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Groundwater-Surface Water Interactions: Implications for Nutrient Transport to Tropical Rivers Prachi Dixon-Jain May 2008 A thesis submitted for the degree of Doctor of Philosophy of The Australian National University
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Groundwater-Surface Water Interactions:
Implications for Nutrient Transport to Tropical Rivers
Prachi Dixon-Jain
May 2008
A thesis submitted for the degree of Doctor of Philosophy of
The Australian National University
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I certify that this thesis does not incorporate without acknowledgment any material previously
published. This work has not previously been submitted for a degree or diploma in any
institution of higher education.
Prachi Dixon-Jain
May 2008
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Acknowledgements
The journey of a PhD is filled with moments of excitement, frustration, uncertainty, satisfaction
and relief. I dearly thank my family and friends for their encouragement during all of these
phases. In particular, I thank my husband Stephen, as well as Mum, Dad and Jij for their faith in
my abilities and unfailing support to see me through to the end.
I would like to acknowledge my supervisors Professor Tony Jakeman, Dr Rebecca Letcher, Dr
Barry Croke (ANU), Dr Richard Cresswell (CSIRO Land and Water) and Dr John Sims (Bureau
of Rural Sciences) for their contributions to different aspects of my research. I’d particularly
like to thank Rebecca for her guidance, flexibility and encouragement. I am also grateful to
Barry for his technical expertise and willingness to act as a sounding board to work through
analytical problems. Support from Richard has been invaluable for guiding me through the
hydrochemical aspects of my research. I especially thank Richard for agreeing to provide
supervision at an advanced stage in the project. Thanks also to John for his encouragement to
pursue a PhD in the first place and supporting me with the appropriate arrangements at BRS to
make it a reality.
I acknowledge Ray Evans (Salient Solutions) for helpful discussions in the initial phases of my
research. I’d also like to thank Dirk Kirste and Bear McPhail (ANU) for sharing their
hydrochemical expertise. Furthermore, I am grateful to Ian White (ANU), Jon Olley (CSIRO
Land and Water) and Peter Baker (BRS) for their interest and enthusiasm in my project.
I am extremely grateful to John Spring and Grahaem Chiles (BRS) for their professional
technical assistance in the field. Thanks also to John Charles (QDNRW) for his assistance
during a reconnaissance of the study area. Laboratory analyses were undertaken by Aleksandra
Plazinska (BRS), John Pengelly (Murray Darling Freshwater Research Centre), staff at CSIRO
Land and Water Laboratories (South Australia) and ECOWISE Environmental, to whom I am
grateful. In particular I’d like to thank Fred Leaney and Megan LeFourner for their assistance
with radon sampling.
The invaluable technical advice of QDNRW staff, especially Ray McGowan, Ian Baker, Geoff
Pocock and Phil Kerr is gratefully acknowledged. For logistical support and supply of
subsidiary data I also thank Peter Martin and Glen Romano (Hinchinbrook Shire Council);
Andrew Wood and Tony Marino (CSR); Anna Forrest and Raymond De Lai (Herbert Resource
Information Centre); Anthony McLoughlin and Ed Stephens (QDNRW); David Post, Peter
Fitch and John Dighton (CSIRO Land and Water); Caroline Coppo (Herbert River Catchment
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Group), Rod Collins (National Parks); Greg Shannon (Bureau of Sugar Experiment Stations);
and Ron Kerkwyk (Herbert Cane Productivity Services).
This research would not have been possible without the kind support of members of the local
community in the catchment. I really appreciate their generous hospitality and willingness to
help out where they could. I’d especially like to acknowledge the landholders who assisted with
sampling and allowed access to their properties, including Margaret and Doug Matthews, Vince
Vitale, Tony Palmas, Michael and Lynn Cristaudo, John Gollogly, Ian Morley, John and Pam
Schmidt, and Trevor Pallanza. In addition, I extend my gratitude to Norm and Karyn Bliesner
for patiently searching through drilling records and advising me on access to key sampling sites.
