakes has been shown to be influenced by morphology (surface area/volume), lithology (surficial and bedrock geology) and atmospheric loadings (Conroy and Keller, 1976). Water bodies such as Balmer Lake, with high surface area to volume ratios and situated in calcareous terrain, tend to exhibit higher productivity. Lower production characterizes acid lakes (pH<5.5) that have been impacted by acid rain, are poorly buffered, and have high volume: surface area ratios. Total nitrogen is enriched in the surface sediments of both stations 1 and 6 (Fig. 4.34C) and is illustrated in the depth distributions of the respective C r g : N 0  wt. ratios (Fig. 4.34D). Lower C o r g / N values in the upper horizons are most likely the result of increased inputs of inorganic-N compounds and not the result of variations in the C:N ratio of introduced organic matter. Increased inputs of  143  terrigenous organics, which are commonly associated with lake-side industry (e.g., forest clearing, etc.), would have resulted in a higher C:N signal. The observed distribution of N can be explained by a combination of two relatively recent inputs. First, the hydrolysis of cyanide introduced by the adjacent ore processing operations has increased the abundance of nitrate and ammonium in the lake. Second, accidental dam breaches have supplied significant loadings of fine-grained, chlorite-rich material to the inherently clay-deficient natural deposits (PDI, personal communication). The exchange capacity of chlorite for cations such as NH4+, C s , R b and K is comparable to that of illite (Grim, +  +  +  1953). Therefore, the mineralogical shift induced by the addition of chlorite could have contributed to the relative abundance of particulate nitrogen in the upper decimetre of the Balmer Lake deposits. The presence of clay-rich deposits in the upper of horizons was established by semi-quantitative X-ray diffraction and examinations of element/Al wt. ratios. XRD peak-height ratios suggest that the quartz content in the upper stratum at station 1 is higher than at depths below 10 cm at this site. However, an increased quartz content cannot represent the sole mineralogic contrast in the upper decimetre because S i / A l wt. ratios decrease in this stratum (Fig. 4.32A). The recent addition of a relatively Al-rich mineral phase to station 1 presents the only plausible explanation that can account for the lower Si:Al ratios. XRD analyses indeed confirm the presence of abundant clay minerals, notably chlorite, in the upper horizons. The data collectively imply that the deposition of detrital chlorite (Si:Al = 1.3) has overwhelmed the input of quartz. The high M g content in chlorite and the higher Mg:Al wt. ratios in the upper stratum at station l(Fig. 4.32C) provide further support for a chlorite input. In contrast to station 1, Si:Al ratios in the upper horizons at station 6 imply a  144  relatively higher quartz content. This is supported by the XRD data: quartz/clay peak height ratios in deposits at station 6 are higher than those at station 1. The locations of historic tailings discharges with respect to the coring locations could be expected to have exerted significant control on the composition of mining-derived material that reached the respective locales. Station 6 is situated tens of metres.from the major points of tailings discharge while station 1 sits over 1 km away. Deposits immediately proximal to such outputs would be expected to be relatively coarser-grained than those more distal. Such textural influences could contribute to the mineralogical differences in the tailings-containing deposits in the two lake areas. Below depths of 10 cm at station 1, sediment concentrations of all trace metals are low, and are comparable to values typical of Canadian Shield lakes (Semkin and Kramer, 1976). In contrast, the lowest concentrations of Ni, Zn, Cu, As and Pb observed in the tailings deposits are still 4, 2, 2, 20 and 10 x greater, respectively, than the natural contents (Fig. 4.35). The upper horizons at both locales, which are characterized by overall increased metal inventories, are punctuated by dramatic metal spikes; however, those at station 1 (i.e., most distal from the tailings discharge) are an order of magnitude greater than comparable enrichments at station 6. This implies that higher metal contents are associated with the more widely dispersable "slimes" fraction of the tailings. Pedersen and others (1993) proposed a textural argument to account for metal distributions observed in Anderson Lake, Manitoba, into which tailings have been deposited since 1979. In their study, a coring site over 1 km distant from the tailings discharge was characterized by bands of metal-rich sediments in the upper facies that had Zn, Cu and Pb concentrations an order of magnitude greater than those observed in pure bulk tailings deposits.  These horizons  hosted ~2 wt.% C u , -12 wt.% Zn and -15 wt.% S, and were marked by high  145  F e / A l ratios  (-4:1),  suggesting metal-pyrite and metal-sulphide associations.  The sulphur distribution at station 1 in Balmer Lake sediments is not clearly matched with the observed metal spikes (Fig. 4 . 3 4 E ) . A broad S enrichment of 1.5 wt.% between 3 and 8 cm, and a minor surficial increase do correlate coarsely with N i , Zn, C u and As profiles. Assuming sulphide stoichiometries of ZnS, NiS and  AS2S3,  Cu2S,  the sulphur inventory present is more than needed to  account for the metal enrichments if they occur as their respective sulphides. F e / A l profiles at station 1, although characterized by general increases in the top stratum, correlate poorly with the observed metal spikes (Fig. 4 . 3 4 F ) . However, the large Fe signal contributed by the clay fraction of the tailings input would potentially overwhelm any contribution of Fe associated with the observed metal spikes. A coarse calculation comparing the S accumulation rate and the downward diffusive flux of sulphate was performed in order to test whether the observed distribution could possibly represent an authigenic sulphide precipitate (Appendix E). The lake-wide averaged diffusive influx of sulphate-S was calculated to be within an order of magnitude of the S accumulation rate, suggesting that in situ sulphide precipitation may partially account for the observed S distribution at station 1. Carignan and Tessier (1988) also observed sedimentary S maxima between 2 and 8 cm below the interface in eight sulphate-enriched Candadian Shield lakes.  Such values, which ranged from  0.2-3.4  wt.%, were attributed to  diagenetic enrichments arising from the downward diffusion of SO4 " followed by its fixation as various S species. Reduced inorganic (e.g., Fe-mono sulphides, pyrite, and elemental S), rather than organic sulphur compounds constituted the major proportion of the total S in these sediments. Organic forms of S were also  146  shown to be insignificant in polluted lakes near Sudbury, Ontario (Nriagu and Soon, 1985). Seasonal profiles of porewater Mn distributions (Section 5.3) indicate that considerable diagenetic cycling occurs in the near-surface sediments of Balmer Lake. Manganous ions remobilized in the reducing porewaters migrate upwards where they may precipitate as manganese oxide in aerobic sediments. At steadystate, the reducing horizon will migrate upward at the rate of linear sediment accumulation. Under such conditions, the cycle of dissolution and precipitation results in progressive solid-phase M n enrichment above the M n oxide redox boundary (Klinkhammer, 1980). In the organic-rich, highly reducing sediments in Balmer Lake, however, the thin veneer of oxic sediments limits the extent to which M n can accumulate. For this reason, the broad, deep, solid-phase M n maximum observed at station 1 does not likely represent a diagenetic profile. The M n enrichment probably stems from the addition of a tailings-derived carbonate or sulphide phase.  The sedimentary geochemistry of M n will be  discussed in greater detail in section 5.3.3.  5.3 Porewater Chemistry  Examination of trace metal porewater distributions (Appendix C), in conjunction with the various metabolite profiles associated with the degradation of organic matter, provides an effective way to study metal behaviour and mobility in sediments.  The dialysis-array samplers used in the Balmer Lake  project yielded high resolution profiles of dissolved constituents across the sediment-water interface, and these are usually sufficient to allow reasonably accurate determinations of the direction and extent of diffusive trace metal fluxes. Furthermore, the high-resolution information provides good support for  147  the interpretation of post-depositional processes of trace metal consumption and remobilization. Various tools can be employed to assist interpretations of porewater profiles.  Saturation indices, for example, are often used to elucidate  solution/precipitation chemistry in marine and lacustrine sediments (e.g., Carignan and Nriagu, 1985; Balistrieri et al., 1992). However, caution must be exercised when comparing the measured ion activity product with the solubility product to deduce the tendency of a solid phase to form, as the kinetics of solid phase formation are often slow, and the measurement of soluble species (e.g., by filtration) may include colloidal forms (Davison, 1993). Furthermore, metal complexation with ligands in solution or those associated with particle surfaces can also promote deviations from predicted equilibria. Due to the lack of reliable thermodynamic data, most studies consider simple mineral phases in solubility calculations.  This can result in further  misinterpretation, as the formation of mixed phases has been proposed to control the concentration of numerous species (Suess, 1979).  Various programs,  including MINEQL, and H Y D R A Q L (Papelis et al, 1988), are often used to constrain chemical speciation and equilibrium in such solubility calculations. Examination of the dissolved trace metal and metabolite porewater profiles in Balmer Lake sediments demonstrate that dramatic changes in the chemical environment occur across the sediment-water interface.  Sharp  gradients for most constituents appear at or shortly below the sediment surface throughout the lake during all seasons. The shapes and depths of trace metal and metabolite profiles, in combination with solid phase distributions, are used below in the interpretation of the porewater chemistry in Balmer Lake. In particular, zones of non-linear concentration gradients delineate the depths of trace metal consumption and  148  remobilization. The absence of porewater HS", PO4 " and HCO3" measurements 2  prevented calculation of the solubility of various authigenic trace metal precipitates. However, due to the high spatial resolution of the profiles and the well-constrained thermodynamic transitions defined by NO3", Mn(II), Fe(TJ) and SO4 " data, a reasonable degree of confidence can be drawn in the predictions of trace metal behaviour. Diffusive fluxes, and hence estimates of reaction rates, were calculated from linear concentration gradients (Appendix F) using Fick's First Law:  J = z  <t>  F  dc dz  2  1  where J = flux; D ° = in situ diffusion coefficient in cm 'sec (Li and Gregory, z  1974); F = formation factor = 1.4 for silty clay (Manheim, 1970); this takes into account the tortuous diffusion path of an ion in wet sediments; 0 = porosity = 0.87, calculated using the measured water content and an average particle density of 2.65 gem" ; and dc/dz = concentration gradient in gem" . 3  4  It is  assumed that porosity and diffusion coefficients are invariant with depth. Diffusion coefficients were not corrected for any random transport mechanisms such as biodiffusion, gas ebullition or wave mixing.  5.3.1 Ammonium and Nitrate  The cycles of the essential elements, C, N , and P, are coupled because they are incorporated into the living biomass of primary producers in relatively constant ratios. The complete mineralization of this biomass must therefore lead +  to the regeneration of the nutrients, H C 0 , N H 3  4  2-  and H P 0  4  in a similar  stoichiometric proportion (Redfield, 1958). The generally concave-downward 149  profiles in Balmer Lake sediments suggest that N H  is actively being liberated  + 4  throughout the sampled sediment column (Figs. 4.12-4.14; 4.17)  The large  porewater N H / gradients evident during the spring, summer and fall, support upward diffusive fluxes towards the sediment water interface. Efflux estimates 0 1 over these periods range from 0.2-1.2 mmolm" d" (Appendix F). The bottom water mining-derived enrichment in the shallow waters of "Tailings Bay" (station 5) in the fall, however, fosters a downward influx of NFL; into the +  sediments (Fig. 4.13D). The combination of a deeper remineralized sediment source at this site results in bi-directional diffusion toward the interface. The nitrification of upward diffusing NFL; at oxygenated horizons during +  the spring, summer and fall, represents the principal NFL;* sink in the interfacial sediments and bottom waters (Froelich et al., 1979).  Oxidation is usually  considered to proceed via the reduction of molecular O2 and Mn(IV) by nitrifying bacteria (Van der Weijden, 1992). From N H / profiles, it is generally assumed that NH4 is not oxidized, directly or indirectly, by Fe(III) oxide (Froelich et al., +  1979). Lower C:N ratios in the upper sediment stratum, however, suggest a portion of the remobilized NFL;* sorbs to the abundant clay fractions or other phases in these horizons (see section 5.2). Over the fall-winter transition, considerable shifts occur with respect to the distribution of porewater and bottom water NH4 (see section 5.1.3). When +  the lake becomes ice-covered and anoxia develops at depth, NH4 is predicted to +  diffuse across the sediment-water interface into the bottom waters.  Diffusive  effluxes are expected to be greatest in the earliest stages of anoxia, and to wane as diffusive equilibrium develops between interstitial and bottom waters. At the time of sampling in late winter, development of such an equilibrium was anticipated to be in an advanced stage.  If the largest diffusive NFL;* flux  calculated above of 1.2 mmolm" d" remained constant over time, it would take 150  approximately 170 days to produce a 200 uiriol'L" enrichment in the lowermost 1  metre of the water column. Although this calculation represents a maximum, it suggests that diffusive processes significantly affect bottom water concentrations during the seasonally anoxic period. The relatively minor influences of mining inputs at station 4 make this site the most suitable to assess winter benthic exchanges due to natural processes (Fig. 4.16A). Gradients normally observed in the sediments migrate above the sediment surface during the winter at this site; the N H / profile reflects diffusive equilibrium across the interface. The concave-upward profile between 5 and 20 cm above the sediment surface implies oxidative removal of N F I / in this zone. Similarly, a concave-upward profile for N0 " mirrors that of NFL;"", suggesting 1  3  active nitrification (4.16B). The seasonal N H / distributions at stations 1 and 2 have been plotted together to illustrate the variability over the course of a full year (Fig. 5.2). Note that spring profiles were not obtained for station 2. The spring, summer and fall profiles clearly illustrate ammonia production below the interface; these contrast strongly with the generally invariant N H i distribution in the winter. The zone +  of oxidative removal migrates above the sediment surface in winter, and is characterized by a dramatic decrease in the concentration gradient. A n unusual winter maximum above the sediment-water interface is proposed to reflect a non-steady-state input associated with mining-related discharges. inspection of the winter N H  + 4  Closer  distributions at stations 1 and 2 (Figs. 4.15A and D)  illustrates the interfacial variation at these sites. It is suggested that variations with respect to the timing, frequency and magnitude of mine-related inputs can account for such variability. The fact that the N H  + 4  concentrations are slightly  higher in bottom waters at station 1 in winter implies that a minor ammonium influx to the sediments exists temporarily under the ice at this location.  151  Fig. 5.2.  Seasonal peeper profiles of ammonium for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 152  Nitrate profiles during the spring, summer and fall exhibit remarkable inter-season and inter-site consistency (Fig. 5.3).  Vigorous nitrate reduction  begins 2-5 cm below the sediment-water interface, resulting in downward fluxes ranging from 0.3-1.3 mmol N'm" d" during the well-mixed periods (Appendix F). Such implied denitrification rates are similar to those observed in organic-rich sediments of other temperate lakes (Rudd et al., 1986), and lie intermediate to 2  1  denitrification rates observed in aerobic deep-sea sediments (3-10 umol N'm" d") 2  1  and eutrophic coastal marine systems (7 mmol N'm" d") (Jenkins and Kemp, 1984). The oxidation of upward diffusing NFL; appears to account for the +  interfacial NO3" spikes seen particularly in summer at sites 1 and 2 (Fig. 5.3). Coastal marine sediments commonly have a sharp NO3" peak close to the sediment surface, from which NO3" is lost to the bottom waters via upward diffusion (Emerson, 1976).  In this case, the peaks apparently occur at or  immediately above the interface, implying that the sediment surface is the site of extremely active nitrification in summer. The winter data indicate that NO3" reduction begins 7-15 cm above the sediment-surface at stations 1,2 and 4 (Figs. 4.15 and 4.16; see section 4.1.3). The near-bottom anoxia implied by these profiles is not illustrated by the nitrate profiles at the shallow (2m) station 6 (Fig. 4.16E), indicating that anoxia is restricted to deeper lake areas.  5.3.2 Sulphate  After aerobic respiration, SO4 " reduction is, on a global scale, the most important process in the diagenesis of organic matter in marine systems. However in lakes, sulphate reduction is relatively insignificant given the dearth  153  Fig. 5.3. Seasonal peeper profiles of nitrate for stations 1 and 2 (A and B, respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 154  of dissolved SO4" (usually < 100 u.mol'L'1). This generalization does not apply to 2  Balmer Lake, however, where relatively recent anthropogenic additions of SO4" 2  have undoubtedly shifted the respective proportions of organic matter oxidized by O2 and SO4 ". This becomes particularly relevant during periods of bottom water anoxia, during which the absence of other oxidants at the sediment-water interface implies that organic matter remineralization proceeds via SO4" 2  reduction alone for possibly several months per year. The dialysis-array data nicely depict the temporal mechanisms that govern the seasonal distribution of SO4 ". In winter, large external SO4 " inputs 2  2  impart a diffusive influx of SO4" into the sediments; the topmost porewaters host 2  in excess of 8 mmol'L" during such periods (Figs 4.15 and 4.16). In spring, 1  average water column values are greatly reduced (to ~ 3.5 mmol'L" ) as a result of 1  water column mixing and an increase in lake volume. The subsurface maximum in spring represents the remnant of the interfacial SO4" excess seen in the late winter (Fig. 5.4).  