I would like to acknowledge the support and friendship of the students that have accompanied
me at various stages of the PhD journey, particularly Karen Ivkovic, Beth Rickwood, Celina
Smith, Amir Sadoddin, Birte Schoettker, John Drewry, Wendy Welsh, Geoff Adams and Sue
Powell. Thanks to Karen and Wendy for their feedback on chapter 4. A special thanks to Karen,
who has not only been a colleague and a friend, but a valuable mentor.
Other technical support was provided by various staff at the ANU. I acknowledge Jason
Sharples for his assistance with manipulating the groundwater database, Paul Sjoberg for his
help with scanning cross-sections, Debbie Claridge and Clive Hilliker for assistance with
graphics, and Karl Nissen and Steve Leahy for IT support. I also thank Sue Kelo for her caring
nature and administrative support.
This research was jointly funded by the Australian National University and the Bureau of Rural
Sciences, to whom I extend my gratitude. I especially thank BRS for their generous funding
towards field work and for allowing flexible work arrangements.
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Abstract
The interaction between groundwater and surface water systems is a key component of the
hydrological cycle and an understanding of their connectivity is fundamental for sustainable
water resource management. Water is a vehicle for mobilising dissolved constituents, including
nutrients, between surface and subsurface waters and between terrestrial and marine systems.
Therefore, knowledge of surface-subsurface linkages is critical not only for water quantity
allocation, but also for water quality and its implications for ecosystem health. In particular,
ascertaining the significance of groundwater fluxes for river nitrogen budgets is an important
motivation for characterising river-groundwater connectivity. This overarching theme is
developed through the course of the thesis.
The marked seasonality of tropical river systems provides a unique opportunity to investigate
groundwater contributions to surface waters, especially when there are minimal overland flows.
The Herbert River in northeast Queensland represents a useful case study in the Australian
tropics for assessing the potential for transport of agricultural contaminants, such as dissolved
forms of nitrogen, between surface and subsurface waters, and between terrestrial and marine
systems, including the ecologically significant Great Barrier Reef World Heritage Area. Whilst
the lower Herbert River catchment, dominated by sugarcane production, is the focus for this
thesis, the research methodology and policy implications for nutrient monitoring and
management are applicable to other tropical catchments.
An extensive water quality sampling program was instigated to collect river and groundwater
samples during low flow conditions, for analysis of a range of conservative and non-
conservative environmental tracers including major ions, stable isotopes of water, radon, and
dissolved inorganic forms of nitrogen. Grab samples were collected during months representing
the beginning and end of the dry season to compare connectivity relationships at contrasting
stages of the stream hydrograph. Hydrochemical data at the end of the dry season is particularly
useful for isolating the groundwater signal in the river and its tributaries. Existing physical and
chemical datasets are also an important source of high temporal resolution information to
supplement the more detailed water quality data collected specifically for this investigation.
An understanding of the dynamics of water movement between river and aquifer storages is
critical for assessing the mobility of dissolved nitrogen between them. A combination of
hydrogeological, hydrometric, hydrological and hydrochemical tools are applied to characterise
the interaction between the alluvial aquifers and the lower Herbert River at a catchment scale.
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Specifically, the potential for hydraulic connection and the direction of flux between the aquifer
system and the river are evaluated through qualitative hydrometric approaches, including: depth
relationships of the river channel with that of the underlying alluvial sediments; historical
groundwater elevation-stream stage relationships; and groundwater flow patterns around the
river. Hydrological techniques such as stream hydrograph and flow duration curve analysis are
utilised to assess the temporal characteristics of flow in the river; the groundwater flux to the
river is also quantified by hydrograph separation. Physical understanding of river-aquifer
linkages is verified and enriched through analysis of surface water chemistry data, in
conjunction with the conceptual hydrogeological model developed from physical and chemical
assessment of the aquifers. The significance of groundwater as a vector for nitrogen is then
evaluated in light of a conceptual process understanding of the river-aquifer system. This
provides a platform for undertaking future catchment-scale nutrient budget studies based on
detailed investigations of nitrogen sources and transformations.