The reduction of SO4 " in the porewaters, combined with 2  diffusional transport, indicate that the sub-surface peak is in a state of progressive decay. Indeed, early summer profiles at station 1 exhibit a highly attenuated SO4 " maximum (Fig. 4.12C), and no evidence of the peak remains by fall (Fig. 4.13C). Diffusive influxes of sulphate over the course of the sampling year ranged O -I  from 0.5-2.4 mmol'm" d" (Appendix F). The lower end of this range agrees well with  Rudd et al. (1986), who reported reduction rates ranging from 0.1-0.4  mmolm" d" in organic-rich sediments in several temperate lakes. The upper limit in Balmer Lake sediments represents enhanced transport across the interface due to the large gradients imposed by sulphate-rich effluents.  The  depth where sulphate reduction commences in the Balmer Lake deposits, defined where d [S0 2  ]/dz > 0, varies seaonally. 2  4  155  The reduction horizon appears to  Fig. 5.4. Seasonal peeper profiles of sulphate for stations 1 and 2 (A and B , respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 156  migrate towards the interface through the summer-fall-winter transition, 2-  indicating that S 0  4  is being consumed at progressively shallower depths. The 2-  presence of abundant electron acceptors other than S 0  4  (e.g., reactive Fe(IJJ) and  Mn(IV)) may sustain microbial processes in the upper sediments during the early summer. However, enhanced production and build-up of organic matter over the summer would cause labile Fe and M n species to become progressively 2-  depleted, thus favouring the consumption of S 0  at shallower sediment depths.  4  2-  Although dissolved C u profiles suggest S 0  4  is being reduced several  centimetres above the sediment-water interface in winter (Fig. 4.23-4.25), non2-  linear S 0  4  concentration gradients only become obvious below the sediment 2-  surface (Fig. 5.4).  The S 0  4  -rich advective inputs most likely overwhelm 2-  consumption signatures in the hypolimnion in this season. The S 0  4  reduction  horizon deepens again in the spring as Mn and Fe oxides precipitate at and just below the reoxygenated interface. In addition to the reduction of sulphate, sulphide is also introduced to porewaters by the hydrolysis of proteins in organic detritus; fresh plankton has an organic-S content ranging from 1-2 wt.% (Jorgensen, 1977). Although dissolved trace metal profiles point to the precipitation of solid-phase sulphides as a sink for H2S, the respective organic and inorganic contributions to the overall accumulation of sulphur remain unclear (see section 5.2).  5.3.3 Iron and Manganese Diagenetic Fe and Mn enrichments are commonly observed at, or near the surface of oxic lake sediments.  Such accumulations normally result from the  upward diffusion of porewater Fe(II) and Mn(II) in anoxic sediments, followed by precipitation as their oxyhydroxides in surficial oxidized horizons (Carignan  157  and Tessier, 1988). In neutral lake waters, the oxidation of Fe is believed to proceed chemically, although its removal below the oxycline indicates that other oxidants in addition to O2, such as NO3" and Mn(II), may be involved. The extent of diagenetic iron oxyhydroxide enrichments in Balmer Lake can not be determined from the solid-phase data due to the large mining-derived signals of Fe and Mn in the top 10 cm of the sediment column (see section 5.2; Figs. 4.34F and 4.35A). However, the respective remobilization depths of porewater Fe and Mn during oxygenated periods provide an indication of the thickness of surficial oxide enrichments and hence help assess the sorptive capacity for trace metals. Dissolved Mn distributions in porewaters during the spring, summer and fall periods suggest that M n oxyhydroxides are remobilized at depths ranging from ~2 to 8 cm (Figs. 4.18-4.27). The shallower depths evident in the fall suggest the M n reduction horizon migrates upwards throughout the summer months, as alluded to in Section 5.3.2.  The addition of dissolved Fe to the summer  porewaters is thermodynamically consistent with the M n distribution; significant gradients are evident 2-3 cm below comparable increases for M n (Figs. 4.18 and 4.19).  Assuming that the Fe oxyhydroxide dissolution front migrates in  accordance with Mn, it can be proposed that the sorptive capacity of the surficial sediments for trace elements decreases through the summer, fall and winter, and increases again in the spring. Although dissolved iron distributions in Balmer Lake sediments exhibit considerable inter-site and inter-season (i.e., non-steady-state) variability, the available evidence indicates that authigenic sulphide precipitation presents a significant control on the distribution of Fe and other trace elements. During the summer period, concavity in the Fe profiles between ~8 and 20 cm depth at Stations 1 and 2 implies removal from pore solution (Figs. 4.18A and 4.19A). Since these zones lie immediately below the nitrate-zero boundary, sulphide  158  precipitation is implicated as a likely Fe sink.  Sulphide precipitation was  proposed to account for similarly shaped profiles in a Canadian Shield lake (Carignan, 1984; Carignan and Nriagu, 1985). The non-steady-state effects on the porewater sulphate distributions in Balmer Lake limit the use of the sulphate profiles in precisely defining dissolved sulphide production zones. In general, however, the sulphate distributions are thermodynamically consistent with the respective sulphate reduction horizons (Figs. 4.12C and F). The migration of the SO4 " reduction horizon towards the interface over the summer implies that authigenic sulphides precipitate at shallower sediment depths in the fall. The relative absence of Fe above 20 cm at station 1 in the fall (Fig. 5.5) can be attributed to two factors: (i) the reducible iron oxide inventory in the upper sediment column in spring is exhausted by late autumn; and (ii) sulphide precipitation removes iron from porewaters in the upper two decimetres as conditions become more reducing through the late summer. Deep sources of dissolved Fe can also be seen in profiles from stations 1 and 2 (Fig 5.5). Such increases at depth may represent the historical addition of dissolved iron to porewaters prior to the mining operations, when the natural dissolved sulphate concentration in the watershed was very low. Under such conditions, limited production of dissolved sulphide during diagenesis  allows high F e  2+  concentrations to persist in porewaters. Winter Fe profiles exhibit considerable inter-site variability (Fig. 5.6). The distribution at station 1 is fairly consistent with the summer and fall profiles at this site.  The high concentrations of dissolved Fe observed in the deep  porewaters at station 2 during the summer and fall (3-4 mg'L" ) are absent in 1  winter (Fig. 5.6). At stations 4 and 6, dissolved Fe levels generally remain < 500 (ig'L" . Whether such seasonal disparity at station 2 represents a sampling 1  artifact, or a combination of diffusion and consumption mechanisms, remains  159  Fig. 5.5.  Seasonal peeper profiles of dissolved Fe for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 160  Winter iron  0  500 1000 1500 2000 2500  Dissolved [Fe], ppb  Fig. 5.6. Winter peeper profiles of dissolved Fe for stations 1, 2, 4 and 6 in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 161  unclear. The rapid oxygenation of Fe(II) in the presence of oxygen does make the sampling of reducing porewaters more susceptible to potential error. The biological processes leading to sulphide mineral formation are generally characteristic of anoxic sediments where the bacterially-mediated oxidation of organic matter is accomplished via reduction of porewater sulphate and production of hydrogen sulphide (Berner, 1984). Under such conditions, metal ions remobilized by reductive dissolution of oxide minerals can react with H2S and other sulphur compounds to form a variety of sulphide minerals. Due to the abundance and ubiquitous nature of Fe, these authigenic precipitates are often dominated by Fe-sulphides. The major iron sulphide minerals found in lacustrine sediments are mackinawite (Feo.995-i.023S), greigite (Fe3S4), pyrite (FeS2) and amorphous-FeS (Morse et ah, 1987). As an assemblage, amorphous-FeS, mackinawite and greigite are generally termed acid volatile sulphides (AVS) because they are readily decomposed by acids. In the presence of HS" at concentrations above pyrite saturation, transformation of these metastable phases usually proceeds to pyrite, a more thermodynamically stable mineral (Berner 1984). However, pyrite can precipitate directly in environments where the more soluble monosulphides are undersaturated with respect to sulphide (Perry, 1993). The amount of pyrite that may form in sediments is controlled by the rates of supply of labile organic matter, reactive detrital iron minerals and dissolved sulphate (Berner, 1984). In most marine systems (those with oxygenated bottom waters), the amount and reactivity of organic matter buried in the sediment represent the primary control on pyrite formation because it is this organic supply which limits the rate of in situ sulphate reduction and thus hydrogen sulphide production (Berner, 1984; Calvert and Karlin, 1991).  Such sediments deposited in oxygenated  162  environments generally show a linear relationship between total or pyritic sulphur and organic carbon (Raiswell and Berner, 1985). In contrast to marine systems, pyrite formation in freshwater regimes is generally inhibited by low concentrations of dissolved sulphate (Berner, 1984). Therefore, much higher concentrations of pyrite typically form in organic-rich marine sediments; this has led to the use of C / S ratios to distinguish freshwater and marine sedimentary rocks (Davison et al., 1985). Experimental data suggest that limitations on pyrite formation are not important at sulphate concentrations greater than 5 mmol L" (in Berner, 1984, from Westrich, 1983). Concentrations of 1  ~3 mmol L" in Balmer Lake deep waters during the summer and fall imply that 1  authigenic sulphide formation may be somewhat limited by sulphate abundance. During the winter and spring periods, however, a plentiful supply of both organic matter and reducible sulphate (i.e., > 6 mmol'L" ) suggest that the amount 1  of reactive Fe-phases limits pyrite formation during these periods. From the above discussion, it can be proposed that extensive precipitation of authigenic sulphides in Balmer Lake sediments is likely a relatively recent phenomenon, reflecting the current unnatural abundance of SO4 " in the lake waters.  Recent sedimentary sulphide additions to S-polluted lakes in the  Sudbury region of southern Ontario have similarly been attributed to anthropogenic sulphate loadings (Carignan and Tessier, 1988).  Diagenetic  sulphur accumulations in excess of 3 wt.% (i.e., greater than those enrichments observed in Balmer Lake) have been reported in such lakes. Since the potential SO4" supply has exceeded the labile-Fe (i.e., non-silicate) supply in these lakes, S 2  fixation as inorganic sulphides appeares to be iron limited (Carignan and Tessier, 1985a; Nriagu and Soon, 1985). Reports of trace metal sulphide phases generally make no clear distinctions between precipitation as the pure metal sulphide or co-precipitation  163  with Fe sulphide. Experimental investigations have revealed that adsorption and coprecipitation of many trace elements with mackinawite (FeS) and pyrite (FeS2) are important processes in marine anoxic sediments (Huerta-Diaz and Morse, 1992; Arakaki and Morse, 1993; Morse and Arakaki, 1993). Furthermore, analyses of lacustrine iron sulphide particles collected from the anoxic bottom waters in a soft-water lake in the U.K. revealed C u and Z n concentrations in excess of 4,000 and 6,000 mg'L" , respectively; such levels were higher than those 1  observed in iron oxide precursors (Davison et al, 1992). Control by phases other than oxide and sulphide precipitates may be affecting the behaviour of Fe species in Balmer Lake, although this is difficult to constrain with the available data. Iron carbonate and phosphate minerals, for example, are potentially important authigenic components of lacustrine sediments.  Iron carbonate phases form in anoxic environments, where the  metabolic generation of inorganic C species leads to supersaturation, favouring the formation of siderite (FeC0 ) (Suess, 1979). Such solubility control has been 3  invoked to account for Fe distributions in several lacustrine studies (Emerson, 1976; Matisoff et al., 1980; Cook, 1984; Carignan and Nriagu, 1985). Although no definite conclusions can be drawn for Balmer Lake, the abundance of reducible sulphate suggest that dissolved Fe species in the anoxic porewaters would be scavenged by free sulphide. The seasonal porewater Mn distributions at stations 1 and 2 suggests that non-steady-state inputs of Mn-rich water are contributing to the temporal variability (Fig. 5.7). The summer and fall profiles are consistent with diagenetic processes; i.e., remobilization in reducing porewaters and precipitation in the oxic surficial horizons. However, the large winter maximum at ~15 cm depth at station 1 cannot represent a post-depositional signature.  The exhausted  inventory of reducible Mn-oxides in the surface sediments in late autumn  164  Fig. 5.7.  Seasonal peeper profiles of dissolved Mn for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 165  effectively excludes the dissolution of such phases as a potential source of the dissolved input. Furthermore, the depth of the dissolved maximum lies within the zone of postulated FeS precipitation, suggesting that labile Mn-oxides are not present at these depths. The above argument implies that an influx from the bottom waters is contributing to the observed winter profile (Fig. 5.7). Implicit in this argument is that at some point between the fall and winter sampling periods (Oct. to May), the sediments at station 1 were in contact with bottom waters hosting M n concentrations in excess of 1400 jxgU . The large mining-derived 1  sedimentary accumulations of Mn in the upper fades (Fig. 4.35) do suggest that tailings pond effluents host abundant manganese. Figure 5.8 illustrates the proposed temporal sequence. A n instantaneous input of dissolved Mn to the fall bottom waters (5.8B), followed by diffusion of Mn into the sediments would foster increased concentrations in the upper porewaters (Fig. 5.8C). By late winter, the strength of the bottom water M n source had decreased. However, due to the longer residence time of dissolved species in pore solution, evidence of the previous input persists (Fig. 5.8D). The high porewater concentrations support an efflux of Mn into the bottom waters during this period (Fig. 5.7).  A similar deep porewater maximum was not  evident for sulphate during winter sampling at stations 1 and 2 which may suggest that either: i) the consumption of sulphate in the interfacial horizons inhibits accumulation in the porewaters; or ii) the mining-related inputs of manganese and sulphate are temporally decoupled. Dissolved M n concentrations progressively decrease from the winter maximum through the spring, summer and fall (Fig. 5.7). The re-establishment of an interfacial Mn sink upon re-oxygenation of the water column in the spring suggests that the progressive consumption of the dissolved Mn near the  166  A [Mn]  B  c  [Mn]  [Mn]  D [Mn]  bottom waters sediments  Figure 5.8. Time series of progressive bottom water and porewater profiles (A-D) over the fall-winter transition in Balmer Lake: A) fall profile, B) instantaneous input of Mn to bottom waters, C) profile after some time of equilibration, and D) late winter profile. Depth and concentration axes are in arbitrary units.  sediment-water interface, in conjunction with diffusional transport, are responsible for the declining values. The minimal influence of mining-related bottom waters at station 4 make this site perhaps best suited to examine the winter M n behaviour (Fig. 4.25B). The Mn-redoxcline has clearly migrated above the sediment surface at this site, and temporary diffusive equilibrium has been established across the sedimentwater interface. The concave-upward profile between 8 and 25 cm above the lake floor is consistent with the oxidative precipitation and removal of upward diffusing Mn(JJ) species. Upward diffusive effluxes determined for Mn ranged from 54-700 ug'm" 2  'd  _1  (Appendix F). The low end represents fluxes calculated from the much  shallower gradients observed during the late-winter period. This is consistent with the hypothesis that once the oxidative sink for dissolved M n is removed  167  during winter anoxia, diffusional processes will cause a progressive decrease in the concentration gradient. Efflux estimates for the spring, summer, and fall exhibit reasonable consistency (Appendix F). Calculated effluxes were used to test the potential influence of such mechanisms on the bottom water composition. At station 4, for example, an excess of approximately 1 x 10 p.g has 5  been added to the lower 0.5 m of the winter water column (Fig. 4.11). Calculations using a constant maximum gradient suggest that diffusive processes alone would take on the order of 140 days to account for such an accumulation. Similar calculations for hypolimnic concentrations at station 1 (~2 x 10 ug excess 5  in the lower 0.5 m) and station 2 (~ 3.5 x 10 u.g excess in the lower 0.5 m), 5  demonstrate that time scales on the order of 280 and 500 days, respectively, 2 1  would be required. Since the maximum efflux of 700 u.g'm~ d~ was used in the above calculations, the estimated time periods most likely represent minimum values.  The calculations suggest that mining-related inputs may have  contributed to the bottom water enrichments at stations 1 and 2. Indeed, bottom water sulphate values are significantly higher at these two stations. The contrasting dissolved Mn profiles observed in the winter at station 6, indicate that Mn is being consumed as opposed to being remobilized in the upper sediments at this site (Fig. 4.26B). Sulphide precipitation as MnS (Suess, 1979) and sorption with mackinawite (FeS) (Heurta-Diaz and Morse, 1992; Arakaki and Morse, 1993) have been suggested to control Mn distributions in marine sediments. The relatively soluble nature of MnS (alabandite, pK = 0.40) suggests that the latter is more favourable in Balmer Lake sediments. Laboratory investigations have demonstrated that adsorption of Mn(U) with fine-grained mackinawite is more important than coprecipitation, and has been attributed to a combination of the high specific surface area of FeS, and a small partition coefficient that limits Mn incorporation into the FeS lattice (Arakaki and Morse, 168  1993). The tailings-rich deposits at station 6 perhaps favour the removal of Mn from porewaters via such adsorption mechanisms. The precipitation of Mn-carbonate phases has been suggested to govern the behaviour of M n in the reducing porewaters of several lake sediments (Matisoff et al., 1980; Carignan and Nriagu, 1985). Supersaturation with respect to carbonate minerals can be achieved in anoxic environments due to the pHbuffering effect by proteolytic formation of H S, other weak acids, and possibly 2  H -ion exchange (Suess, 1979). However, solubility considerations indicate that +  the porewaters at station 6 are likely undersaturated with respect to rhodochrosite.  5.3.4 Nickel  Peeper profiles suggest that two factors control the distribution of nickel in Balmer Lake porewaters: (1) release to solution from labile Ni-bearing phases at or near the sediment-water interface; and (2) the precipitation of authigenic Nibearing sulphide phases at depth. Each will be discussed in turn. Surface or near-surface maxima of dissolved N i can be observed throughout the year in almost all profiles collected from Balmer Lake sediments. Although the enrichments are generally small in relation to extent of removal at deeper depths, the peaks are well-constrained and appear to represent real signatures (Figs. 4.18-4.27). The magnitudes of these signals above the respective water column background concentrations average ~40 ppb and range from 30200 ppb. Concentration profiles suggest that N i is simultaneously diffusing from the interface to the sediments, and from the interface to the overlying water. The fact that these profiles exhibit reasonable seasonal consistency suggests that the spikes are representative of steady-state conditions.  169  Remobilization of trace metals at the sediment-water interface has been typically attributed to remineralization of organic matter and/or dissolution of Fe and Mn oxyhydroxides (Kh'nkhammer, 1980). In the case of Balmer Lake, the presence of dissolved N i maxima 2 to 5 cm above the predicted horizon of Mn(IV) reduction during the summer sampling session suggests that N i is released from the oxidation of organic matter, and not by dissolving oxides (Figs. 4.18D and 4.19D). Enrichments derived from organic matter oxidation would be expected to be observable at or near the sediment-water interface due to the higher rates of remineralization in the surface fades that stem largely from the greater reactivity of freshly deposited organic matter (Middleburg, 1989). In the fall period, the M n and Fe redox horizons are immediately proximal to the sediment-water interface, and as such, a biogenic origin cannot be uniquely isolated (Figs. 4.13-4.14). The cycling of organic matter has been invoked as a control on the distribution of N i because the metal has a nutrient-like distribution in the oceans (Bruland, 1980). The principal mechanisms of incorporation into particles are active algal uptake and indiscriminant complexation by high-affinity surface ligands. It has been proposed that some non-essential constituents function as chemical analogues to those that are essential, and hence, are sequestered and transported in a similar manner (Morel and Hudson, 1985). Although reports of N i concentrations in porewaters are rare, the existing information sheds important light on the early diagenetic behaviour of the element. Westerlund et al. (1986), for example, used benthic flux chambers to assess trace metal transport across the sediment-water interface in Swedish coastal deposits. They showed that Cu, N i , Zn and Cd, were released to the bottom waters in spite of the presence of abundant sulphide at shallow sediment depths. These trace metals were shown to be associated with phases from which  170  they were released via oxidation, rather than by reduction as was the case for Mn, Fe and Co. Similarly, Klinkhammer et al (1980) and Tsunogai et al. (1979) attributed post depositional migration of N i and C u to release from labile organics during early diagenesis in pelagic sediments. Interfacial enrichments on the order of ~ 10 ppb greater than immediately overlying water column values have been observed for some of the more biologically important elements (e.g., Zn; Morfett et al., 1988).  However,  oxidative additions in excess of 30 ppb for elements having less strong associations with organic matter (e.g., Ni), have not been previously reported. It is proposed that the high concentrations of N i in the water column of Balmer Lake (averaging >300 ppb) facilitates its sorption onto and/or incorporation with labile particulates in concentrations great enough to support interfacial releases of such large magnitude. Indeed, the particulate N i concentrations observed in Balmer Lake (Section 4.1.4.2) are comparable to the highest ever reported (see Nriagu et al, 1982). In addition to its affiliation with organic matter, nickel has been shown to have a high sorptive affinity for oxide surfaces (Vuceta and Morgan, 1978; Young and Harvey, 1992), and is commonly associated with diagenetic M n and Fe enrichments in lake sediments (Carignan and Nriagu, 1985; Cornwell, 1986). For these reasons, diagenetic accumulations of Fe and Mn in Balmer Lake sediments may be expected to liberate exchangeable N i species at their respective reduction horizons. Indeed, the large shallow spike of dissolved N i in the fall at station 1 (Fig. 4.20D) might reflect contributions from both organic and oxide sources. In addition, the slightly deeper N i enrichments observed in the spring (at ~5-7 cm), correspond more closely with the apparent remobilization depths of Fe and M n (Fig. 4.27D). Remobilized porewater N i maxima in excess of 8 ppb have been reported in coastal marine sediments (Shaw et al, 1990).  171  Rapid sulphide  complexation/precipitation of liberated N i may, however, prevent the formation of a dissolved N i excess at deeper depths in the summer. Indeed, the reducing environment in Balmer Lake sediments suggests that the zones of trace metal remobilization (due to Mn(IV) and Fe(IH) reduction) and sulphide precipitation almost overlap. The downward concavity below ~5 cm seen in all profiles indicates that the sediments of Balmer Lake provide a sink for dissolved N i throughout the year. Two observations support the precipitation of Ni-bearing sulphide phases as the most probable removal mechanism. First, the N i removal zones in general correspond with SO4 " reduction horizons.  Second, authigenic sulphide  precipitation has been reported for N i in several marine and lacustrine sediments (Morse and Arakaki, 1993; Green et al., 1989; Belzile and Morris, 1995), including reports from Canadian Shield lakes (Carignan and Nriagu, 1985). Peeper profiles of dissolved N i exhibit remarkable inter-site and interseason consistency (Fig. 5.9).  While sulphate distributions follow a seasonal  pattern consistent with the development of anoxia in winter (i.e., reduction depths are shallowest in late winter and deeper in spring and fall), N i removal profiles do not obey this sequence. In general, the predicted zones of N i removal occur deeper in the winter than in the fall and summer. Reasons for this are not clear; however, a decoupling of N i and sulphate during the winter months might be expected considering the variable rates of discharge and poorly constrained composition of mining inputs. Non-steady-state N i profiles associated with variable mining discharges like those seen for NFL; and Fe, are not easily +  discernible from the winter data. The predominantly uni-directional diffusive transport perhaps makes this element less susceptible to such influences. Solubility calculations indicated that the porewaters in Canadian Shield lakes studied by Carignan and Nriagu (1985) were supersaturated with respect to  172  Fig. 5.9.  Seasonal peeper profiles of dissolved Ni for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 173  millerite (NiS). Diffusional influxes of Ni, which ranged from 10-51 u.g'cm~ y" , 2  1  represented 59-161 % of the calculated N i accumulation rate. Similar influxes are 2  1  calculated for Balmer Lake sediments: values range from 2.6-14.2 p-g'cm" y" (Appendix F). On average, concentration gradients for nickel in Balmer Lake porewaters are steepest in the summer and fall; due to temperature-dependent diffusivities, this translates to the greatest diffusive influxes being observed in 2  1  the summer. Assuming an average influx of 6 u,g'cm" y" over the entire lake basin, a mean lake depth of 2.4 m, and an average water column concentration of 380 ug'L" , it would take on the order of 15 years for the sediments to consume 1  the whole-lake inventory of dissolved nickel. This corresponds to -1% sediment retention of the total dissolved nickel burden with respect to a lake residence time of -230 days, and suggests that sulphide precipitation has little effect on reducing the large burden of N i in Balmer Lake.  Furthermore, downward  diffusion alone cannot account for the large sedimentary N i accumulations. The accumulation rate, based on 0.3 wt.% N i , a dry bulk sediment density of 0.35 gem" , and a sedimentation rate of 0.3 em'y" , is estimated to be ~400 jj.g'cm^'y" 3  1  1  (Appendix G). Nevertheless, the gradients observed in the porewaters will support significant accumulation of N i in periods when tailings over-flow is absent.  5.3.5 Copper  The general harmony between the porewater and bottom water distributions of C u and N i , suggest that their behaviours are influenced by similar controls (Figs. 4.18-4.27). Unlike N i and Zn, however, dissolved C u profiles exhibit no evidence of any interfacial remobilization above the depth of sulphide precipitation. For all seasons, the behaviour of Cu appears to be largely  174  dictated by authigenic sulphide precipitation at shallow sediment depths during the spring, summer and fall, and in the anoxic bottom waters during the winter period. During the well-mixed periods of the spring, summer and fall, peeper profiles of dissolved Cu exhibit good inter-season and inter-site agreement (Figs. 4.18-4.22,4.27). The minimal deviations reflect slightly different removal depths, bottom water concentrations and diffusive influxes. Because there is no evidence of any regeneration from labile solid phases in the upper sediment horizons during these three seasons implies that either particulate Cu is transported to the sediments in refractory phases that are not subject to remobilization, or C u liberated to the shallow porewaters is rapidly precipitated. The biogeochemistry of Cu, and the conditions in Balmer Lake, suggest that the latter hypothesis is most tenable. Copper is considered to be biologically more important than N i with respect to algal metabolic requirements and is characterized by nutrient-like distributions in the oceans (Bruland, 1980; Sanders, 1983). Copper associated with labile organic matter appears to be released during early diagenesis in a wide range of sedimentary environments including lacustrine sediments (Morfett et al, 1988; McKee et al, 1989b; Reynolds and Hamilton-Taylor, 1992; Young and Harvey, 1992), coastal marine deposits (Elderfield, 1981; Westerlund et al, 1986; Shaw et al, 1994; Kerner and Geisler, 1995), pelagic sediments (Tsunogai et al, 1979), suboxic hemipelagic sediments (Klinkhammer, 1980) and marine minetailings deposits (Pedersen, 1985). Examinations of C u partitioning in various sedimentary phases have demonstrated that up to 80-90 % of the Cu burden may be organically bound (Elderfield, 1981; Young and Harvey/1992). The rapid scavenging of liberated species may account for the lack of dissolved C u additions to pore and bottom waters at the sediment-water  175  interface. The extremely insoluble nature of CuS (pK = 35.4; Stumm and Morgan, 1981) as compared to NiS (pK = 26.7), may favour the preferential removal of Cusulphide phases in the interfacial sediments. This proposal is consistent with the geochemistry of C u in the Balmer Lake water column.  The chemical  environment in the winter hypolimnion favours the removal of Cu presumably as sulphide precipitates; such a mechanism is not obvious for Ni. A fundamental criticism of the previous argument is the fact that the postulated Cu removal is occurring at depths shallower than those predicted for sulphate reduction. However, sulphide precipitation has been implicated for trace element removal in sub-oxic sediments. The removal of porewater C d at a suboxic front in abyssal Atlantic sediments, for example, was attributed to the precipitation of authigenic CdS (Rosenthal et al., 1995). They suggested that the diffusion of free-sulphide from underlying anoxic sediments was sufficient to support sulphide formation. In addition, Sorenson and Jorgenson (1987) found that sulphate reduction can take place within the microenvironment of organic aggregates in the oxic zone. It is possible that the organic-rich sediments in Balmer Lake promote such development of heterogeneous redox conditions in the "oxic" horizons. The tendency of a metal to form ligand complexes is usually explained in terms of ligand field crystal-stabilization theory (Stumm and Morgan, 1981). This led to the Irving-Williams order for complex stability of plus II cations, which increases in the order: M n  2 +  < Fe  2+  < Co  2+  < Zn  2 +  < Ni  2 +  < Cu  2 +  . The  sequence presented here reflects increases in the capability of the cation to take up electrons (i.e., increasing ionization potential of the metal). Complexation, steric hindrances, entropy effects and site specificity can distort this hierarchy (Benjamin and Leckie, 1981). Based on analyses of porewaters collected in Anderson Lake, Manitoba, Pedersen et al. (1993) suggested that the high sorptive  176  affinity of C u for dissolved organic ligands, which would tend to keep C u in solution, could account for Cu being removed at deeper depths than other metals such as Zn and Cd. This suggestion appears not to apply to the porewater C u and N i distributions in Balmer Lake, however, as the depth of C u removal is always shallower than that for Ni. The anoxic sediments in Balmer Lake clearly provide a permanent sink for dissolved Cu throughout the year. Comparisons of seasonal profiles for stations 1 and 2, for example, suggest active sulphide precipitation occurs at depths between 2-10 cm from spring through fall (Fig. 5.10). Unlike dissolved N i , the seasonal migration in the depth of Cu removal is consistent with that of sulphate; precipitation zones migrate upwards over the summer-fall-winter transition and descend again in the spring. Sulphide precipitation has been frequently proposed to limit C u concentrations in marine and lacustrine porewaters (Elderfield, 1981; Pedersen, 1985; Gaillard et al, 1986; Pedersen, 1988; Pedersen et al, 1990; Hamilton-Taylor and Davison, 1994) , including reports for shield lakes (Carignan and Nriagu, 1985; Pedersen et al, 1993). Carignan and Nriagu (1985), for example, observed dramatic depletions of dissolved C u at depths consistent with sulphide production in one acidic (pH ~ 4.5) and one slightly alkaline Canadian Shield lake (pH ~ 7.5); solubility calculations indicated that the porewaters were supersaturated with respect to covellite (CuS). Similarly, Pedersen and others (1993) attributed the precipitous decline of C u in the shallow porewaters of contaminated Anderson Lake, Manitoba, to the precipitation of authigenic discrete or mixed metal sulphide phases.  Indeed, framboidal pyrite was  observed by these authors to be ubiquitous in the surface sediments. Diffusional influxes clearly support significant accumulations of C u in some lacustrine sediments. In the Shield lakes studied by Carignan and Nriagu  177  B  Station 1 40  Station 2  E o of o ca t  0) CO  $  i +-*  c  e d) CO  E o M—  CD O  c  CO  ^—•  CO  b Spring  -60  0  50  1 0 0 150 2 0 0 2 5 0 3 0 0  Dissolved [Cu], ugL"  Fig. 5.10.  1  -60  0  50  100 150 200 250 300  Dissolved [Cu], fig L"  1  Seasonal peeper profiles of dissolved Cu for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 178  (1985), for example, diffusive influx estimates ranged from 0.5-5.1 ug'cm^'y" and 1  represented 1-52 % of the estimated C u accumulation rate. In Balmer Lake, diffusive C u influxes remained reasonably constant over the four seasons, with values ranging from 2.3-6.7 M-g'cm" y" (Appendix F). In general, no obvious spatial or temporal trends are evident with respect to the inter-site and interseason flux estimates. The rate of Cu removal by sediment influxes appears to have little effect on the total lake burden of dissolved C u in Balmer Lake. Estimates assuming an average influx of 4 ug'cm" y", a mean lake depth of 2.4 m, and an average water column concentration of 200 M-g'L' , suggest that -5% of the 1  total dissolved Cu is fixed within the lake residence time of 230 days. Downward diffusive fluxes of C u in Balmer Lake do not appear to represent the principal depositional mechanism responsible for the large sediment concentrations. As suggested for Ni, the calculated rate of diffusion of dissolved C u species from the bottom waters into the sediments is not sufficient to account for the bands of Cu-rich sediments (~1 wt.%) at station 1 (Fig. 4.35). The accumulation rate implied by the enriched horizons is estimated to be on the order of 1000 ug'cm^'y" (Appendix G). However, since the last ephemeral 1  episode of accidental tailings overflow (~1970), diffusional transport has most likely played a significant role in C u accumulation. Indeed, the steep negative concentration gradients are comparable to the magnitudes observed in the cited previous studies in which the importance of such mechanisms has been stressed.  5.3.6 Zinc  The bottom water and porewater distributions of dissolved Z n in Balmer Lake appear to be controlled by two principal mechanisms: 1) intense remobilization from labile, Zn-rich sedimentary phases in the first 5 cm below  179  the interface; and 2) precipitation of Zn-bearing sulphide minerals at greater depths. Each will be described in turn. Peeper profiles of Zn are characterized by shallow pronounced porewater maxima during all seasons (Figs. 4.18-4.27). Steep concentration gradients either side of the spike promote bi-diffusional transport away from the interfacial horizons. The concentrations of the surficial peaks average 75 ug L" greater than 1  concentrations in the overlying bottom waters (Fig. 5.11). Typically, the largest interfacial spikes and gradients were observed during the fall period. Reasons for this are not fully clear, although the degree of enrichment will be related to the rate of supply and the rate of removal of dissolved Z n species. Specifically, such processes will be influenced by the supply of organic matter, rates of organic matter remineralization and rates of authigenic sulphide formation. Diffusive effluxes, which ranged from 11-330 ug'm^'d" , were computed in 1  order to assess the potential contribution of remobilization on the Zn enrichment in the winter hypolimnion (Appendix F). The following discussion ignores potentially significant influences from advecting mining-flows.  