The research approach used in this thesis highlights the value of combining analytical
techniques, not provided by any one method, to inform and verify different aspects of a complex
water resource problem involving both surface and groundwater systems. The application of
multiple environmental tracers, at varied spatial and temporal resolution, is particularly
instructive for distinguishing between the key processes that influence the chemistry of the river
in space and time. Furthermore, the spectrum of tracer techniques provides both qualitative and
quantitative information regarding the flux of groundwater along the length of the lower Herbert
River. Whilst the absolute groundwater fluxes determined have a degree of uncertainty, mass
balances of radon and selected solutes highlight the value of quantitative estimates in
combination with qualitative trends to characterise river-aquifer relationships. The analyses
demonstrate that discharge of groundwater from the alluvial aquifers is a dominant influence on
both the flow and chemistry of the lower Herbert River in the dry season. In particular,
groundwater is a key vector for the delivery of nitrate to the river during low flow conditions.
This provides a new perspective for monitoring and management of nutrients in tropical rivers
where there is good connectivity with the underlying groundwater system.
Key recommendations arising from this research include: (1) water quality sampling should be
undertaken at recognised periods on the stream/groundwater hydrograph, with an understanding
of temporal and spatial river-aquifer connectivity relationships; (2) surface and subsurface
sources of water and dissolved nutrients must be considered, including identification of nutrient
hotpots in both surface water and groundwater systems; (3) sampling locations should capture
the longitudinal variation in river nutrient concentrations, not simply end-of-river monitoring;
(4) appropriate water quality guideline values must be set to account for seasonal changes in
both the sources and forms of nutrients transported to surface waters.
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Table of Contents
ACKNOWLEDGEMENTS ....................................................................................................... v
ABSTRACT............................................................................................................................... vii
TABLE OF CONTENTS .......................................................................................................... ix
LIST OF FIGURES .................................................................................................................. xv
LIST OF TABLES ................................................................................................................. xxiii
CHAPTER 1 RESEARCH CONTEXT .................................................................................... 1
1.1 Introduction ............................................................................................................... 1
1.2 Integrated Catchment Management ........................................................................ 2
1.3 Water Reform in Australia....................................................................................... 4
1.3.1 Water quality management...................................................................................... 4
1.3.2 Groundwater policy and conjunctive management ................................................. 4
1.3.3 National Water Initiative......................................................................................... 5
1.4 Motivation .................................................................................................................. 5
1.4.1 Nitrogen in surface water and groundwater ............................................................ 5
1.4.2 Tropical Australia and nutrient delivery ................................................................. 6
1.5 Scope of the Thesis..................................................................................................... 8
1.6 Thesis Outline ............................................................................................................ 9
CHAPTER 2 CONNECTED WATER RESOURCES .......................................................... 11
2.1 Introduction ............................................................................................................. 11
2.2 Mechanisms of Interaction ..................................................................................... 11
2.2.1 Physical interactions.............................................................................................. 12
2.2.2 Chemical interactions............................................................................................ 15
2.3 Framework for a Nutrient Budget ......................................................................... 18
2.4 Assessment Methods................................................................................................ 20
2.4.1 Characterising river-groundwater interactions...................................................... 20
2.4.1.1 Hydrogeological approaches ........................................................................ 20
2.4.1.2 Hydrometric approaches............................................................................... 21
2.4.1.3 Hydrological approaches .............................................................................. 21
2.4.1.4 Hydrochemical approaches .......................................................................... 22
2.4.1.5 GIS-based approaches .................................................................................. 22
2.4.1.6 Modelling approaches .................................................................................. 23
2.4.2 Characterising nutrient mobility in water.............................................................. 24
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2.4.2.1 Field-based studies........................................................................................24
2.4.2.2 Nutrient modelling ........................................................................................25
2.4.3 Implications for the research approach..................................................................26
2.5 Chapter Summary....................................................................................................27
CHAPTER 3 RIVER-AQUIFER INTERACTIONS IN THE WET TROPICS .................29
3.1 Introduction..............................................................................................................29
3.1.1 Climate...................................................................................................................29