From water  column profiles, bottom water enrichments in the lower metre of the water column at stations 1, 2 and 4, were estimated to be ~ 5 x 10 , 3.5 x 10 , and 1.5 x 4  4  10 ug, respectively. Assuming a constant efflux of -300 ug'm^'d" as seen in the 4  1  fall, upward diffusion of Zn species at these stations would take -170,120 and 50 days, respectively, to produce the concentrations observed in the bottom waters. These values represent minimum time estimates, as effluxes are expected to decrease as gradients lessen during ice-cover. They do, however, illustrate that benthic effluxes can have a significant impact on the metal content of the water column during the seasonally anoxic period. The remobilization of Zn from the oxidation of organic matter is considered to be the most important source of labile Zn species in Balmer Lake  180  Fig. 5.11.  Seasonal peeper profiles of dissolved Zn for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 181  sediments.  Four strands of evidence are offered to support this. First, zinc  appears to be generated at shallower horizons than the respective remobilization depths of Fe and Mn during the summer period. The putative organic source is more difficult to distinguish in the more reducing fall sediments. Second, of the elements measured, zinc exhibits the best correlation with POM in water column particulates. Third, a pronounced surficial dissolved Zn spike was evident at station 1 during the winter at a time when Fe and Mn oxide sources should have been exhausted. And fourth, reports of strong Zn-organic associations abound in the literature.  With respect to the latter point, compositional analyses of  phytoplankton demonstrate that Zn is typically the most abundant trace metal in both marine and lacustrine algae; extensions of Redfield ratios to include micronutrients have revealed stoichiometrics of Cn3N15P1Zno.03-o.06 (Sigg, 1985; Reynolds and Hamilton-Taylor, 1992). However, porewater profiles that provide direct evidence of surficial enrichments derived from remobilized organic matter are scarce in the literature.  Data from seasonally anoxic Esthwaite Water  demonstrated that elevated dissolved Zn concentrations at the sediment surface were attributable to a remobilized source from decaying algal material (Morfett et al., 1988). Transient interfacial Zn spikes coincided with large accumulations of organic material at the sediment surface, and were characterized by constant Zn:Cu ratios in a stoichiometry consistent with that of phytoplankton (Sigg, 1985; Reynolds and Hamilton-Taylor, 1992). Similarly, effluxes of Zn in Swedish coastal sediments were attributed to releases during the breakdown of labile organic matter at or near the sediment-water interface (Westerlund et al., 1986). Comparisons of the magnitudes of near-interface dissolved Z n contents between Balmer Lake and the studies cited above reveal considerable disparities. The greatest enrichments reported by Morfett and others (1988) and Westerlund (1986) were respectively ~ 12 and 20 ug'L" above background water column 1  182  concentrations; those observed in this study ranged from 15-200 ug'L" . It must 1  be pointed out, however, that Balmer Lake is unique in that it hosts dissolved Zn concentrations over an order of magnitude higher than either Esthwaite Water or Swedish coastal waters. More importantly, particulate Zn concentrations are among the highest recorded anywhere (Table 4.6). Therefore, one can argue that potentially large burdens of labile, Zn-bearing particulates are deposited in Balmer Lake; the remobilization of such phases may facilitate the formation of the anomolously large Zn spikes in the surficial horizons. Due to the compressed redox depth scales in Balmer Lake, dissolved Zn contributed by near-interface reductive dissolution of Fe and M n oxides is difficult to distinguish from Z n released during organic matter remineralization. A spatial decoupling of the two sources is evident only during the summer. Previous studies which have assessed the partitioning of zinc between the water column and the sediments in lakes suggest that the binding of Zn by Fe and Mn oxides may be important in the seasonally oxic horizons in Balmer Lake (McKee et al, 1989a; Tessier et al, 1989; Williams, 1992; Young and Harvey, 1992). A n examination of eight Canadian Shield lakes by Tessier et al (1989) demonstrated a strong pH-dependence with respect to the distribution coefficients of Z n between particulates and solution phases (i.e., 1Q increased with pH).  Their results agree well with theoretical and experimental  complexation models of cation sorption to oxide surfaces (Benjamin and Leckie, 1981). For the lakes assessed (pH values ranged from 4 to 8.4), the p H interval 5.5 to 6.5 defined a region of maximum potential change in % Z n sorbed to Fe oxyhydroxides; the results suggested that a large portion of the potentially available or mobile Zn present should be associated with the surficial sediments above p H 7, and be present in the water column below p H 5. A similar study of oxic  lake  sediments  from  seven  183  other  shield  lakes  examined  the  sediment/porewater partitioning (i.e., Ka) of Zn together with the speciation of the bound fraction (Young and Harvey, 1992). Enhanced sorption between p H 5.0 and 6.5 was also observed in this work. Although Zn concentrations in Balmer Lake are complicated by mining inputs, the p H range of 7-7.4 (i.e., the upper end of the sorption edge) suggests that a significant portion of the sedimentary Zn inventory should be associated with Fe and/or M n oxides in oxic sediment horizons. However, the role of these phases in the diagenetic cycling of Zn in the lake sediments is not clearly reflected in the porewater zinc profiles. Below the horizons of near-interface remobilization, dissolved zinc species are rapidly consumed from Balmer Lake porewaters presumably by the precipitation of authigenic sulphides. The precipitation of Zn-sulphides is a widely recognized aspect of Zn behaviour in lacustrine systems (Pedersen, 1983; Carignan and Tessier, 1985b; Pedersen, 1988; Tessier et al, 1989; Davison et al, 1992; Williams, 1992; Pedersen et al, 1993), and has been the subject of a fairly recent review (Hamilton-Taylor and Davison, 1994). For all seasons and sites, Zn and N i are removed at very similar depths, which may reflect their relatively comparable solubility products with respect to sulphide. As for N i , sulphide precipitation in the hypolimnion does not appear to present an important removal mechanism during winter. During the summer, winter and spring periods, Zn concentrations in the interstitial waters are at least a factor of two lower than in the overlying bottom waters (Fig. 5.11), suggesting there is a net diffusive transport into the sediments. In the fall, however, diffusive equilibria between the sediments and water column are evident at stations 1 and 2 (Fig. 5.11). Diffusive influxes of dissolved Zn below the near-interface peaks exhibited little inter-season variability with seasonal averages ranging from 2.2-3.9 ug'cm^y" . Except for the fall period, 1  184  concentration gradients associated with downward diffusion were greater than those corresponding to effluxes (Appendix F). Tessier et al. (1989) examined a large number of oligotrophic, shield lakes of varying p H and degrees of Z n contamination. In the more polluted lakes (pH<6), downward influxes driven by pronounced gradients (0.39-1.7 x 10" gem" ) suggested dissolved Zn species were 8  4  being efficiently trapped as sulphide phases at depth; these gradients agree well with those seen in Balmer Lake deposits, which range from 2.2-5.3 x 10~ gem" 8  4  (Appendix F). In lakes of pH>6, Tessier et al. (1989) observed that the vertical concentration gradients were generally much lower (0.65-1.9 x 10" gem" ). The 9  4  greater Zn inputs to the acidic lakes in combination with more vigorous sulphide precipitation were suggested to contribute to the larger gradients. In a related study, porewater Zn profiles were obtained by Carignan and Tessier (1985) in order to assess the relative importance of diffusion mechanisms in two acidic lakes. Loci of Zn removal at depths of 2 cm coincided with the presence of measurable porewater sulphide; in both sediments, the porewaters were slightly supersaturated with respect to sphalerite. Carignan and Tessier estimated that the diffusive influxes in the lakes (1.0-1.7 ug'cm'^y" ) accounted for 1  76 and 52 % of the respective sedimentary Zn accumulations. The importance of diffusive influxes to the overall Zn accumulation in Balmer Lake deposits cannot be established so rigorously given the significant seasonality in the porewater distributions. Nevertheless, the data collectively point to an important conclusion: early diagenetic reactions in Balmer Lake sediments render the deposits below 10 cm depth a significant and permanent 2  1  sink for the metal. Calculations using an average influx of 3 M-gcm" y" over the entire lake benthic area suggest that 13% of the total dissolved Z n inventory is retained with respect to the lake residence time of -230 days.  185  5.4 Factors Controlling the Diagenetic Behaviour of As  The ubiquity of Fe and M n oxyhydroxides, and their potential role in regulating trace element concentrations in natural waters, has stimulated numerous investigations on As behaviour in the presence of these substrates. The importance of Fe-bearing oxides has been repeatedly inferred to exert the principal control on the distribution and behaviour of As in marine and lacustrine systems (Crecelius, 1975; Edenborn et al., 1986; Agett and Kriegman, 1988; Belzile and Tessier, 1990; De Vitre et al., 1991; Widerlund and Ingri, 1995) while Mn has received relatively little attention (Takamatsu et al., 1985; Anderson and Bruland, 1991). The apparent strong associations between As and M n observed in Balmer Lake sediments imply that under certain conditions, manganese may also play a significant role in As geochemistry.  The section  below assesses the importance of M n and Fe with respect to the diagenetic behaviour and mobility of As in sediments, and is presented within the contexts of experimental and theoretical considerations. The argument that is developed will be then applied to the seasonal geochemistry of As in Balmer Lake.  5.4.1 Experimental and Field Observations  Theoretical and experimental studies were reviewed in order to determine the geochemical influences of Fe and Mn on the diagenetic behaviour of As. A number of questions receive particular attention in the following paragraphs: 1) how do the sorptive properties of Fe and Mn oxide phases compare? 2) what are the As species predicted to be associated with these phases? and 3) what is the nature of the adsorption mechanism?  186  As discussed in sections 5.1.5.1 and 5.1.5.2, natural Fe and M n oxides represent a diverse assemblage of phases that encompass a broad spectrum of sorptive characteristics. Due to variability in the chemical environment during their formation and their ability to incorporate foreign ions into their structures, significant natural heterogeneity can develop in oxides with respect to reactive surface areas, types of binding sites, site densities and steric effects.  A  comprehensive body of literature has been devoted to their sorptive properties. Although most experiments commonly use relatively pure Fe and M n oxide phases, which may exhibit considerable compositional and morphological differences compared to natural particles, useful comparisons can be drawn. Poorly ordered M n and Fe oxides are typically characterized by loosely hydrated structures which are permeable to ions (Davison, 1993).  The  incorporation or adsorption of minor amounts of foreign ions in the formation of these phases appear to restrict crystal development, thus contributing to more random growth (Taylor, 1987). The permeability means that sorption reactions are not restricted to external sites as is the case for more crystalline solids. Thus, the reactive surface areas for the mostly amorphous phases can be large. BET analyses of synthetic M n 0 and Fe(OH) suggest that they share similar specific 2  3  2  1  surface areas. Estimates for the former range from 230-320 m g" (Vuceta and Morgan, 1978), while those for amorphous Fe-oxide range from 182-600 ir^'g"  1  (Davis and Leckie, 1978b; Davis and Leckie, 1978a; Benjamin and Leckie, 1982). Similarly, good agreement has been observed between M n and Fe solid phases with respect to experimentally derived site densities and chemical free energies of adsorption for trace elements (Vuceta and Morgan, 1978). More definitive contrasts between Fe and M n oxide phases are evident upon comparison of the nature of their respective surface charges. The surfaces  187  of many oxides become relatively more negatively charged with increasing p H (and vice versa) due to variations in the adsorption of potential-determining ions (H 0 3  +  and OH") (Stumm and Morgan, 1981). The p H at which the surface  assumes a net zero surface charge (the point of zero charge, or pzc) can be used estimate whether a surface is likely to be positively or negatively charged at a given p H (Kinniburgh, 1981). The pzc for goethite for example is 7.5-8.6 while that for birnessite is 1.5-2 (Kinniburgh, 1981); goethite (a-FeOOH) and birnessite (5-Mn02) represent two of the more abundant oxides of Fe and M n in natural systems.  Therefore, at neutral p H , M n oxide is predicted to be strongly  negatively charged and thus will repel anions. This argument has been proposed by several groups to support stronger associations of Fe oxides with arsenate (H As0 ") at natural p H levels (De Vitre et al, 1991). 2  4  To constrain further the diagenetic behaviour of As in sediments, it is necessary to define the chemical species involved and compare their behaviour. Thermodynamic arguments suggest that inorganic As in oxidized aquatic systems should be present as arsenate [As(V)] (Ferguson and Gavis, 1972); arsenic acid (H3ASO3, p K  al  = 2.2, p K  a2  = 6.9) is predicted to be largely dissociated  at neutral pH's. Conversely, data in the literature show that arsenite [As(Hl)] typically predominates in anoxic interstitial waters (Edenborn et al., 1986; Peterson and Carpenter, 1986; Agett and Kriegman, 1988); arsenous acid (H As0 , p K 3  4  a l  = 9.2) is expected to be predominantly neutrally charged in  natural waters. For simplicity, organo-arsenicals will be ignored in the following discussion. Both arsenite and arsenate are removed from solution in the presence of amorphous Fe-oxides, although As(V) is more readily sorbed than As(IH) (Pierce and Moore, 1982; Bowell, 1994). The contrast has been attributed to charge  188  differences between the two complexes. Although accounts of solid-phase As speciation are rare, it is generally assumed that As(V) predominates in oxic sediments (Belzile and Tessier, 1990). Dissolution experiments using diagenetic Fe and As samples collected from 12 Canadian lakes, for example, demonstrated that only arsenate was present (De Vitre et al, 1991). From the above discussion, it is apparent that the scavenging of As (III) by oxide surfaces involves an oxidation step.  For example, As(III) removal from the porewaters in oxic  horizons may involve oxidation to As(V) prior to adsorption. Alternatively, the oxidation of As(III) to As(V) may occur during or after its adsorption (Gulens et al, 1979). Indeed, the sorption of As has been shown to occur simultaneously with the oxidation of As(III) to As(V) in several Saskatchewan lake sediments (Oscarson et al, 1980). Molecular O2 and Fe and Mn oxyhydroxides are the most likely oxidants of arsenite in sedimentary porewaters. However, oxidation by molecular O2 is kinetically slow, and thus is unlikely to be important (Cherry et al, 1979). In contrast, Oscarson et al (1980) observed that synthetic, poorly crystalline birnessite (5-Mn02) was very effective in oxidizing As(III). Microbial inhibition via the addition of HgCl2 did not significantly retard the reaction, indicating that such oxidation was largely an abiotic process.  Despite the thermodynamic  favourability, a redox reaction between As(III) and Fe(III) oxide was not observed within three days in a similar experiment (Oscarson et al, 1981a); the results were attributed to slow kinetics.  However, in another study Fe  oxyhydroxides (both natural and synthetic) were found to oxidize As quite rapidly; typically, 50-75% of spiked As(III) was oxidized with two days, irrespective of p H (De Vitre et al, 1991). The apparent discrepancy between the two experiments may be explained by the much higher Fe:As ratios (3,400 vs. 10)  189  in the latter study.  Thus, it appears that both Fe(IIJ) and Mn(IV) serve as  effective oxidants of As(IU) given suitable conditions. Experiments have shown that both coulombic attraction and specific adsorption are important in the uptake of arsenate by Fe-oxyhydroxides. In this context, specific adsorption refers to uptake that cannot be accounted for solely by electrostatic interactions, regardless of the sign of the surface charge (Stumm and Morgan, 1981). "Specific adsorption" which is a term equivalent to "ligand exchange", thus reflects the specificity in bonding of different ions due to their charge, size and polarizability (Taylor, 1987). Several strands of evidence have been offered to support specific associations of arsenate with Fe-oxyhydroxides. First, the specific adsorption of anions induces a greater negative charge on oxide surfaces which results in a shift in the point of zero charge (pH ) to a lower p H value. The amount of shift pzc  depends on the particular ion sorbed, its concentration, and the nature of the solid surface (Stumm and Morgan, 1981). Several groups have demonstrated pH  p z c  shifts of several p H units for Fe-oxyhydroxide surfaces with sorbed  arsenate, indicating specific interactions (Pierce and Moore, 1982; Fuller et al, 1993; Bowell, 1994; Hsia et al, 1994). Second, adsorption due to electrostatic processes is usually very rapid (i.e., seconds). Equilibration periods on the order of hours observed for As-Fe systems indicate formation of chemical bonds between the As species and the adsorbent (Oscarson et al, 1980; Pierce and Moore, 1982; Fuller et al, 1993). Additional evidence for direct coordination of arsenate has been derived from various other chemical and physical techniques including energy dispersive X-ray analysis (EDAX) and Fourier transform infrared (FTIR) methods (Hsia et al, 1994). Laboratory demonstrations which have established the importance of Fe in As cycling are consistent with other studies of marine and lacustrine  190  sediments. A very convincing data set of six Canadian lakes (pH between 4.0 and 8.4) demonstrated the importance of Fe on the behaviour of As (Belzile and Tessier, 1990). Tightly correlated porewater profiles for all lakes indicated that As was mobilized during the reduction and dissolution of Fe-oxyhydroxides. Moreover, binding constants derived from the field data compared well with those obtained in the laboratory for the adsorption of arsenate on amorphous Feoxides. Similar conclusions were drawn for comparable associations of dissolved Fe and As in a Canadian Shield lake (De Vitre et al., 1991). In general, field studies which have alluded to an As-dependence on M n have not differentiated between Mn and Fe-oxide control. Lake Ohakuri in New Zealand, for example, is a seasonally anoxic lake that receives unusually high As inputs from geothermal sources. The accumulation of dissolved Fe, M n and predominantly As(III) in the hypolimnion over the course of anoxia was attributed to sediment release via the reductive dissolution of Fe(III)- and perhaps Mn(IV) oxides (Agett and O'Brien, 1985; Agett and Kriegman, 1988). Indeed, the large interfacial gradients supported sediment effluxes for all three elements. Relatively fewer laboratory studies of arsenate sorption onto Mn-oxide phases exist in the literature.  Oscarson et al. (1981) demonstrated that  synthetically prepared birnessite could efficiently oxidize arsenite to arsenate; the reaction was accompanied by increases in the concentrations of remobilized Mn(II) and K from the oxide lattice. However, while a significant portion of +  As(V) was removed from solution by Fe(III) oxide in an identical experiment, very little of the added As was sorbed by Mn02. Since the pzc of birnessite is close to p H 2 (Oscarson et al, 1981b), it was assumed that the high negative charge density at experimental p H (pH=6.9-7.7) inhibited the sorption of the anionic arsenate species (largely H2ASO4" and H A S O 4 ").  191  In a similar experiment with synthesized hydrous MnG^, Takamatsu et al. (1985) demonstrated that arsenate was not appreciably adsorbed in the p H range 6 to 8. However, in the presence of large concentrations (20 mg'L" ) of a divalent 1  cation, a substantial increase in the amount of adsorbed arsenate was observed; the percent adsorbed increased from <20% to 100 % at p H 6 with each of the added trace elements (Mn , Sr , Ba 2+  2+  2+  and Ni ). The results were explained in 2+  terms of a shift in the surface charge of the Mn-oxide surface in the presence of abundant cations. Divalent cations such as M n  2 +  and N i  2 +  readily sorb to Mn-  oxides, and in the process, displace protons to solution from the oxide surface (Stumm and Morgan, 1981). In this study, the amount of metal adsorbed relative to the extent of proton release was approximately 1:1 on a molar basis. Takamatsu and others (1985) suggested that since the ratio of charge equivalents released to charge equivalents adsorbed was <1, cation adsorption would increase the pzc of the oxide surface, and hence, enhance arsenate sorption. In another experiment, a slurry of oxidized As-bearing surface sediment and seawater was incubated with added plankton material to simulate the sequential depletion of oxidants that occurs during early diagenesis (Edenborn et al., 1986). The sequential release of reduced M n and Fe species, as well as the timing of SO4 " reduction, were thermodynamically consistent with measured 2  redox potentials. The release of As(IU) to the porewaters occurred concomitantly with the release of Fe(II), suggesting strong Fe-As associations. It should also be noted that the As release came after the dissolution of almost all of the reducible Mn. The thermodynamics of reductive dissolution of Fe and Mn oxyhydroxide phases has further implications with respect to their relative influences on the diagenetic chemistry of As.  Theoretical examinations of organic matter  diagenesis predict that Mn(IV) should be reduced at a higher redox potential  192  than Fe(III) (Froelich et ah, 1979). This implies that in reducing sediments, the addition of manganous ions to porewaters should occur at shallower sediment depths than corresponding ferrous ion inputs. In mildly reducing sediments where thermodynamic gradients are shallow (i.e., the redox horizons for the various interstitial metabolites extend over many centimetres), such spatial decoupling of Fe and M n reduction can be observed (Edenborn, et ah, 1986). Therefore, even assuming As is associated with Mn oxyhydroxides, it is possible that any release of As to the porewaters upon Mn dissolution could be masked by re-adsorption onto Fe oxyhydroxide phases. Hence, the final expression of As release to the porewater would be more closely associated with increases of dissolved Fe . Such an argument was put forth by Peterson and Carpenter (1986) to account for strong associations observed between M n and As solid phases in surface sediments, and the concomitant release of dissolved Fe and As in deeper horizons. Because intensely reducing sediments are less likely to foster a F e / M n redox decoupling, the mechanism of Fe control described above may be less important in such regimes. In summary, contrasts in the surface charge characteristics of Fe and M n oxide phases influence the sorption of charged species. The heterogeneity of these particulates makes comparisons of other parameters (e.g., specific surface areas, site densities, etc.) more difficult to assess. In addition, experimental data demonstrate a pH-dependence of arsenate sorption for which both coulombic and specific mechanisms are involved; however, the exact nature and variability of the chemi-sorptive processes are not completely understood. Speciation studies suggest that arsenate is the dominant species in oxic sediment horizons and that arsenite predominates in reducing interstitial waters; the removal of As(III) from solution appears to involve its oxidation by Mn(IV) and probably Fe(III) oxides. Finally, limited experimental data demonstrate weak associations  193  of As with M n oxides; significant sorption has only been reported for high cationic-strength solutions in which the charge properties of the oxide surface have been modified. Conversely, similar studies offer conclusive evidence that Fe oxide phases exert significant control on the behaviour of As.  5.4.2 The Behaviour of As in Balmer Lake Sediments  Seasonal porewater profiles suggest that two principal controls are governing As distributions in Balmer Lake sediments.  First, intense  remobilization at shallow sediment depths indicate a redox-controlled release from labile, As-bearing particulates; and second, lower concentrations in the deeper horizons suggest the consumption of As as authigenic sulphide phases. Each of these processes will be described in turn. During the well-oxygenated periods of the spring, summer and fall, porewater As profiles are characterized by steep gradients that begin just below the sediment surface (Figs. 4.18-4.22; 4.27).  The depths of remobilization  typically overlap with corresponding dissolved Mn production horizons. This suggests that a labile portion of As is associated with Mn-oxyhydroxide phases in the upper horizons, both of which follow a redox-pattern of dissolution in the suboxic zone, upward diffusion, and precipitation near the sediment-water interface. Where dissolved Fe gradients are clearly defined, corresponding As slopes are shallower (Fig. 5.12), suggesting that the dissolution Fe solid phases is not responsible for the porewater additions of dissolved As at sediment depths above than the Fe(III) redox boundary. The chemically analogous behaviours of phosphate and arsenate, and hence the potential for biological incorporation, might suggest that the remineralization of organic matter may contribute to the dissolved As enrichments; however, the decoupling between the oxidation of  194  Summer Station 2 D i s s o l v e d [ M n ] , ug L " 200  0  400  600  800  1000  2000  3000  1  1000  4000  D i s s o l v e d [ F e ] a n d [ A s ] , ug L "  1  Fig. 5.12. Summer peeper profiles of dissolved Fe, M n and A s for station 2 in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 195  organic material and the dissolution of oxide phases as seen for Zn during the summer period is not clearly evident for As. A linear regression was performed to assess further the dependence of As on the redox behaviour of Mn in Balmer Lake porewaters. Molar concentrations were compared at depths extending from the interfacial horizons to depths of ~20 cm in order to encompass the remobilization depth range (Fig. 5.13). Coefficients of determination (r ) can be interpreted as the proportion of the variation in Y that is "accounted for" by linear regression of Y on X. In other words, an r value of 0.9 means that 90% of the variation in porewater As 2  concentrations can be accounted for by variations in Mn. The r coefficients for the spring, summer and fall average 0.85 (0.70-0.96), indicating that the diagenetic cycling of M n is significantly affecting the behaviour of As. Corresponding values for Fe and As over these periods are slightly lower (mean = 0.78(0.42-0.96)). The fact that the A s / M n slopes in Fig. 5.13 are not linear demonstrates that mechanisms other than the dissolution of Mn-oxide phases are involved; for each profile, uniform A s / M n ratios of <1 are evident to ~3-4 cm and then generally increase with depth to ~2 in the deeper horizons. The preferential removal of Mn from pore solution or the addition of As (e.g., via the dissolution of Fe-oxide phases) may account for the observed increases. It should be stressed that the weaker correlations between As and Fe do not imply that Mn is exerting greater control on the distribution of As in Balmer Lake porewaters. Instead, the dearth of ferrous-Fe in the shallow porewaters may suggest that Fe is rapidly consumed as authigenic sulphide phases. Indeed, the zones of Fe(III) and SO4 " reduction are predicted almost to overlap. It is 2  therefore possible that As is remobilized upon the reductive dissolution of Fe(IH) oxides, but Fe consumption from porewaters hinders conclusive interpretations.  196  Summer T—|—i—I  60 F-  |  40  I  |  I  I  I  |  •  St. 2, ^=0.87  •  SI. 2, ^=0.96  A  81.1.^-0.92  O  St. 1,^=0.90  I  I  I  |  •  25  I—r  •  1 1  i  1 1 1 1  i  "5 E oo  A  'd '  i • ' i 1  1  •  •i  0 ^  12  14  1 1 1 1  .. n  "  •  =t 10  O  5  16  *  D  St. 1,^=0.71  D  • 5  18  •  St. 5, r'-o.ea •  •1-1-*-1. .1. i  10  1 1 1 1  D  15f-  < 8  ' i  1 1 1  20 f-  A  «F 20  Fall  B  • • • •  1  10  1  • • • •  15  1  • • • •  20  1  •  • • • •  25  30  35  -1  [Mn], umol L"  [Mn], umol L"  1  Fall ^  V  _J  40  1 1 1 1 1 St. 2, ^=0.94  1 1  •  30  St. 2, ^=0.70  1 1 1  CO  •  • O • O  •  10  %  •S»  I  * 1 1 1  2  1  CP .. i . . . i . . . i . . . i . . . i . . .  4  6  8  10  12  [Mn], umol L'  14  :  16  1  Spring  Winter  D  i * * * i *  1  1  i ' * * i * '* i *  o  1  1  i  1  1  -I—I—I—I—|—I—I—I—I—|—'—I—I—I—|—I—I  '  30  10  o E  3.  ^<  •  • 20  •-  Q  o  E  1  <o°o  @%  O  •  St. 1.1^=0.85  •  St. 1.1^=0.81  I  I—|—I—'—i—r-  1  1  20  6  In  <  0  St. 2, ^=-0.07  •  St. 1,^=0.50 St. 4, ^=0.01  1 •  yf 10 <  •  •  i  •  I  • •—i—i—i—i  8  12  16  20  24  28  -1  I  I  I  I  I  I  •  •  '  10  I  I  1- .1  1  1  1  15  1  I  20  [Mn], umol L"  1  [Mn], umol L"  Fig. 5.13. X - Y scatter plot illustrating the dependence of dissolved As on the distribution of dissolved Mn. The values (molar concentraions) represent a depth interval spanning from ~3 cm above the sediment surface to ~20 cm below. Coefficients of determination (r ) from linear regression analyses for the respective stations are included. 197  I  25  Similar correlative examinations performed for the winter period (r values of -0.01-0.5) illustrate much weaker associations of As with M n (5.12). This decoupling is consistent with the seasonal development of anoxia and the concomitant change in the redox chemistry at the sediment water interface. The sulphate reduction evident in the winter hypolimnion (section 5.3.2) implies that reactive Mn(IV)-oxides were absent in surface deposits during the time of sampling. Hence, the capacity of the surface sediments to sorb upward diffusing As species should have been greatly reduced. Under such conditions, the diagenetic cycle would have been interrupted; both As and M n species would have diffused freely across the sediment-water interface. Comparison of the porewater plots for Mn (Fig. 5.7) and As (Fig. 5.14) suggests that different mechanisms influence the seasonal distributions of the two elements. Mn porewater maxima, for example, progressively decrease from a winter maximum through the spring, summer and fall.  The winter  enrichments were suggested to stem from inputs of dissolved M n to the porewaters from the overlying bottom waters (see section 5.3.3). Conversely, arsenic profiles decrease from maximum porewater values in the summer to minima in late winter (Fig. 5.14). The disparity implies that the non-steady-state inputs of Mn-rich water to the winter bottom waters did not host significantly elevated concentrations of dissolved As. Instead, the distribution of arsenic may reflect seasonal variability in the kinetics of remobilization and consumption of dissolved As in the porewaters. More specifically, variable rates of reductive dissolution of As-bearing oxides, consumption of dissolved As in the oxic horizons, and authigenic sulphide formation at depth, all contribute to the observed seasonality. Large sub-surface porewater gradients in Balmer Lake sediments support year-round upward diffusion of dissolved As towards the sediment-water  198  Fig. 5.14.  Seasonal peeper profiles of dissolved As for stations 1 and 2 (A and B,  respectively) in Balmer Lake. Replicate samples are represented by double symbols at specific single depths. 199  interface. Rates ranged from 5-175 ug'cm" 'y" ; the large gradients observed in 2  1  the fall account for the upper limit of this range (Appendix F). These values are much greater than those reported for the Laurentian Trough (0.07-0.24 ug'cm^'y" ; 1  Belzile, 1988) a temperate estuary (2.2 ug'cm^'y" ; Widerlund and Ingri, 1995) and 1  various coastal and lake sediments (0.37-2.4 ug'cm^'y" ; Peterson and Carpenter, 1  1986). The largely invariant concentrations of As across the sediment-water interface imply that upward migrating species are consumed by M n and/or Fe oxides in the surface sediments (Figs. 4.18-4.27). However, the reduced sorptive capacity of the surface sediments during periods of bottom water anoxia might suggest that dissolved As species are periodically allowed to diffuse into the bottom waters. In addition, the uncertainty in the position of the sediment-water interface (± 2 cm), and the proximity of As gradients to the sediment surface, suggests that the upper gradient boundary at some sites may extend slightly into the bottom waters. By taking into account a maximum observed As gradient, and neglecting consumption mechanisms of upward migrating species, an upper measure of bottom water input can be estimated. For example, an efflux of 175 ug'cm" 'y~, entering an average water column depth of ~3 m, would result in the addition of -600 ug'L" in one year. Although this level represents a maximum, 1  the result indicates that diffusive transport mechanisms have potential significance within the residence time-scale of Balmer Lake (-230 days). The removal of porewater As below sediment depths of 20 cm at some sites in Balmer Lake is probably due to precipitation of As-bearing sulphide phases (Figs. 4.18-4.19; 4.27). Although the precipitation of  AS2S3  and mixed Fe-  As sulphides have generally not been distinguished, incorporation of As into pyrite has been clearly shown to be an important removal mechanism in several anoxic marine sediments (Huerta-Diaz and Morse, 1992).  Similarly, the  precipitation of a mixed-sulphide phase was proposed by Peterson and  200  Carpenter (1986) to account for pronounced decreases in dissolved As below the sulphate redox transition in near-shore marine sediments. In summary, the geochemical cycle of Fe has a dominant influence on the diagenetic behaviour of As in marine and lacustrine systems, but the influence of Mn on As behaviour is less clear. The lack of understanding partially stems from the dearth of appropriate data in addition to the inherently complex nature of natural systems. It must be realized that the high ionic strength of the polluted water column makes Balmer Lake unique in comparison to "pristine" water bodies. More specifically, the combined burdens of dissolved metal cations (Mn, Ni, Cu, Zn and Fe) exceed 2.5 mg L" in near-interface oxic porewaters. Given 1  such circumstances, sorption of the metals to Mn-oxide phases may shift the surface charge sufficiently to encourage the sorption of anionic As species (Takamatsu et al, 1985).  201  VI. Summary and Conclusions  The distribution and behaviour of trace metals in Balmer Lake are influenced by both natural biogeochemical processes and inputs of miningderived effluents.  The formation of ice in early winter in conjunction with the  high oxidant demand of the organic-rich sediments fosters the gradual development of stratification over the 5 month period of ice cover.  Such  seasonality results in spatial and temporal changes in the redox potential in the bottom waters and sediments. Non-steady-state inputs from the tailings circuits hosting elevated levels of trace elements, sulphate, cyanide and ammonium also contribute to the dynamic character of the system. Dissolved and particulate metal concentrations in Balmer Lake are high, and compare with or surpass values reported for other polluted lacustrine environments. Homogenous distributions of all measured parameters evident during the summer, fall and spring are in marked contrast with the stratified water column conditions in late winter. The trace metal, nutrient and sulphate profiles during this period reflect the combined influences of redox-mediated reactions and inputs of dense mine-waters which migrate laterally across the lake bottom. Although the non-steady-state nature of the latter prevented definitive interpretations of the winter water column, it appears that variability with respect to the duration and intensity of reducing conditions in the bottom waters, and the proximity to effluent discharges contribute to the inter-station contrasts. At sediment depths shallower than 10 cm, the deposits in Balmer Lake comprise contributions from both organic matter- and feldspar-rich natural detritus, and carbonate-, chlorite- and metal-rich tailings inputs. Enrichments of solid phase C u , N i and Zn at shallow depths were significantly lower immediately adjacent to the tailings discharge than at a site situated over 1 km  202  distant. The distribution likely reflects textural sorting of the tailings material, with the finer grain-size fractions hosting higher metal concentrations. Both remobilization and scavenging mechanisms influence the behaviour of trace metals in the sediments and associated porewaters. Early diagenetic reactions in Balmer Lake sediments, for example, render the deposits below 5 cm depth a significant and permanent sink for dissolved C u , N i and Z n . Consumption horizons occur at depths consistent with zones of sulphate reduction, suggesting these elements are sequestered as sulphide precipitates. The extent of authigenic sulphide formation likely reflects the relatively recent addition of mining-derived sulphate to the lake waters. Diffusive transport mechanisms have relatively little effect on the lake concentrations of Cu and Ni. Such processes can account for the removal of only 5 and 1% of their respective inventories with respect to a lake residence time of 230 days. The lake sediments provide a relatively greater diffusive sink for Zn; approximately 13% of the Zn inventory is consumed in this manner. Moreover, sulphide precipitation for these three elements contributes little to their overall accumulation rates. Episodic inputs of tailings material and the deposition of metal-rich particulates represent the principle contributors to the total benthic flux of these elements. Inputs of remobilized Zn and N i from labile particulates result in pronounced spikes near the sediment-water interface. Since such peaks are shown to lie well above the Mn(IV) reduction horizon at some sites, it is suggested that the enrichments stem largely from the oxidation of labile organics, and not from the reduction of Fe and Mn oxides. Contrary to most reports in the literature, the distribution of As appears to be largely dictated by the redox behaviour of Mn. In porewater profiles where the redox horizons for Fe(III) and Mn(IV) are clearly decoupled, dissolved As gradients closely match those of Mn. It is suggested that the large concentrations  203  of divalent cations in the shallow porewaters may significantly alter the surface charge properties of M n oxides to encourage the sorption of anionic arsenic species. The largely invariant concentrations of As across the sediment-water interface during well-mixed periods imply that upward migrating species are consumed by presumably M n oxides in the surface sediments. However, the reduced sorptive capacity of the surface sediments during periods of bottom water anoxia might suggest that dissolved As species are periodically allowed to diffuse into the bottom waters. It is a reasonable expectation that once milling effluents are no longer discharged to Balmer Lake, concentrations of C u and N i should quickly diminish.  This conclusion stems from the predominantly uni-directionally  diffusive transport of these elements and the relatively short residence time of the basin. The same can be perhaps implied for Zn, although a fraction of the total benthic Zn-flux will continue to be regenerated back into the bottom waters. In addition, the settling of abundant organic and detrital particulates will continue to provide an effective transport mechanism of trace metals to bottom sediments.  The apparent involvement of As in the redox cycling M n has  important ramifications with respect to the future water quality in the basin. 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Anthropogenic perturbation of metal fluxes into the atmosphere. In: Changing metal cycles and human health., pp. 27-42. SpringerVerlag.  219  VIII. APPENDICES  Appendix A . Typical ICP/MS operating conditions.  Parameter  Typical value  RF power (kW)  1.35  Argon gas flow rate (L/min): Cooling gas  13.83  Auxiliary  0.983  Nebulizer  1.032  Sampling Position (mm from load coil)  13  Sampler Cone (nickel) from orifice (mm)  1.0  Skimmer Cone (nickel) from orifice (mm)  0.8  Ion lens settings (volts): Extraction  -118  Collector  -2  LI  -2  L2  -45  L3  +5  L4  -45  Pole Bias  -3  Operating Pressure (mbar): Expansion  2.2  Interface  < x lO-  4  Analyzer  2 x lO"  6  Resolution: AM  5.3  220  Appendix B. Quality Assessment/Quality Control (QA/QC) The Q A / Q C program included the analysis of quality assurance samples to assess the precision and accuracy of measurements of the various parameters. Several types of samples were included: laboratory blanks ( D D W with added reagents), field process blanks, laboratory replicates and certified reference materials. Overall, the results for the Q A / Q C samples are indicative of good reproducibility, accuracy and contamination control. I. Dissolved Trace Metals Laboratory blanks for all trace metals were very low, and indicates good contamination control. Blank values were used to determine the detection limits which are reported as six times the standard deviation of the blank results. Table A l . Detection limits and machine precision for trace metal determinations. Detection limits represent six standard deviations of 10 process blank values. Element  Method  Detection Limit  Precision  (ppb)  (%)  Fe  GFAAS  1.8  2.0  Mn  ICP/MS  0.06  2.5  M  ICP/MS  0.10  2.7  Cu  ICP/MS  0.14  2.5  Zn  ICP/MS  0.84  2.6  As  ICP/MS  0.22  2.0  Pb  ICP/MS  0.50  2.0  221  Table A 2 . Analytical results for field blanks. Values represent the average of four samples. Fe  Mn  <1.8  porewater <1.8  Water column filter  Ni  Cu  Zn  As  Pb  <0.06 <0.1  <0.14  1.1  <0.22 <0.5  <0.06 <0.1  <0.14 <0.84 <0.22 <0.5  blank Core blank  Table A 3 . Expected (Exp) and Observed (Obs) values for standard reference materials supplied by the National Research Council (SLRS) and the National Water Research Institute ( T M Series). Observed values correspond to the mean of 10 measurements. The error in the observed values represents two standard deviations. Element  SLRS-2 Exp  TM-21 Obs  Exp  TM-02 Obs  Fe  Exp 50.0 ± 1 6  Mn  10.1 ± 0 . 3  9.46±0.38  6.2 ± 2.0  6.1±0.14  M  1.03 ± 0 . 1 0  1..17±0.10  6.6 ± 2.4  6.47± 0.30  Cu  2.76 ± 0 . 1 7  2.80±0.16  7.6 ± 2 . 2  6.93±0.28  Zn  3.33 + 0.15  3.63*0.32  7.5 ± 2.2  7.80*0.40  As  0.77 ± 0 . 0 9  0.81±0.14  Pb  0.129*0,01  0.148±0.01  5,5 ± 1 , 6  5,20±0,20  222  Obs 50.3±2.0  II. Particulate Trace Metals Table A 4 . Trace metal determinations of the sediment reference material BCSS. The analysis of seven replicates included digestion followed by ICP/MS (Mn, N i , Cu, Zn, As and Pb) and G F A A S (Fe). The error in the observed values represents ± two standard deviations. Digestion  Fe  Mn  M  Cu  Zn  As  Pb  1  32,755  225  78  21  101  19  23  2  37,523  204  69  21  100  17  23  3  33,290  203  64  20  97  17  24  4  177  65  20  95  17  22  5  216  65  18  96  17  23  6  28,073  152  60  21  89  16  21  7  34,273  202  64  19  104  17  24  Mean  33,183 ± 6,800  197 ± 4 6  67 ± 1 1  20 ± 2  98 ± 12  17 ± 2  23 ± 2  Expected  32,900 ± 980  229 ± 15  55 ± 4  19 ± 3  119 ± 1 2  11 ± 2  23 ± 3  Table A 5 . Laboratory blanks for digested polycarbonate filters. A l l values are expressed as p.g L"'.  Digestion  Fe  Mn  Ni  Cu  Zn  As  Pb  1  0.02  0.02  0.01  0.01  0.25  0.04  0.00  2  0.02  0.01  0.01  0.00  0.52  0.00  0.00  3  0.07  0.03  0.06  0.00  1.13  0.13  0.03  4  0.06  0.02  0.01  0.00  4.85  0.09  0.00  5  0.06  0.02  0.02  0.00  3.92  0.10  0.01  6  0.07  0.02  0.09  0.08  0.68  0.10  0.23  7  0.02  0.03  0.06  0.09  2.42  0.10  0.01  Mean  0.05  0.02  0.04  0.03  1.97  0.08  0.04  223  III. Nitrate and Sulphate Table A 6 .  Detection limits and machine precision for nitrate and sulphate  determinations. Detection limits represent six standard deviations of 10 process blank values. Anion  Method  Detection Limit  Precision  NO3-  IC  0.1 pmol-L~l  1.0%  SO4 -  IC  0.1 umol-L"!  1.0%  2  Table A7. Expected (Exp) and Observed (Obs) values for standard reference materials supplied by the National Research Council and the National Water Research Institute. Observed values correspond to the mean of 10 measurements. The error in the observed values represents two standard deviations. Ani-04  N0 -  Ion-96  Exp  Obs  Exp  Obs  Exp  Obs  145 ± 43  141 ± 6  335 ± 3 8  347 ± 24  1629 ± 10  1 6 6 5 ± 100  1170 ± 110  1126 ± 66  1044 ± 1 2 2  1084 ± 8 4  1550 ± 10  1626 ± 120  3  4  Dionex  224  Appendix C. Porewater data  Porewater concentrations of dissolved Fe, M n , N i , Cu, Zn, As, Pb, N H , N 0 " +  4  3  and S0 " in Balmer Lake for the summer, fall, winter and spring field seasons. A l l values 2  4  are reported ugL" , except for N H 1  + 4  (umolL ) N 0 " (pmolL ) and S0 " 1  1  3  2  4  (mmolL ). 1  Positive values for sampled horizons (cm) for both peepers and cores ( C I , C2 and C3) refer to depths above the sediment-water interface.  Note, that due an angled peeper  insertion, the depth scales for the metal and nutrient data differ for the last spring peeper profile.  225  Appendix C cont. Summer-Station 2 Depth  Mn  Summer-Station 2 Fe  NI  Cu  (cm)  C9L'  ugL'  pgL'  35.5 30.3 23.9 18.7 18.7 14.8 13.5 ^3 11.0 9.7 8.4 7.1 7.1 5.8 5.2 4.5 4.5 3.2 2.6 1.9 1.3 0.6 0.0 0.0 -0.6 -1.3 -1.9 -1.4 -3.2 -3.9 -4.5 -4.5 -5.8 -7.1 -8.4 -9.7 -11.0 -14.8 -18.7 -18.7 -22.6 -26.4 -31.6 -36.8 -36.8 -41.9  45  415 429 417 421  173 174 169 175  45 49 43 51  403 408 413 402 417 405 408  163 163 160 161 162 161  48 62 53 63 47 49  407 392 410 408 413  163 153 162 163 167  51 42 45 46 51  404 424 421 402  155 158 169 157  52 49 59 55  281  285  290  297  306  306 313  57  35 45 37 44 60 39 42 57 54 47 33 34 40 71 90 66 57  Zn  As |igL'  279  289  297  301  309  Pb  NH4  0.5 0.4 0.2 1.0 0.0 0.5 0.4 0.3 0.2 0.3 0.5 0.9 0.2 0.3 0.2 0.4  48  79 90  98  342  3.3  87  354 350  3.4 3.3  0.3  320  100  407 396 437 366 429  148 132 139 108 120  63 73 107 99 124  390 591 567 983 992  0.3 0.4 0.2 0.4 0.4  110 490 1050 1400 1890 1510 1410  244 154 110 101 96 76 59  50 15 5 7 20 4 4  69 72 30 30 40 26 20  2430 4080 4960 5044 4862 3896 3028  0.8 1.1 0.5 0.6 1.9 0.9 1.0  558 574 660 618 622 642  881 908 888 909  1630 2150 2860 3240  49 54 56 67  2 1 1 11  13 17 13 27  2561 2828 2405 3175  0.6 0.8 1.0 1.5  655 671 647 622  986  3850  86  6  33  3854  1.1  666  680 820 940 986 961 979 908  75  3.4  3.4 3.3 3.1  79 77 77 80 73 70  0.6 0.3  350  354 341 348  64 72 128 209 269 348 354  49  S04  1  44 338 380 412 460 494  N03  ugL f»molL>molL"'mmoiL"  346 373 358  3.2 3.3 3.3  433 371 374 400 381 329 194 180 68 92 91  3.3 3.2 3.3 3.2 3.3 3.2  51 77 35 41 36 15 14 23 15 14 26 27 13  Depth  31.0 27.1 25.8 24.5 23.2 21.9 20.6 19.4 19.4 18.1 16.8 15.5 14.2 10.3 9.7 7.7 7.1 6.5 5.8 5.2 4.5 3.9 3.2 2.6 1.3 0.0 -1.3 -1.9 -2.6 -2.6 -3.2 -3.9 -4.5 -5.2 -5.8 -8.5 -7.1 -7.7 -8.4 -9.0 -10.3 -10.3 -11.6 -12.9 -14.2 -15.5 -19.4 -23.2 -23.2 -27.1 -31.0 -36.1 -41.3 -41.3 -46.4  3.1 3.1 3.0  2.6 2.6 2.4 2.2 1.8 1.2 1.2 1.0 0.8 0.6 0.6 0.6 0.3  226  Mn  (cm)  Fe  NI  eg!-' 289  65  Cu  Zn  cgL' 421  164  47  As  Pb  egL'  figL"  NH4  285  0.5  1  30  N03  S04  pmotL''pmolL'mmolL 72  344  3.3  366 341  3.3 3.3  75  288 285  26  415 414  161 154  45 42  266 278  0.4 0.4  298  60 42 54 33 36 37 35 34 48 32 40 56 47 57 38 79  430  171  52  296  0.3  72 78  343  3.2  421  151  46  285  0.4  100  340  3.3  423  160  48  0.4  105  343  3.3  156 147 143 135 131 123 112 88 81 51 28 18 13 10 6 18  61 86 125 90 88 143 88 100 52 60 34 45 28 32 41 20  0.4 0.3 0.4 0.5 0.5 0.5 0.4 0.6 0.6 0.5 0.5 0.5 0.5 0.4 0.3 0.2  90 93 95 87  1677 1544  412 485 460 484 489 461 427 322 295 218 209 171 167 142 123 116 113 111 105  7 6  23 21  286 293 297 372 384 523 530 446 451 885 1049 1502 2135 2536 3003 3319 3525 3985 3895 3794 3798  1688 1460 2056  105 79 72  29 7 2  38 22 23  1464 1535 2923 3175  72 67 59 73 74 129  1 1 1 0  26 23 16 22  12  79  368 360 401 385 364 327 344 179 194 87 80 35 71 52 56 54 59 54 29 30 32 10 14 14 9 8 9 8 4  3.3 3.3 3.2 3.1 3.2 3.1 2.3 2.8 2.8 2.6 2.4 2.5 2.3 2.3 2.0 2.0 2.1 1.9 1.7 1.7 1.4 1.0 1.0 0.8 0.6 0.5 0.4 0.4 0.2  291  287 299 304 361 366 421 414 446 451 506 517 549 756 757 858 844 916 908 918 891 855 869 806 800 812 779 764 814 845 828 909  43 58 60 78 197 381 744 837 1598 1457 1490  4230  3725 2946 2808 2767 2099 1870 1977 2208 2192 2355  0.3 0.1  85 85 67 128 272 333 360 354 456 441 439  0.2 0.3 0.5  510 580 494 609 628 636  0.4 0.3 1.2 0.2  686 650 610 588  0.5  606  Appendix C cont. Summer-Station 1  Summer-Station 1  Depth Mn (cm) (igL  NI Cu M9L' pgL'  Zn  34.8 33.5 32.3 31.0 29.7 29.7 23.2 23.2 18.1 18.1 14.2 14.2 12.9 12.9 11.6 11.6 10.3 9.0 9.0 7.7 7.7 6.5 5.2 4.5 4.5 3.9 3.2 3.2 2.6 1.9 1.3 0.6 0.0 -0.6 -0.6 -1.3 -1.3 -1.9 -1.9 -2.6 -2.6 -3.2 -3.2 -3.9 -3.9 -4.5 -4.5 -5.2 -5.2 -6.5 -6.5 -7.7 -9.0 -9.0 -10.3 -11.6 -12.9 -12.9 -14.2 -14.2 -15.5 -16.8 -19.4 -19.4 -21.9 -21.9 -23.2 -24.5 -24.5 -28.4 -28.4 -33.5 -37.4 -37.4 -42.6 -42.6  423  57  1  237  237 243 242  427 441 426  154  As  Pb  C9L' 285  0.0  150  44  287  0.0  49 47 48  295 292 294  0.0  72  316  0.1  242  421  155 154 153  256  455  167  0.0  253  428  157  67  298  0.1  253  434  158  51  304  0.0  244 237  437 428  162 160  50 50  307 310  0.0 0.0  234  438  155  48  295  0.0  245 239 233  428 443 432 424 431 438 418 429 443 434 397  45 52 47 48 46 45  317 302 303 311 301 313  0.0 0.1  240 234 230 229 224 242 256 260  164 155 156 163 149 161 150 159 151 156 144  44 51 47 50 60  304 294 300 288 298  0.0 0.0 0.0 0.0 0.0 0.0  333 344  469 376  143 114  60 54  315 335  0.0 0.0  395  382  505 513 537  213 220 205  114 104  58 59  0.0 0.0 0.0  670 613  0.0 0.0  51 42  34 30  481 433 690 689 718  106  12  18  1875  0.0  118  13  20  1481  0.1  0.0  752  86  8  15  2663  0.0  765  85 84 81 75  6 6 6 7  13 14 15 26  2809  0.0 0.0 0.1 0.1  106 78  49 17  57 21  810 819 764  86 88 86 84 77 70  16 13 13 12 11 14  39 23 22 26 24 24  2871 2757 2794  740 739  67  6  17  701  67  6  17  768 787 793 799 812 815 803  688 693 699 687 740  56 57 54  2 2 2  14 26 14  67 122  5 86  21 82  3086 2720 2685 2419 2501 2561 2598  0.3 0.2 0.1  Fe N04 S04 NH4 MSL' umolL" mmolL jimolL' 36  31 35 342 330 191 178 151 134 76 41 37 33 31 35 35 29  354  3.4  77  401  3.4  85  77  3.4  390  414  3.5  3.4  91  83 111 110  447  3.3  115 108  31 29 30 29 34 34  403  3.4  107 115 100 101 100 87  35 31 31 162 75 65 475 462 300 267 1023 980 1307 1423 1445 1365 1845 1536 1620 1596  0.1  1792  2066  0.1  1442 1523 1775  0.6  3.5  31 29  2673 2705  1391 1390 1375 1694  63  360  1633 1510 1467 1619  0.1 0.0 0.1 0.1  3.4  31 28 32  0.1 0.1 0.1 0.1 0.1  1424  350  596 396 406 412 365 356 360  3.4 3.4 3.4 3.4 3.5 3.5  185  197 199 149  3.6 3.5 3.6  298  53  3.4  450  44 27 27 29 13 15 14 31 31 32 31 33  28 28  3.2 3.7  130  397  447 511 611  3.1 3.6 3.6 3.4 3.3 3.2  682 684 660  3.0  642  2.6 2.2 2.2 1.8 1.2  2482 2173  23  0.8  3848 3963  17  0.6  655 624  570 570 646  565 567 593  553  227  Depth Mn Ni (cm) ugL' (igL 30.3 29.0 23.9 22.6 22.6 17.4 17.4 16.1 14.8 13.5 12.9 12.3 11.6 11.0 10.3 9.7 8.4 7.7 7.7 7.1 6.5 6.5 5.8 5.2 4.5 3.9 3.9 3.2 2.6 2.6 1.9 1.3 0.6 0.0 0.0 -0.6 -1.3 -1.3 -1.9 -2.6 -3.2 -3.9 -4.5 -5.2 -5.2 -5.8 -7.1 -7.1 -8.4 -8.4 -9.7 -9.7 -11.0 -11.0 -12.3 -12.3 -13.5 -14.8 -16.1 -16.1 -17.4 -18.7 -20.0 -22.6 -23.9 -27.7 -29.0 -32.9 -34.2 -38.1 -39.3 -41.9  Cu Zn As Pb Fe N04 S04 NH4 pgL' egL' cgL' cgL' &*gL" (*molL"'mmotL>molL" 1  291 250  459 392  166  57  276 310  428 479  161 159 161  53 47 52  290 281 279  272  420  159  48  284  276  0.2 0.1 0.1 0.6  59 32  355  44  392  3.3  32  349 364  3.3 3.5  89  366  3.4  92  352  3.4  93  32 43 50 34  365  3.4  97  93  389  3.4  95 97  560  3.4  421 414 418 358 321  3.4 3.4 3.5 3.4 3.4  87 84 85 82  198 115 73 34 21 29 18 18  3.6 3.6 3.7 3.7 3.7 3.7 3.8 3.7 3.7  580 600  19 18  3.7 3.5  529 646  27  3.3  641  35  3.2 3.2 3.0 2.4 2.7  649  0.2  312  467  218  83  339  2.2  288  448  162  48  285  0.2  43  296  447  154  46  278  0.2  32  270 305 303 322 355  426 472 451 463 581  162 169 161 159 166  49 49 50 47 49  288 287 290 275 289  0.2 0.2  36  337 355 329 297 295 305 279 364  589 602 489 450 439 460 412 514  157 165 163 157 163 166 162  47 49 49 49 49 48 49  276 281 283 279 291 288 294  356 307 297 301  510 431 414 414  170 189 167 163 169  50 52 53 60 59  279 325 299 295 306  324 397 384 440 443 599 637 702 712  422 511 457 488 394 368 304 169 176  147 152 163 133 129 74 67 18 16  65 80 85 91 113 83 77 30 36  308 344 342 449 494 625 610 844 942  838 861 861 890 984  126 102 105 88  20 21 20 40  1469 1731 1683 1858  16 107  1962 1115  1.4 0.3 0.3 5.7  0.2 0.2 0.2 0.2 1.6 0.2 0.2 0.4 0.8 0.8 0.5 0.2 0.2 0.2  34 33 32  87 135 39 41 35  1.4 0.6  36  0.2 0.2 0.2 0.2 0.2 5.4  44 40  1115  76 127  7 6 7 7 5 5 63  0.3 0.2  1017  79  15  31  1679  1.1  1085 1012 1101  76 76 73  6 5 4  17 18 17  1415 1415 1381  0.3 0.3 0.2  71 87  79 694 541 1799 1726 1826 1689  1650 1628 1432 1370 1394  83 85  65 59 67  135 190 277 355 423 459 488 542  620 601 528 626  1106  77  4  18  1362  0.2  959 1038 987  84 60  4 4  18 14  1835 1742 1066  0.3 0.2  844 975  66 48 71  3 5 2 2  15 14 15  1096 1265  0.3 0.4 0.4 0.4  1311 1151  71 71 82  2.4  558  1.8  489  1499  80 73  1.4 1.4 1.1  545  2092 2615  68  0.9  505  2437  68  0.9  503  510  Appendix C cont. Fall-Station 2  Fall-Station 2  Depth  Mn  (cm)  MflL'  30.3 18.7 18.7 13.5 12.3 11.0 9.7 5.8 5.2 4.5 4.5 3.9 3.2 2.6 1.9 1.3 0.6 0.0 0.0 -0.6 -0.6 -1.3 -1.9 -2.6 -3.2 -3.9 -4.5 -4.5 -5.2 -5.8 -5.8 -6.5 -6.5 -7.1 -7.7 -8.4 -8.4 -9.0 -9.7 -9.7 -11.0 -11.0 -12.3 -12.3 -13.5 -14.8 -14.8 -16.1 -17.4 -17.4 -20.0 -23.9 -23.9 -27.7 -27.7 -32.9 -32.9 -39.3 -45.8 -45.8  205' 200 202 206  285 288 289 285  201 202  NI  Cu  Zn  As  C9L-'  egL'  106 106 107 108  55 45.8 45.2 55.2  202 209 212 203  0.3 0.2 0.2 0.4  102 90  304 301  3.6 3.7  58 62  90  302  3.6  64  271 263  100 97  40.5 41.1  211 201  0.2 0.4  95 102  289 289  3.6 3.5  62 62  205  271  100  43.5  195  0.3  109  268 275  3.2 3.2  62  201  272  100  42.3  206  0.3  293 285 291 295 292 294 288 287  3.6 3.4 3.6 3.6 3.6 3.6 3.5 3.5  63 63 60 61 60 62  198 196 203 215  281  105  44.6  285 265  107 98  49.7 52.7  217 474 426 508 445 517 439  256 258 203 216 156 166 144 109  92.5 92.1 24.4 29.9 27.1 28 30.2 20.2  53.7 54.3 123.5 131 101 130 84.5 47.8  594 632  122 98.2  31.5 16.6  754 745 736 681 822 822 689 775  128  815 823 822 832 791 759 697 690 741 876 943 947 957 958 999  Pb  Fe  N03  S04  NH4  f*9t-" (imolL mmolL pmolL 1  86 93 103 100 93  205 194 215 222  0.3  233 561 565 745 591 714 859  0.9 0.8 0.6 1 0.6 1.1 0.8 1.2  63 103 75 105 91 186  292  3.5  52  162 128 39 53  85 95 139 140  67.2 38.5  1302 1712  3 1.7  476 579  28.2  76.1  1.1  224  1.9  185  79.8 113 75.9  18.6 27.2 16.3  28.9 67 24.1  2.5 2.1 2  959 404 986  20 23 25  19.7 10.2  42.3 26.6  2.2 2.3  460 1237  259 262  90.9 92.6 67.4  17.6 17.7 10.3  54.7 48.5 23.4  1.1 1.1 1  662  21 29 28 18  1.9 1.7 1.8 1.8 1.6 1.7  204 225 220  101 67  1769 1763 2183 1689 2202 2153 1735 2098  19 20 20 . 35  3.0 2.9 2.6 2.5 2.6 2.2 2.1  1.6  332  931  16  1.5  364  58.4 64.5 69 82 110  9.6 24.7 26 38 28.6  21.7 32.8 33.7 56 97.5  946 810  16 2  1.4 1.2 1.2 1.1 1.1  384 408 418 436  61.5 58.2 60.2 32.3  18.3 13 13 3.8  39.3 37 34.2 13.5  0.9 0.7  458 502  34.5  4.6  14.5  2791 2840 2781  64.7 51  7.2 7.4  26.6 29  2797 3257  2250 2364 2410 2646 2255 2066 1946 1938 1983 2848  0.6 0.5  0.6 0.6 0.7 0.8 0.8  112  645 427  57  166 170  1.1 1.1 1.1 1.2  632 1010  1 4 4 2 2  2108  2  0.6  513  1.4  2455  13  491  0.8 1.8  2495 3009 2997  14 8 8  0.6 0.6 0.5 0.3 0.3  497 507  228  Depth  Mn  (cm)  MgL  Ni  30.3 18.7 18.7 9.7 5.8 3.2 2.6 2.6 1.9 1.3 0.6 0.6 0.0 -0.6 -0.6 -1.3 -1.3 -1.9 -1.9 -2.6 -3.2 -3.2 -3.9 -4.5 -5.2 -5.2 -5.8 -6.5 -7.1 -7.1 -7.7 -7.7 -8.4 -8.4 -9.0 -9.7 -9.7 -11.0 -11.0 -12.3 -12.3 -13.5 -14.8 -14.8 -16.1 -16.1 -17.4 -17.4 -18.7 -20.0 -21.3 -22.6 -23.9 -27.7 -32.9 -39.3 -39.3 -45.8 -45.8  202 206 198 196 191 191 200  Cu  Zn  C9L'  1  As  Pb  cgL'  cgL'  0.4 0.4 0.3 0.3 0.3 0.3  295 290 284 283 296 285  116 117 109 112 114 109  65 56 71 51 50 51 54  238 235 238 230 226 229 229  111 109 106 107  55 80 56 55  240 239  198  289 308 280 274  416  326  63  431  302  44  459 448 451 398  223 198 135 134 122 106 84 84 46 70  17 13 20  34 5 25  55  15  543  62  648 629 612 609  207 223  354 376 319 389 426 416 502  67 66  302 304 300 309 310  3.4 3.5 3.3 3.4 3.4  67 66 64 65  310 313  3.4 3.4  63 60 62  331 335  3.4 3.6  45 34  258 255 230  3.3 3.2 2.9  78  39.4 51.9  155 126  2.4 2.5  135 142  41.6 62.1 92.5  70 56 27 27 25 18 15  2.1 2.1 1.9 1.9 1.9 1.7 1.7  174 158 185 181 213 219  216.5 212.1 121.8 123.9  22  1.6  254  19  1.6  244  447.9 423 337.5  21 20 20  1.5 1.5 1.5  287  421.2  16  1.4  356  13 6  1.3 1.1  382 421  320  0.5  217  353  0.5  49  170  43 38 37 26 29 19  238 77 77 90 48 37  546 518 578 767  50.7 47.6 57.1 50.3 42.2  747 1005 975  0.7 0.1 0.9  20  17  29  1988  1  97  29  63  2048  1.7  98 100 91 89  80 81 38 37  63 64 59 58  1903 1900 1950  1.4 1.4 1.3 1.3  21 20  63 61  2.3  67.4 119.6  0.3 1941 1893  1.1 1  NH4  3.5 3.6  203  0.5  S04  311 312  218  1197 1595 1768 1757 2222  83 84  92.4 94.9 93.7 97.3 94.6 84.4 92.2 91 78.9 106.9  0.5 0.6 0.6 0.5 0.8 1.1  N03  **gL"' f*motL""mmolL>molL"  0.3 0.3 0.4 0.4  52 687 671  Fe  453.9  79  346  6  1.1  445  7 7  0.8 0.8  463  695  47  16  28  1912  0.9  650.7  11  0.9  489  756 793 830 839 823 934  42 39 23 24 24 34  8 8 1 2.3 2.5 4  28 22 10 9 11.5 15  3287 2049 2164 1749 1730 4062  0.8 1.7 0.8 0.9 0.9 1.8  919.2 1301 2417 3504  8 17 14 15  537 526 547 533  3780 3934  10 11  0.7 0.6 0.5 0.5 0.5 0.3 0.3  581  Appendix C cont. Fall-Station 1  Fall-Station 5  Depth Mn (cm)  Ni Cu Zn As CSL-' M9L' C9L' ugL'  33.5 21.9 21.9 12.9 9.0 8.4 7.7 7.1 8.5 6.5 5.8 5.2 5.2 4.5 3.9 3.9 3.2 2.6 1.9 1.3 1.3 0.6 0.0 -0.6 -1.3 -1.3 -1.9 -2.6 -2.6 -3.2 -3.2 -3.9 -3.9 -4.5 -4.5 -5.2 -5.8 -6.5 -6.5 -7.7 -7.7 -9.0 -10.3 -10.3 -11.6 -12.9 -12.9 -14.2 -14.2 -15.5 -16.8 -18.1 -19.4 -20.6 -20.6 -24.5 -24.5 -29.7 -29.7 -38.1 -42.6 -42.6  240 203 207 208 209  280 284 288 282 287  117 121 122 120 121  85 48 46 49 51  214 214 218 218 218  0.17 0.48 0.53 0.17 0.17  119 93.7 92.8 90.5 90.3  357 352 350 351 349  3.8 3.7 3.7 3.7 3.8  63 63  204  274  117  44  216  0.23  94.8  350  3.8  63  202  283  118  48  212  0.18  354  3.8  63  251 204  339 282  136 117  55 46  264 211  0.18 0.17  95.2 97 99.8 96.2  290  120  58  0.11  269 281 286 283 280 286 295 291 297 529  431 477  463 366  113 119 118 116 118 118 118 114 111 60 59 57 45  52 56 64 62 57 56 60 63 71 277 275 265 133  212 211 211 205 214 205 212 215 209 208 204 190  3.8 3.7 3.7 3.6 3.7  62 59  205 206 202 207 213 210 207 210 205 214 221 466  355 355 349 350 349  482  370  39  439 439 465  292 288 267  509 523 622  Pb Fe N03 S04 NH4 ugL i»molL mmolL umolL 1  88.4 86.4  64 64  66 64  195 287  0.2 0.2 91.7 357 0.17 91 355 0.26 94.5 353 0.16 94.9 358 0.16 93.3 350 0.31 120.4 348 0.17 88.4 354 0.17 86.9 330 0.19 37.6 234 0.2 0.26 34.4 147 0.25 35.1 95  121  327  0.37  41 42 29  88 89 62  392 383 431  0.27 0.3 0.41 37.96  235 232 211  31 25 38  50 40 68  430 476 477  0.42 0.53 0.38  45.4 40.1 54.8  536 537 570 505  242 251 236 223  42 42 28 36  84 84 74 83  676 685 712 686  0.36 0.33 0.48 0.4  38.9 34.1 45.9 51.3  627 617  181 179  19 33  62 81  834 894  597 590  185 177  30 29  83 76  852 845  0.34 62.3 0.24 70.44 69.28 0.31 74.24 0.24  561  164  42  78  811  0.4  93.24  19  513  207  47  93  710  0.36  10  1.5  430  564 570 714  90 94 38  27 28 12  42 42 14  829 830 801  0.58 0.66 0.65  10 11 10  1.4  404  1.1  408  680 722 719  32 36 36  5.8 29 29  10 19 17  775 813 816  0.31 0.94 1  45.9 46 152.8 145.3 494.4 530 803.1 969 1002  11 7 7  0.8 0.6 0.6  398 365  45 43.2 34.3  70 70 25 30 30  3.7 3.7 3.7 3.8 3.7 3.7 3.8 3.7 3.7  65 65 64 63 57 65 56 130  3.3 3.8 3.7 2.9  126 193  2.9  207  173  2.7  232  2.7 2.5 2.7 2.6 2.4  226 275 271  46 23 23 11 8  2.3 2.1  279 290  2.0 2.0  314 329  8  1.9 1.8  354  1.9  388  19 16 13  290  229  Depth Mn NI Cu (cm) egt-' cgL' cgL'  Zn As CSL"' cgL'  Pb Fe N03 S04 NH4 eg!-' (igL- (imolL'mmotL>motL  36.1 24.5 24.5 19.4 15.5 11.6 11.6 7.7 7.7 7.1 6.5 6.5 5.8 5.2 4.5 4.5 3.9 3.9 3.2 3.2 2.6 1.9 1.3 1.3 0.6 0.6 0.0 -0.6 -0.6 -1.3 -1.3 -1.9 -1.9 -2.6 -3.2 -3.9 -5.2 -5.2 -6.5 -7.7 -7.7 -9.0 -9.0 -10.3 -10.3 -11.6 -11.6 -12.9 -14.2 -15.5 -16.8 -16.8 -18.1 -18.1 -19.4 -19.4 -24.5 -29.7 -29.7 -34.8 -34.8 -40.0 -40.0  1.3 0.3 0.3 0.3 0.4 0.3 0.3 0.3  502 120  313 326  3.7 3.7  63 65  98 96 97  313 323 313  3.7 3.7 3.6  67 67 75  110 104  319 319  4.4 4.4  94  196 184 179 186 192 204  1  292 294 279 296 298 307 300 309  118 109 104 110 110 110 107 96  189 188 181 89.7 73.7 57.5 55.6 80.3  213 204 209 215 223 255  264 257 254 259 254  324  100  91  0.3  108  318  4.7  105  323 323 324  96 94 92  117 79.5 75  382 370 369 376 376  0.3 0.3 0.1  319 320 318  4.7 5.1 5.1  106 110 109  250 252 244  323 327 321  93 94 91  72 71 67  371 376 359  0.2 0.2 0.2  99 112 112 100 121  317  5.0  111  103  324 321 305 304 307  92 91 82 83 57  65 70 60 62 37  365 346 306 312 277  0.2 0.2 0.2 0.2 0.1  317 317 317 320 313  5.1 5.1 5.2 5.0 5.1  109  256 260 253 256 310  312  5.0  114  352 353  199 125  43 21  42 19  234 220  0.3 0.3  78  14  26  406  0.7  4.8 4.7 4.8 3.1  92 107  577  309 295 300 159  377 376  60 60  9.7 9.7  12 12  266 272  0.4 0.4  146  3.3  58  771 383 744 747 619 1008  75 69 79  11.5 11.2 13  18 42 25  0.5 0.6 1  96 100  9 9  2.3 2.3  39 50  69 91  13 17  43 35  428 375 590 598 725 873  0.9 1.1  209 191  5 5  52 55  1312  106  20  31  886  1.2  210  1384  112  22  32  1265  1.2  2.7  68  1470  120  25  35  1539  1.1  227 222 206  4 4 4  2.1 1.9 2.0 2.3  729 967  102 124  18 18  29 34  1285 1258  1.1 1.2  1639 1745  151  24  44  1388 1441  1.4  1468  133  18  36  1808  1784 904  84 43  23 5  20 9  3041 2862  683  40  5  15  2498  235  338  97 98 102 104 77 74 74 66 142 146 69 68  110 112 114  38  51  3  2.1  92  2 2  2.4 2.3  125 129  264 257  2  2.4 2.4  158  1.1  203  218  221 219  2 2 1 2  2.8  1 0.9  243 207  1.8  485 449  2.6 2.3 2.3 2.2 2.4 2.4  187  211  2 2  199  Appendix C cont. Winter-Station 2  Winter-Station 1  Depth Mn (cm) ugL  NI C9L'  Cu  27.7 27.7 22.6 16.1 16.1 14.8 14.8 9.7 9.7 8.4 7.1 7.1 6.5 5.8 5.8 5.2 4.5 4.5 3.9 3.2 2.6 2.6 1.9 1.3 1.3 0.6 0.6 0.0 -0.6 -0.6 -1.3 -1.3 -1.9 -1.9 -2.6 -2.6 -3.2 -3.2 -3.9 -4.5 -5.2 -5.8 -5.8 -6.5 -7.1 -7.7 -8.4 -8.4 -9.0 -9.7 -1Z3 -12.3 -13.5 -13.5 -14.8 -14.8 -16.1 -17.4 -17.4 -18.7 -18.7 -21.3 -25.2 -26.4 -26.4 -31.6 -31.6 -36.8 -36.8 -41.9 -48.4 -48.4  432 439 483 501  95 97 74 56  1  487 486 671 812  942  420  24  Zn As Pb C9L' i*gL' cgL' 187 190 141 181  197  194 193 209 238  414  0.6 0.6 0.5 0.3  0.2  954  410  21  198  488  0.1  941 970 937 977 969 930 988 930  417 402 402 399 408 395 404 404  22 21 20 21 22 20 20 21  218 216 224 212 219 251 211 269  512 516 488 553 551 497 565 504  0.2 0.1 0.1 0.3 0.3 0.1 0.1 0.2  972 965  406 400  21 21  270 280  589 535  0.2 0.2  970 978 978 915 926 961  408 404 399 405 403 469  21 22 23 19 19 19  256 285 275 241 245 257  616 606 530 606 622 468  0.1 0.2 0.4 0 0 0.2  949  418  17  248  532  0.1  1044  446  17  243  536  0.05  Fe N03 S04 NH4 MgL umolL mmolL (imolL' 1  66  249  6.9  288  57 71 74  193  9.0  284  34  10.1  7  10.3  393 389 457  132 129 218 220  5  10.5  449  6 8 10 10  10.3 9.8 10.5 10.2  401 413 393  7 14  10.0 9.3  413 433  398  11 9 9 7  9.2 9.5 9.5 8.9  337 455  10 9  9.1 8.6  393 377  234 219 286 291  8  8.1  393  8  7.8  417 425 401  259 270 306 268 329 281 285 374 311  217  23 24  4.5 4.5  429 413  1117  365  16  223  580  0.06  163  8  3.9  433  1125 1083  293 298  15 15  180 181  647 630  0.1 0.1  158  8  3.1  461  1003  220  11  132  780  0.1  287  6  2.8  453  1001  154 154  9.7 9.5  95 95  835  0.4 0.3  333  885 731  115 68  11 19  78 54  736 610  0.3 0.2  668  43  23  33  632  0.2  182 63 65 66 67  11 13  2.2 1.7  489 481  10 9  1.5 1.5  477  656  37  28  38  703  0.2  72  590 560 608  43 42 43  26 25 35  52 52 42  594 557 613  0.2 0.2 0.2  43 40  9  0.8  649 667 641  40 48 43  25 13 24  32  579  337  6  0.7  41  363  0.4 1.6 0.7  9  0.6  501  644 643 573 651 653  39 39 36 46 46  23 23 31 22 21  33 33 143 52 50  300 295 272 314 323  1.1 1.1 0.3 0.9 0.9  204 201 192 193 75 241  8  0.4  501  7 5  0.3 0.3  493 497  12  1.1  485 473 481 485  230  Depth Mn (cm) (jgL  1  25.2 20.0 13.5 13.5 12.3 12.3 9.7 8.4 7.1 5.8 4.5 4.5 3.9 3.2 3.2 2.6 1.9 1.3 0.6 0.0 0.0 -0.6 -0.6 -1.3 -1.9 -2.6 -3.2 -3.9 -4.5 -4.5 -5.2 -5.2 -5.8 -5.8 -6.5 -7.1 -7.7 -8.4 -8.4 -9.0 -9.7 -10.3 -11.0 -11.6 -12.3 -14.8 -14.8 -16.1 -16.1 -17.4 -18.7 -18.7 -20.0 -20.0 -21.3 -21.3 -23.9 -27.7 -29.0 -34.2 -34.2 -39.3 -39.3 -44.5 -44.5 -51.0 -51.0  531 641 868 843 887  NI Cu Zn As Pb gL' cgL-' cgL' MgL (.gL  1  1  M  367 407 452 422 414  67 65 59 50 39  88 116 147 149 161  159 172 198 204 255  0.27 0.23 0.51 0.14 0.15  Fe N03 S04 NH4 MgL' (imolLmmolL>molL1  41 42 48 40 39  _  256 259  10.3 11.3  437 487  132  13.0  52  12.8  524 528 582  910  425  41  164  266  0.14  42  46  12.6  578  986  462  37  162  275  0.14  46  48 49  12.1 12.3  536  981 996 891 967 938 981 947  457 458 422 425 444 408 441  37 37  158 157  0.25 0.22  12.1  549  156 169 153 154  71 68 49 60 45 65 44  51  34 41 34 39  280 283 260 271 268 268 260  465 483 495 482 571 466 447 421  36 36 40 34 45 36 49 35  161 163 185 174 248 200 223 174  298 298 269 301 297 295 278 297  0.11 0.11 0.12 0.12 0.13 0.12 0.12 0.17  10.4 11.6 10.4 10.8 10.4 10.1 10.4  528 549 520 528 520  1046 1058 989 1067 1055 1073 1024 1090  35 46 33 37 31 31 27 25 24  9.9 8.3  520 495  21 19  7.5 8.9  499 528  19 19  8.2 8.2  536 532 495  0.17 0.15 0.13 0.14  66 65 45 72 49 81 53 87 83  528  1143  342  28  136  297  0.17  110 115  14  6.3  1182  264  19  98  369  0.2  300  12  5.5  499  1234 1232  204 206  14 14  70 68  411 405  0.22 0.22  477 467  20  4.7  507  1258  176  10  59  394  0.13  502  15  4.2  503  1342  136  9  45  385  0.4  532  1309 1372  115 93  8 7  40 30  369 410  0.23 0.21  400 359  72  7.4  34  408  0.11  183 171  3.4 2.7 2.7  507 495  1325  16 15 16  1214 1215  41 42  4 4.2  17 20  441 450  0.27 0.31  493  7  2.7  495  7  1.6  483 491  30 8  1.1 0.9  491 478  5  0.7  474  2 2 5  0.5  445  0.4  416  3  0.2  404  789  31  5  22  401  0.11  776  24  2.9  13  387  0.63  718 739  25 24  1.9 2.4  ias 9.8  365 476  0.43 0.58  697  29  4  17  386  0.53  628  33  6.3  21  250  0.59  718  33  4  16  491  491  134 126 666 683 1476 1438 1017 692 655 2081 2186  Appendix C cont. Winter-Station 4 Depth Mn Ni (cm) M9L' cgL' 25.2 20.0 20.0 16.1 12.3 12.3 9.7 9.7 8.4 5.8 5.8 4.5 3.9 3.2 3.2 2.6 2.6 1.9 1.3 1.3 0.6 0.6 0.0 -0.6 -1.3 -1.9 -2.6 -3.2 -3.2 -3.9 -4.5 -5.2 -5.8 -5.8 -6.5 -7.1 -7.7 -8.4 -8.4 -9.0 -9.7 -10.3 -11.0 -12.3 -13.5 -14.8 -16.1 -17.4 -18.7 -18.7 -20.0 -21.3 -25.2 -25.2 -26.4 -30.3 -30.3 -31.6 -31.6 -35.5 -36.8 -41.9 -41.9 -43.2 -47.1 -47.1 -51.0 -52.2 -52.2  356 371 372 454 556  491 450 443 441 425  Winter-Station 6 Cu MSI."'  290 256 255 209 145  Zn As Pb Fe N03 S04 NH4 MgL' cgL' MgL' (igL MmolL mmolLMmolL  Depth Mn Ni Cu Zn As Pb Fe N03 S04 NH4 (cm) MgL MgL' MgL' MgL' cgL' MgL' MgL' (»molL mmoiL MTIOIL  97 94 96 121 138  26.4 21.3 20.0 18.7 18.7 17.4 14.8 13.5 11.0 11.0 9.7 9.7 8.4 7.1 7.1 5.8 5.2 4.5 4.5 3.9 3.2 2.6 2.6 1.9 1.9 1.3 0.6 0.0 -0.6 -0.6 -1.3 -1.9 -2.6 -3.2 -3.2 -3.9 -4.5 -5.2 -5.8 -5.8 -6.5 -7.1 -7.1 -8.4 -9.0 -9.7 -11.0 -12.3 -13.5 -14.8 -14.8 -16.1 -16.1 -17.4 -18.7 -20.0 -23.9 -23.9 -25.2 -25.2 -29.0 -29.0 -30.3 -34.2 -35.5 -40.6 -41.9 -45.8 -45.8 -49.7 -51.0  -1  897 966 963 956 998 967 974 950  320 315 281 316 299 304  43 39 35 38 39 36  141 151 144 147  274  41  959  275  1023 1009 972 1028 1005 1047 987 1024 1021 989  368 354 353 330 315  0.4 0.3 0.3 0.3 0.3  0.1 0.3 0.3 0.1 0.3 0.1  157  302 305 299 287 298 317 316 277  39  161  274  0.2  304 306 291 353 331 343 356 336 338 338  36 36 36 40 36 36 33 31 31 43  151 153 167 159 181 147 168 123 124 181  308 324 282 342 339 357 405 409 416 489  0.1 0.1 0.2 0.2 0.3 0.2 0.4 0.2 0.2 1.3  957 1009 1017  280 257 255  24 26 26  98 94 94  501 463 459  0.3 0.2 0.1  988  211  27  81  480  978 991  146 141  27 27  86 64  472 491  152  0.3  176 141  477 424  4.6 4.7  130 118  132 115  364 290 283 229  4.9 5.5  142 171  5.9  200 200  67 151 76 128 80 147 144 75 124 74 125 78 151 71 153 75 65 141  64  11  5.3  299  0.3 0.3  71  31 30  5.1  311  12  4.5  481  0.2  79  480  0.2  77  921  29  524  45 47  3 3.1  25 26  436  841  45  7.2  32  684  70  14  65  729 726  53 55  7.2 7.5  43 45  0.3 0.4  162 125  0.2 0.2  125  283  0.2  71  191  0.2  64  211 206  291  0.3  44  2  291 286  77 55  41  50  6.7 6.7 6.6 7.2  307 295  15  940  274  6.7 6.7 6.5 6.7 5.7  11  502  6.4  23 17 17 21 13 17 12  107  32  286 291 291 295  291 291 291  104  4  33 21 22 27  6.8 6.8 7.0 6.6 6.8  6.8 6.9 6.8  958  79  55 55 31 24 24  20 21 20  992 972  -  0.2 0.2  5.9  45 25  12 6.8  46 19  235 294  0.1 0.5  52 597 615  705 696  26 26  7.1 7  20 19  134 131  0.5 0.5  436  712 715  112 110  7.2 6.8  64 57  244 249  0.7 1.3  471  303  303  13  3.2  303  15  2.5  344  27  1.9  303  25 21  1.6 1.4 1.4  299 282  3  1.0  254 278  4  0.8  237  80  623 741  295 295 315  4 4  0.6 0.4  196 155  231  1  317 287  410 365  185 148  125 81  325 387  0.6 0.5  325 315  418 408  185 179  85 80  382 355  0.5 0.5  313  414  182  83  375  0.5  112  314 284  410 376  185 163  86 81  379 365  0.4 0.5  112 112  313  410  178  89  408  0.4  108  302 311 317 310 332 314  397 409 419 404 435 415  164 168 156 157 170 157 170  94 91 104 98 94 112 99  465 435 552 540 458 612  0.4 0.5 0.6 0.5 0.5 0.5 0.6  101  314 318 321 311 314 327 298 321 324 334 324  419 419 426 409 413 421 431 407 409 378 422  378 324  313 311  1350 1714  0.4 0.4 0.6 0.5 1 0.5 0.7 0.6 0.4 0.4 0.4 0.4 0.7 0.6 0.7 0.6 0.6 1 0.6  87  355 376 341 348  145 137 114 118 122 131 135 116 144 111 126 128 101 186 147 124 124 82 114  695 696 545 732 613 879 578 721 952 839 1085  334 327 343 325  157 162 163 162 154 158 172 149 157 135 148 151 126 138 126 130 129 97 101  302  253  72  84  1974  1  282  211  56  70  2264  1  81  198  120  25  36  3028  0.7  60  186 155  108 103  23 26  19 19  3507 3237  1  88 89  998 1334 1158 1468  129 90  93 94 99  91 90 84 85 83 97 86 84 132 123 134 84 84 141 75  126  76  17  15  4875  1  158  95  61  14  7  6034  1  317  50 50  18 18  4 4  4 3  8536 8483  0.5 0.6  127 129  45  14  4  3  9314  0.5  108  39 36  10 10  3 3  6 2  10160 0.3 10428 0.3  101 112  36  12  4  2  11813 0.3 11478  129  35  13  4  3  13909 0.6  216  436 432  4.1 4.0  98 94  473  4.8  98  427  4.0  94  454  4.4  98 94  439 434 440 435 459  4.1 4.2 4.2 4.4 4.5  90  463  4.5  453 457 457  4.4 4.2 4.3  78 42 66  465 475 464  4.5 4.4 4.4  58 54 50  440 426 417 405  4.4 3.9 3.7 4.1  50 46 50 50  385 379  4.2 4.0  62 58  312  3.6  66  221 221 236  3.3 3.3 4.0  86  191  3.6  98  108  3.1  113  66  2.4  125 129  32 20 20  2.1 1.9 1.9  141 149  24  1.8  153 161  7  1.9  165  2  1.9  161  5  2.1  173  90 82 86  7o  82  Appendix C cont. Winter-Core 1  Winter-Core 2  Depth Mn Ni (cm) C9L' H9L"'  Cu  570 583  456 458  740 839 863 874  1.0 1.0 1.0 -0.3 -1.3 -1.3 -1.8 -1.8 -2.5 -2.5 -3.5 -4.5 -4.5 -5.5 -5.5 -7.0 -7.0 -9.0 -11.0 -11.0 -13.0 -13.0 -15.5 -18.5 -18.5 -24.5 -24.5 -30.5 -30.5  Zn c9L'  As  107 104  142 135  223 223  1.5 1.5  491 621 634 617  67 25 25 20  126 122 126 179  345 419 434 446  0.3 0.1 0.1 0.4  1082 288 1057 283 986 121 1017 42  66 69 12 1.3  54 56 23 45  438 448 1021 1343  0.6 0.6 0.5 0.1  1033 1023 990  13 12 9.4  1.3 2 1.1  4.6 8.6 3.3  1020 0.03 1035 0.08 709 0.05  252  989 943  6.6 6.3  1.1 1.1  3.8 4.8  533 434  0.02 0.02  222 221  848 873 848 743 737 682  7 6.5 4.8 4.5 4.5 3.4  1 0.8 0.6 0.8 0.9 0.6  16 12 6.9 323 318 26  688 685 423 320 321 83  650 743  2.9  0.6  20  17 17  0.07 116 0.02 116 0.02 436 0.2 794 0.2 0.08 2411 2469 0.2 3074 2963  Pb Fe N03 S04 M9L' Hfl-L"' (imolL mmoll^ 83 82 86 50 31  267 278  10.6 11.2  58 6  12.0 11.0  56 57 76  14  11.1  2 3 3 7  10.7 10.9 8.2 7.6  2  7.6  2 3 2 3 2 2  6.0 6.0 5.8 4.7 4.8 4.0  3 3 2  3.2 2.6 2.5 1.1  6 6  0.5 0.5  104 387 408 461  Depth Mn Ni Cu Zn As Pb (cm) (•SfL- CSL' C9L' eg<-' pgL' cgL  Fe N03 S04 pgL' fimouVmmolL  1.0 1.0 -0.3 -0.8 -0.8 -1.3 -1.3 -1.8 -1.8 -2.5 -2.5 -3.5 -3.5 -4.5 -4.5 -5.5 -7.0 -7.0 -9.0 -9.0 -11.0 -13.0 -13.0 -15.5 -15.5 -18.5 -18.5 -24.5 -24.5 -30.5 -30.5  102 104 44 44  272 308 54 19  10.3 8.9 12.5 12.3  47 49 53  12 11 6  10.9  60  7 7 3 3 3  1  1  232  755 561 692 561 1083 663 1020 676 1071 694 1041 694  129 163 71 68 70 51  145 141 129 142 154 173  234 237 386 374 383 400  1 1.4 0.4 0.4 0.4 0.5  1096 686 1139 744 1354 586  40 40 35  114 114 74  434 445 472  0.6 0.6 0.8  1348  167  22  28  740  0.3  1365  44  5  7.9  1020  0.2  1284 1319 1401 1346  13 6 6 5  4 4 3.6 4  7.7 3.9 3.9 4.4  735 483 505 358  0.2 0.2 0.2 0.3  1216 1103 1039 897  8 4 6 4  5.5 3.5 4.8 3.9  4.1 4.1 4.2 3.9  309 268 254 271  0.2 0.2 0.2 0.2  897  4  3.6  4  207  0.2  722 735 897  2.7 3 3.3  4.4 4.6 4.2  5.4 5.7 4.3  44 44 9  0.3 0.3 0.3  90 91 469 460 440 324 331 291 316 390 408 837 2197 3094 3121  12 1 1 2  11.3 11.2 10.2 8.7 8.1 6.7 7.1 6.5  1 1  5.2 5.3 4.9 3.9  1  3.1  3 3 5 5 4 4  2.2 0.9 0.4 0.4  Appendix C cont. Winter-Core 3 Depth Mn (cm)  193 1.0 1.0 190 -0.3 281 -0.8 397 -0.8 -1.3 364 -1.3 -1.8 264 -1.8 259 -2.5 420 -3.5 457 -3.5 456 -4.5 451 -4.5 -5.5 308 -5.5 -7.0 186 -7.0 183 -9.0 72 -9.0 -11.0 33 -11.0 -13.0 32 -15.5 32 -15.5 -18.5 27 -18.5 27 -21.5 -21.5 30 -24.5 37 -24.5 35  Spring-Station 1 Ni  Cu Zn As egL' C9L' MgL' (jgL'  Pb egL'  Fe  6 3 95 295  9 7 37 147  1888 1878 2658 776  0.5 0.4 0.7 0.9  56 46 45 105  395  147  88  1815  0.5  300 301 502 457 452 276  47 46 65 66 65 41  23 3291 25 3263 114 2151 96 1289 94 1290 49 1365  0.5 0.5 1 0.8 0.9 0.5  21  17  54  11  8  1780  0.4  S04  cgL' jimotLmmolL  243 252 321 516  100  N03 340 418  4.5 4.7  467  4.4  57 55 24  421  4.3  83 80  365 318  45  230 229 124 126 42  0.4  31 30 40  13  2  73  2202 2212 5017  0.3  74  8.2  1  22  7553  0.3  6.9 6.3  1 0  4 2  7933 9424  0.3 0.2  82 81 86 108  5.1 4.9  1 1  2 3  5.9 6.1 5.9  1 1 1  2 3 2  9928 0.1 10491 0.1 10142 0.4 11649 0.4 11303 0.5  4.4  15 13 21  131  24 28 27 30  118  25  109 106  28  4.1 3.9  3.4 3.0 1.1 2.8 2.8 2.7 2.7 2.6 2.7 2.7  233  Depth  Fe  (cm)  cgL'  NI  Cu  MgL'  MgL'  MgL  MgL'  31.5 31.5 26.4 19.9 19.9 18.6 18.6 13.5 12.2 10.9 10.9 9.6 9.6 8.3 7.0 5.7 4.5 3.8 3.2 2.5 1.9 1.9 1.2 1.2 0.6 0.6 -0.1 -0.7 -0.7 -1.4 -2.0 -2.0 -2.6 -3.3 -3.9 -4.6 -4.6 -5.2 -5.9 -5.9 -7.2 -8.5 -9.7 -9.7 -11.0 -12.3 -12.3 -14.9 -14.9 -16.2 -17.5 -17.5 -20.1 -21.4 -21.4 -22.6 -23.9 -25.2 -25.2 -26.5 -26.5 -29.1 -29.1 -31.7 -34.3 -34.3 -36.8 -36.8 -39.4 -42.0 -42.0 -44.6 -45.9 -45.9  82 78 77 82  290  409  266  61  224  1.7  265  3.5  187  295 290 295  405 397 415  259 247 260  57 59 60  233 228 233  1.8 1.5 1.4  263  3.5  185  269  3.6  75  301  401  241  60  231  1.3  262  3.6  191 187 185  72 72  288 290  400 403  240 243  60 60  229 227  1.1 1.1  73  287  401  236  61  226  1.1  259 261  3.5 3.5  72 73 80 68 82 68 70 82  288 291 302 307 302 303  415 400 401 400 404 407  246 238 239 237 238 241  61 63 63 64 63  222 222 230 233 228 227  0.96 0.97 0.12 1.3 1.3 1.1  265 256 260 255 266 259 260  3.6 3.4 3.4 3.5 3.6 3.4 3.6  189 185 187 185 191 185  305  426  254  70  236  1.7  299 300 310 309 306 305 312 314 432 415 505 528  417 411 420 415 424 399 409 420 420 421 426 446  249 246 249 242 249 240 236 241 219 220 156 150  69 70 69 70 86 78 89 88 97 95 111  232 230 244 237 233 251 241 238 284 260 435 424  1.2 1.2 1.5 1.5 1.4 2.2  262 264 259  3.6 3.8 3.5  185 189 178  267  3.5  178 178  1.1 1.7 1.3 1.9 1.6  283 268 272 262 175 148 118  3.7 3.7 3.7 3.9 4.2 4.7 5.0  180 185  724 830  426 474  85 66  93 114  657 701  0.89 1.1  65  54 59 185 189 277 133 127 366 366  966 444 963 360 1094 203 1071 204 1104 107 1067 128  42 29 14 15 10 21  129 120 61 61 28 42  877 1.4 1022 1.2 1747 1.6 1743 1.5 2090 0.75 1769 1.6  30 20 8  5.7 5.7 5.9 6.0 5.9  10 7  5.5 4.7  1015  107  7.2  26  2316 0.87  8  3.9 4.0  201  972 967  136 134  24 23  56 46  2001 1995  3.3 3.2  2  3.4  515  146  887 913  98 89  25 14  44 21  1532  2.7 1.7  11  2.9  511  9  2.6  498  3  2.4  505 498  9 9  1.8  466  11  1.3 1.3  440  g  1.1  412  4  1.1  397  74 84 74 91 104 85 88 79 77 70 52  111 112  Mn  762  109  54  As  Zn 1  1  68  706  Pb  N03  S04  NH4  MgL' nmolLmmolLnmolL  1.1  3.4  175  785  90  31  55  637  2.2  169  731 714  101 100  31 30  65 63  298 282  2.6 2.2  239 238  800  124  30  93  298  2.7  89  513 501  191 192  56 57  117 112  889 879  1.2 1.2  185  193 230 277 331 410 453 485 503 511 511 503  Appendix C cont. Spring-Station 1 Depth Mn Ni (cm) cgL' fgL 31.8 30.7 27.2 27.2 26.0 21.4 21.4 20.3 15.6 15.6 14.5 9.9 8.7 4.7 4.1 3.5 2.9 1.8 1.2 0.6 0.0 -0.5 -1.7 -2.3 -2.8 -3.4 -4.0 -4.0 -5.2 -5.2 -6.3 -6.3 -7.5 -7.5 -8.6 -9.8 -9.8 -10.9 -12.1 -13.2 -14.4 -14.4 -15.6 -15.6 -16.7 -17.9 -19.0 -20.2 -20.2 -21.3 -22.5 -22.5 -23.6 -24.8 -26.0 -29.4 -30.6 -31.7 -32.9 -32.9 -37.5  Cu 1  Zn As egL' pigL'  M  Pb Fe gL' (igL'  277  360  198  100  199  0.7  74  274 272  355 363  194 197  56 58  203 207  0.5 0.5  72  272  335  180  53  200  0.4  74 76  PH  7.61 7.77 7.81  276 272  335 352  186 196  62  201 211  0.8 0.6  72  278  380  204  60  213  0.7  74  280  363  199  64  215  0.8  72  7.70 7.78 7.95  275  352  200  66  203  0.9  69  8.02 7.99  275  360  197  74  192  0.9  69  8.00  278  345  194  69  201  1  76  346  364  184  85  216  1  82  401  396  173  85  262  0.9  402 409 596  373 379 369  173 176 132  77 87 89  266 277 349  1 1 1  79 79 83  747  339  97  81  448  0.8  858 842 834  259 209 209  53 44 44  66 61 62  637 765 747  0.6 1.3 0.14  792 776 792 816 735  180 142 112 112 114  47 42 27 27 26  75 60 40 39 63  742 748 916 914 812  1.4 1.5 0.9 0.8 1.3  720 755  109 96  23 19  57 55  848 939  1.4 1.3  80 75 93 92  762 767  92 93  13 13  35 37  900 917  1 0.9  144 138  747  75  11  34  664  0.7  188  699  67  6.3  42  562  1  311  662  64  5.2  33  402  1.1  513  651 662 682  57 54 67  1.7 1.7 7.5  18 22 19  451 457 738  0.3 0.3 0.5  834 829 925  64 50 47 46 70 70 83 65 74  7.80  7.80  7.80  7.70  7.54 7.53 7.57 7.54  7.46 7.57 7.35 7.56 7.41 7.32  7.56  Depth N03 S04 NH4 (cm) |*motL" mmolL" jimoU"' 39.3 34.7 28.9 23.1 23.1 17.3 15.0 14.4 13.9 13.9 13.3 12.7 12.1 11.6 11,0 10.4 9.8 9.8 9.2 8.7 8.1 7.5 6.9 6.4 6.4 5.8 5.2 4.6 4.0 3.5 3.5 2.3 2.3 1.2 1.2 0.0 0.0 -1.2 -2.3 -2.3 -3.5 -4.6 -5.8 -6.9 -6.9 -8.1 -8.1 -9.2 -10.4 -11.6 -12.7 -12.7 -13.9 -15.0 -15.0 -16.2 -17.3 -19.6 -24.3 -28.9 -28.9  260 254 254 239  3.6 3.5 3.6 3.2  254  3.5  253 247  3.5 3.5  183  245  3.5  185  252 248 253 251 243 251 249 253 250 286 253 252 280 251 247  3.6 3.5 3.5 3.6 3.5 3.6 3.6 3.6 3.6 3.6 3.6 3.6 3.6 3.5 3.5  185 185 188 185  244  3.5  173  255  3.7  188  265 260 267  3.8 3.8 3.7  192  186  4.7  296  57  5.9  437  12  188 190 188 185 181 185  183 181 181 179 183 179 183 177 177  198  520  5  5.2  529  19  2.9 3.0  488  23 78 17  2.0 1.4 1.4 1.4  458 392 439  234  Appendix D. Sediment Data Compositional sediment data arranged by sample depth for each core. Major elements, C a C 0 , Organic C 3  and N are in wt.%; minor elements are in ppm. Depth (cm) Corel station i  Core 2 Station 1  0.25 0.75 1.25 1.75 2.50 3.50 4.50 5.50 7.00 9.00 11.00 13.00 15.50 18,50 24.50 30.50  0.25 0.75 1.25 I. 75 2.50 3.50 4.50 5.50 7.00 9.00 II. 00 13.00 15.50 18.50 24.50 30.50  Core 3 Station 6  0.25 0.75 1.25 1-75 2.50 3.50 4.50 5.50 7.00 9.00 11.00 13.00 15.50 18.50 21.50 24.50  CaC0 N Si Ti Al Fe C a Mg Na K P (wt.%) (wt.%)(wt.%)(wt.%)(wt.%)(wt.%)(wt.%)(wt.%)(wt.%)(wt.%)(wt.%)(wt.%) 3  0.65 0.96 1.25 0.98 0.92 0.82 0.85 0.49 0.04 0.01 0.00 0.00. 0.00  5.80 e.85 3.41 2.28 1.45 2.97 3.08 4.29 8.78 8.84 8.64 8.59 8.48  0.00 0.00  7.75 7.54  1.015 1.595 0.859 0.739 0,360 0.630 0.567 0,730 0.860 0.787 0.755 0.740 0.725 0.687 0.676 0.663  0.67 0.98  5.93  1.059  6.23 3.66 2.06 1.53 2.74 3.04 5.44 8.55 8.81 8.78 8.68 8.43 8.09 7.86 7.89  1.379 0.956 0.661 0.370 0.577 0.593 0.744 0.774 0.754 0.750 0.730 0.697 0.672 0.658 0,685  1.18 0.98 0.93 0.80 0.76 0.46 0.05 0.01 0.01 0.01 0.00 0.00 0.00 0.00  2.44 1.78 2.13 2.70 2.36 2.23 2.41 2.56 2.64 3.06 3.18 3.14 3.20 3.25  48.14  0.56  11.67  8.77  4.15  3.57  48.44 50.19 50.45 48.54 49.47 48.89 55.37 46.79 55,96 53.40 53.27 54.57 54.72 55.62  0.59 0.63 0.65 0.58 0.59 0,56 0.55 0.51 0.54 0.54 0.55 0.57 0.59 0.59  11.26 13.05 13.93 13.28 13.59 13,65 12.48 9.80 12.34 11.56 11.86 12.55 13.00 13.17  10.54 11.87 11.63 10.77 10.83 10.16 5.98 5.09 4.30 4.27 4.46 4,46 5.44 4.61  5.08 5.05 5.15 4.45 4.64 3.84 3.14 2.84 2.94 2.88 j 2.89 2.81 2.70 2.70  4.71 4.95 5.06 4.47 4.57 3.81 2.01 1.59 1.84 1.77 1.77 1.87 2.00 1.76  46.09 47.94 48,93 50.38  0.56 0.58 0.63 0.65  11.16 11.24 12.65 13.81  9.91 10.62 11.52 11.71  4.73 4.95 5.02 5.13  4.07 4.73 4.67  49,43 48.18 53.13 52.78 55.14 51.67 51.73 49.22 51.78 53,33  0.58 0.55 0.54 0.53 0.54 0.53 0.54 0.54 0.57 0.58  13.65 12.37 12.25 11.55 12.30 11.15 11.34 10.95 12.11 1Z60  10.69 8.76 5.13 4.41 4.50 4.32 4.31 4.45 4.65 4.52  4.44 3.78 2.98 2.88 2.90 2.85 . 2,85 2.76 2.68 2,62  4.33  1.16 0.75 0.72 0.60 0.49 0.58 0.64 0.76 1.28 1.37 1,41 1.42 1.46 1.53 1.61 1.62  1.96  0.22  1.63 1.63 1.67 1.73 1.85 1.87 2.23 2.02 2,15 2.16 2.16 2.33 2.35 2.36  0.16 0.15 0.16 0.19 0.15 0.17 0.25 0.22 0.26 0.23 0.26 0.25 0.23 0.24  1.65 1.66 1.70 1.68  0.23 0.16 0.15 0.18  1.87 1.92 2.04 2.08 2.07 2.05 2.24 2.34 2.44 2.30  0.15 0.18 0.23 0.23 0.25 0.22 0,23 0.22 0.21 0,23  1.15  5.00  3.49 2.06 1,75 1.96 1.73 1.77 1.65 1.85 2,07  0.81 0.75 0.55 0.56 0.55 0.59 0.86 1.32 1.36 1.37 1.40 1.48 1.54 1.58 1.55  S  0.32 0.43 0.26 0.28 0.39 1.20 1.19 1,43 0.93 0.57 0.47 0.43 0.39 0.29 0.20 0.16  As ppm  4744 5029 4677 4450 4053 1272 318 173 115 74 25 0 0  0.37 0.42 0.24 0.33 0.42 0.92 1.23 1.34 0.67 0.48 0.45 0.42 0,35 0.36 0.20 0.15  4196 5014 5021 4751 4415 3505 751 265 181 114 54 30 0 0  1.43  0.150  51.02  0.51  9.15  10.26  8.02  6.32  0.43  1.08  0.07  0.32  2192  3.48 2.91 0.97 1.14 0.87 0,64 0.60 0.56 0.33 0.35 0.06 0.05 0.03  0.515 0.388 0.108 0.170 0.101 0.050 0.053 0.037 0.017 0.019 0.011 0.007 0.002 0.018 0.014  50.92 49.81 49.64 50.32 49.55 48,94 49.09 47.39 48.73 42.62 49.92 48.88 49.47 52.23 48.05  0.54 0.57 0.56 0.57 0.57 0.58 0.58 0.56 0.56 0.52 0.58 0,57 0.56 0.59 0.57  9.96 10.32 9.40 9.93 10.53 10.56 10.71 10.06 10.25 8.28 10.43 9.50 9.30 10.46 9.39  9.80 10.75 10.19 10.88 11.07 10,99 10.86 10.88 11.16 10.41 10.97 10,36 10.19 11.64 11.12  6.33 7.23 8.27 7.53 7.02 7.50 8.01 8.22 9.16 9.08 9.26 9.22 9.34 9.05 8.55  4.93 5.95 6.45 6.37 6.23 6.66 7.04 6.98 6,97 5.68 7.70 7.20 7.13 7.45 6.59  0.64 0.44 0.30 0.40 0.34 0.31 0.25 0.20 0.25 0.26 0.22 0.23 0.25 0.23 0.24  1.24 1.23 1.14 1.15 1.23 1.28 1.24 1.24 1.19 1.18 1.14 1.05 1.06 1.16 1.13  0.11 0.10 0.07 0.08 0.07 0.07 0.06 0.06 0.04 0.05 0.05 0.06 0.06 0.07 0.05  0.33 0.19 0,07 0.09 0.11 0.11 0.04 0.14 0.05 0.07 0.06 0.06 0.04 0.14 0.07  3756  235  2527 3193 3570 3104 2813 2387 2022 1978 1976 1827 2072 2217 2451  Appendix D cont.  Core 1 Station 1  Depth  Ba  Co  Cr  Cu  Mn  Ni  Pb  Rb  Sr  V  Y  Zn  Zr  (cm)  ppm  ppm  ppm  ppm  ppm  ppm  ppm  ppm  ppm  ppm  ppm  ppm  ppm  0.25 0.75 1.25 1.75 2.50 3.50 4.S0 5.50 7.00 9.00 11.00 13.00 15.50 18.50 24.50 30.50  322 209 189 154 177 200 201 237 453 492 522 546 533 565 585 558  130 310 179 98 73 317 208 203 64 27 24 21 18 18 17 17  273 372 478 399 352 334 325 285 115 95 90 96 91 95 96 97  8280 6893 2687 1530 1412 18136 4060 3782 708 132 58 49 39 30 22 14  1136 1457 1503 1600 1850 1645 1601 1391 718 690 683 694 654 651 624 620  2177 4615 4105 2759 1748 4516 3302 2793 827 206 127 102 86 69 52 42  132 129 105 171 196 146 159 209 29 12 12 13 16 14 15 12  68 67 45 32 28 0 0 7 58 72 78 76 74 82 83 87  158 117 119 103 99 112 114 127 192 211 220 225 224 233 228 237  167 196 220 247 277 249 253 250 110 96 94 97 95 102 105 104  22 22 27 27 31 58 56 43 23 17 15 18 18 21 19 19  2509 4781 2959 5856 2753 2222 1922 2902 1013 255 184 158 145 127 129 122  83 62 59 57 60 64 67 76 111 132 142 145 143 150 148 156  Core 2  0.25  335  140  291  8539  1215  2413  122  68  162  177  26  2837  81  Station 1  0.75 1.25 1.75 2.50 3.50 4.50 5.50 7.00 9.00 11.00 13.00 15.50 18.50 24.50 30.50  239 197 150 184 198 204 280 478 525 515 547 527 549 601 575  253 165 91 74 226 238 170 40 26 22 25 18 21 16 19  393 467 378 366 336 324 257 117 95 92 93 89 93 100 97  6122 2928 1443 1128 11739 6433 2939 438 122 64 44 30 23 15 17  1438 1462 1705 1860 1694 1629 1261 732 681 676 684 649 644 648 627  4370 4045 2411 1775 3498 3780 2368 468 200 125 93 75 64 41 43  123 118 187 205 165 174 117 42 24 13 18 17 14 10 8  60 42 35 23 0 0 22 65 72 75 71 77 84 90 84  126 116 104 99 111 113 142 202 213 217 216 225 223 228 229  200 217 259 275 252 264 199 111 96 90 92 89 96 108 100  28 24 26 25 48 56 43 19 19 17 17 20 17 21 17  4095 3769 5596 2797 2101 2112 2610 492 225 180 152 136 131 124 123  64 58 59 59 62 64 83 118 130 134 140 142 146 150 145  Core 3  0.25  89  66  609  1163  2060  406  90  12  99  205  19  428  37  Station 6  0.75 1.25 1.75 2.50 3.50 4.50 5.50 7.00 9.00 11.00 13.00 15.50 18.50 21.50 24.50  131 110 70 74 77 76 76 68 65 56 52 42 40 49 44  141 111 58 52 47 33 38 34 24 23 26 29 26 27 40  511 586 634 599 610 637 648 663 676 681 675 663 658 658 644  5160 3525 1041 1873 1097 436 423 250 64 24 36 28 26 69 57  1677 1922 2128 1901 1958 2124 2166 2212 2313 2255 2315 2354 2289 2240 2169  882 780 448 764 768 549 420 313 196 176 194 178 174 204 237  77 107 107 126 141 161 159 162 154 153 131 115 117 144 124  27 12 12 13 12 12 14 13 8 14 9 B 4 11 4  111 99 98 88 91 90 91 90 89 91 93 89 89 93 96  196 221 225 220 244 256 247 246 249 231 238 228 222 237 234  23 24 21 22 20 22 21 22 20 25 24 20 19 21 22  875 770 480 994 1145 822 727 534 319 282 305 276 260 359 373  51 42 36 40 41 38 40 38 34 32 37 36 31 37 34  236  Appendix E. Core Logs  Core BL-1 Notes: - Collected 2 April 1994 - Clear skies, light breeze, -3°C - Core collected with Pedersen corer through 1 m of ice - Station 1,4.0 m - Core length 38 cm  Core Log: Whole core: Very fine-grained, silty clay Top 3 mm: Brownish floe, very fine-grained. 0 to 2 cm: Finely laminated alternating light/forest green; millimetre scale or less. ~0.9 cm: Distinct dark lamina. ~ 1 to 2.5 cm: Alternating light/dark fine laminations, cohesive clay. 2.5 to 4.0 cm: Homogeneous grey zone, less cohesive. 4.0 to 8.0 cm: Black layer, evidence of methane bubbles. 8.0 to bottom: Homogeneous, grayish/ chocolate brown.  Core BL-2 Notes:  - Collected 3 April 1994 - Overcast skies, 15 knot breeze, blowing snow, -5°C - Core collected with Pedersen corer through 1 m of ice - Station 1,4.0 m - Core length 41 cm  237  Core Log: Stratigraphy identical to Core BL-1  Core BL-3 Notes: - Collected 3 April 1994 - Clear skies, 15 knot breeze, blowing snow, -5°C (-15 to -20 with wind chill) - Core collected with Pedersen corer through 1 m of ice - Station 6, 3.0 m - Core length 54 cm  Core Log: 0 to 3 cm: Fine silty gray tailings material 1 to 1.5 cm: Separation plane 1 mm thick with < 1 mm brownish layer on separation surface. 2.5 cm: Small separation 2-3 cm wide with thin brownish layer on planar separation surface. Myriad of tiny bubbles (< 1 mm) appeared progressively in the top several em's of core during core logging (temperature effect?). ~3 to 4 cm: Very fine-grained irregular amber/ chocolate/ gray brown layer. Layer thickness unclear due to smearing down sides of core barrel. 4 to 25 cm: Very fine-grained light gray silt. ~26 cm: Separation plane ~ 1 mm thick, discontinuous laterally, water-filled. ~27 cm: Separation plane < 1 mm, with ~1.5 cm thick dark gray layer (medium silt) immediately above (i.e., ~25-27 cm). 27 to 41: Variously gray-coloured medium silt to very fine silt tailings material. Irregular coarsening below 35 cm. Slight variations in colour but no obvious stratification.  238  Appendix F. Diffusive Fluxes  The diffusive flux (J ) can be estimated using Fick's First Law (ignoring both bioturbation z  and advective transport):  J = Dj (j) dc F dz z  where: J = flux z  Dj = in situ diffusion coefficient (Li and Gregory, 1974) F = formation factor & 1.4 for silty clay (Manheim, 1970) (j) = porosity & 0.87, assuming an average particle density of 2.65 g-cm~3 dc = concentration gradient dz The following data tables present diffusive influx estimates (positive values) for dissolved Zn, Cu, N i , N 0 " and S0 ", and effluxes (negative values) for dissolved Zn, N i , 2  3  Mn, As and N H .  4  Diffusion coefficients were calcuated from in situ surface sediment  4  temperatures measured during the summer (17°C), fall (3°C), winter (3°C) and spring (10°C) sampling periods. For the sake of convenience, fluxes are expressed in both grams or moles per cm y" (Flux 1) and grams or moles per m d" (Flux 2). 2  1  2  239  1  Appendix F cont. Downward Influxes Season  Summer  Fall  Winter  Spring  Summer  Fall  Spring  Summer  Fall  Winter  Spring  Station  Element  Dj  Gradient  Fluxl  Flux 2  g/crrr  i4j/crrr7y  ng/m /d  4  2  2  Zn  5.98E-06  2.4E-08  2.8  77.4  2  Zn  5.98E-06  3.3E-08  3.8  105.1  1  Zn  5.98E-06  2.1 E-08  2.5  67.4  1  Zn  5.98E-06  2.3E-08  2.7  75.3  2  Zn  3.81 E-06  2.4E-08  1.8  48.3  2  Zn  3.81 E-06  5.3E-08  3.9  108.0  1  Zn  3.81 E-06  1.1E-07  8.3  226.6  5  Zn  3.81 E-06  2.3E-08  1.7  46.3  2  Zn  3.81 E-06  3.3E-08  2.4  66.9  1  Zn  3.81 E-06  3.3E-08  2.5  68.2  4  Zn  3.81 E-06  3.1 E-08  2.3  63.9  6  Zn  3.81 E-06  2.6E-08  2.0  53.5  1  Zn  4.89E-06  3.7E-08  3.5  96.6  1  Zn  4.89E-06  8.0E-09  0.8  21.0  2  Cu  5.74E-06  2.1 E-08  2.3  63.3  2  Cu  5.74E-06  2.7E-08  3.0  81.9  1  Cu  5.74E-06  4.6E-08  5.2  141.8  1  Cu  5.74E-06  5.9E-08  6.7  183.3  2  Cu  3.82E-06  5.7E-08  4.3  116.7  2  Cu  3.82E-06  3.3E-08  2.5  68.0  1  Cu  3.82E-06  3.5E-08  2.6  71.2  5  Cu  3.82E-06  3.3E-08  2.4  66.9  1 1  Cu  4.78E-06  7.1 E-08  6.7  182.4  Cu  4.78E-06  3.6E-08  3.4  92.8  2  Ni  5.66E-06  7.3E-08  8.0  220.5  2  Ni  5.66E-06  7.2E-08  8.0  218.2  1  Ni  5.66E-06  1.3E-07  14.2  389.0  1  Ni  5.66E-06  8.6E-08  9.5  261.0  2  Ni  3.56E-06  3.1 E-08  2.2  59.9  2  Ni  3.56E-06  7.4E-08  5.2  141.2  1  Ni  3.56E-06  9.1 E-08  . 6.4  174.2  5  Ni  3.56E-06  1.2E-07  8.4  230.4  2  Ni  3.56E-06  5.8E-08  4.0  110.8  1  Ni  3.56E-06  6.0E-08  4.2  115.4  4  Ni  3.56E-06  3.7E-08  2.6  70.9  6  Ni  3.56E-06  3.7E-08  2.6  71.6  1  Ni  4.61 E-06  4.0E-08  3.6  97.8  1  Ni  4.61 E-06  5.7E-08  5.1  139.9  240  Appendix F cont. mol/crrv  mmol/cm /y  mmol/m7d  1.3E-07  0.04  1.1  4  Summer  Fall  Spring  Summer  Fall  Winter  Spring  2  2  N03  1.57E-05  2  N03  1.57E-05  1.3E-07  0.04  1.1  1  N03  1.57E-05  1.5E-07  0.05  1.3  1  N03  1.57E-05  1.3E-07  0.04  1.1  2  N03  1.08E-05  9.8E-08  0.02  0.6  2  N03  1.08E-05  8.3E-08  0.02  0.5  1  N03  1.08E-05  8.9E-08  0.02  0.5  5  N03  1.08E-05  7.8E-08  0.02  0.5  1  N03  1.33E-05  4.2E-08  0.01  0.3  1  N03  1.33E-05  5.6E-08  0.01  0.4  2  S04  8.68E-06  1.3E-07  0.O2  0.6  2  S04  8.68E-06  1.1E-07  0.02  0.5  1  S04  8.68E-06  1.2E-07  0.02  0.6  1  S04  8.68E-06  1.1E-07  0.02  0.5  2  S04  5.65E-06  3.6E-07  0.04  1.1  2  S04  5.65E-06  4.7E-07  0.05  1.4  1  S04  5.65E-06  3.9E-07  0.04  1.2  5  S04  5.65E-06  7.8E-07  0.09  2.4  2  S04  5.65E-06  3.3E-06  0.37  10.0  1  S04  5.65E-06  6.4E-07  0.07  1.9  4  S04  5.65E-06  4.2E-07  0.05  1.3  6  S04  5.65E-06  2.1E-07  0.02  0.6  1  S04  7.17E-06  2.9E-07  0.04.  1.1  1  S04  7.17E-06  2.9E-07  0.04  1.1  Element  Dj  Gradient g/cm"  Flux 1 jxg/crrf/y  Flux 2 ng/m A  pward Effluxes Station  Summer  Fall  Winter  Spring  Fall  2  2  Zn  5.98E-06  -1.7E-08  -2.0  -53.5  2  Zn  5.98E-06  -1.8E-08  -2.1  -56.2  1  Zn  5.98E-06  -1.6E-08  -1.9  -51.8  2  Zn  3.81 E-06  -4.1E-08  -3.1  -84.0  2  Zn  3.81 E-06  -7.8E-08  -5.8  -159.3  1  Zn  3.81 E-06  -1.6E-07  -12.0  -328.3  2  Zn  3.81 E-06  -2.1E-08  -i.5  -42.0  1  Zn  3.81 E-06  -2.9E-08  -2.2  -60.3  6  Zn  3.81 E-06  -1.6E-08  -1.2  -33.0  1  Zn  4.89E-06  -8.2E-09  -0.8  -21.5  1  Zn  4.89E-06  -4.3E-09  -0.4  -11.4  1  Ni  3.56E-06  -3.6E-07  -24.9  -682.2  241  Appendix F cont. Summer  Fall  Winter  Spring  Summer  Fall  Winter  Spring  Summer  Fall  Spring  2  Mn  5.60E-06  -1.3E-07  -13.8  -377.0  2  Mn  5.60E-06  -1.6E-07  -17.8  -487.4  1  Mn  5.60E-06  -1.2E-07  -13.6  -372.4  1  Mn  5.60E-06  -1.1E-07  -11.6  -318.0  2  Mn  3.50E-06  -2.0E-07  -13.6  -371.5  2  Mn  3.50E-06  -1.4E-07  -9.3  -256.0  1  Mn  3.50E-06  -3.8E-07  -25.9  -708.3  5  Mn  3.50E-06  -1.4E-07  -9.5  -259.3  2  Mn  3.50E-06  -6.5E-08  -4.4  -121.4  1  Mn  3.50E-06  -2.9E-08  -2.0  -54.0  1  Mn  4.55E-06  -1.7E-07  -15.0  -411.5  1  Mn  4.55E-06  -1.3E-07  -12.0  -327.6  2  As  7.50E-06  -1 2E-06  -174.6  -4782.7  2  As  7.50E-06  -7.8E-07  -115.0  -3151.5  1  As  7.50E-06  -9.7E-07  -142.9  -3916.1  1  As  7.50E-06  -4.2E-07  -61.0  -1672.2  2  As  4.79E-06  -5.1E-07  -47.8  -1309.6  2  As  4.79E-06  -3.9E-07  -36.4  -998.3  1  As  4.79E-06  -8.3E-08  -7.8  -213.3  5  As  4.79E-06  -1.1E-07  -10.6  -289.3  2  As  4.79E-06  -5.2E-08  -4.9  -132.9  1  As  4.79E-06  -4.4E-08  -4.1  -112.8  4  As  4.79E-06  -4.2E-08  -4.0  -108.3  6  As  4.79E-06  -2.4E-07  -23.0  -629.9  1  As  6.08E-06  -1.8E-07  -21.9  -599.6  1  As  6.08E-06  -1.4E-07  -16.4  -449.9  mol/cnr  mmol/cm /y  4  2  mmol/m  2  NH4  1.64E-05  -8.8E-08  -0.03  2  NH4  1.64E-05  -1.3E-07  -0.04  -1.2  1  NH4  1.64E-05  -1.1E-07  -0.03  -0.9  -0.8  1  NH4  1.64E-05  -1.1E-07  -0.03  -0.9  2  NH4  1.10E-05  -2.9E-08  -0.01  -0.2  2  NH4  1.10E-05  -2.7E-08  -0.01  -0.2  1  NH4  1.10E-05  -3.2E-08  -0.01  -0.2  5  NH4  1.10E-05  -1.6E-08  0.00  -0.1  1  NH4  1.37E-05  -7.0E-08  -0.02  -0.5  1  NH4  1.37E-05  -6.3E-08  -0.02  -0.5  242  2  Appendix G. Accumulation Rates  The following section provides very rough approximations of accumulate rate for various sedimentary components in Balmer Lake. The accumulation rate at a sediment depth Z ( A R ) , can be defined by: Z  A R = Wt.% constituent x Dry Bulk Density (DBD) (g-cnr ) x Sedimentation 3  Z  Rate(cm-yr~l)  Where: i) D B D = (1- (j)) psediment particles -3  <> j = porosity a 0.87 assuming an average particle density of 2.65 g-cm ii) Sedimentation Rate approximated to be ~ 0.3 cm-yr'  1) Sulphur S ^ = 1.5 wt.% x 0.35 g.cm" x 0.3 cm y" = 2 x 10 3  1  3  gcm'V  2) Nickel Ni  = 0.4 wt.% x 0.35 g.cm x 0.3 cm y" = 4 x 10" kg cm" y 3  m  1  7  2  1  3)Gopper ;  C u ^ = 1 wt.% x 0.35 g.cm x 0.3 cm y" = 10 x 10" kg cm '~y 3  1  243  7  l  

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