3.1.2 Features of streamflow and groundwater...............................................................30
3.1.3 Water quality and nutrients....................................................................................31
3.2 Selection of Case Study Area ..................................................................................33
3.2.1 Catchment water quality issues..............................................................................34
3.2.1.1 Previous N studies on surface water .............................................................35
3.2.1.2 Previous N studies on subsurface water........................................................37
3.3 Catchment Characteristics......................................................................................38
3.3.1 Climate...................................................................................................................41
3.3.2 Rivers and aquifers ................................................................................................42
3.4 Research Approach..................................................................................................43
3.5 Data Collection .........................................................................................................44
3.5.1 Existing data and applicability...............................................................................44
3.5.2 Sample types for this study....................................................................................46
3.5.3 Site selection..........................................................................................................47
3.5.4 Timing of sampling................................................................................................48
3.5.5 Logistics, materials and methods...........................................................................50
3.5.5.1 Sampling technique.......................................................................................50
3.5.5.2 Sample preparation and preservation ............................................................50
3.5.5.3 Analytical techniques....................................................................................51
3.6 Chapter Summary....................................................................................................51
CHAPTER 4 HYDROGEOLOGICAL FRAMEWORK ......................................................53
4.1 Introduction..............................................................................................................53
4.1.1 Key concepts and definitions.................................................................................53
4.2 Geologic Characterisation .......................................................................................55
4.2.1 General depositional environment .........................................................................55
4.2.2 Lithostratigraphic interpretation ............................................................................56
4.2.2.1 Relationships with the river ..........................................................................61
4.2.3 Boundary of the study area ....................................................................................61
4.3 Hydraulic Properties of the Aquifers .....................................................................62
4.3.1 Summary of Cox’s interpretation ..........................................................................62
4.3.2 Interpretation based on lithostratigraphy ...............................................................63
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4.4 Vertical Flow in the Subsurface ............................................................................. 63
4.4.1 Data preparation .................................................................................................... 64
4.4.2 Inter-aquifer connectivity...................................................................................... 67
4.4.3 Rainfall response................................................................................................... 74
4.4.4 Classification of the aquifers................................................................................. 76
4.5 Lateral Flow in the Subsurface .............................................................................. 76
4.5.1 Deep aquifer system.............................................................................................. 77
4.5.1.1 Flow pattern.................................................................................................. 77
4.5.1.2 Recharge-discharge characteristics............................................................... 77
4.5.2 Shallow aquifer system ......................................................................................... 81
4.5.2.1 Flow pattern.................................................................................................. 81
4.5.2.2 Recharge-discharge characteristics............................................................... 81
4.5.3 Relationships between aquifers ............................................................................. 85
4.6 Chapter Summary ................................................................................................... 87
CHAPTER 5 HYDROGEOCHEMICAL FRAMEWORK .................................................. 89
5.1 Introduction ............................................................................................................. 89
5.1.1 General principles ................................................................................................. 90
5.1.1.1 Environmental tracers................................................................................... 90
5.1.1.2 Ion chemistry ................................................................................................ 90
5.1.1.3 Isotope chemistry.......................................................................................... 92
5.1.1.4 Nitrogen chemistry ....................................................................................... 94
5.1.2 Methods of interpretation ...................................................................................... 95
5.2 Hydrochemical Patterns.......................................................................................... 96
5.2.1 Compositional groups ........................................................................................... 98
5.2.2 Linear trends........................................................................................................ 103
5.2.2.1 Deep aquifer ............................................................................................... 104
5.2.2.2 Shallow aquifer........................................................................................... 107
5.2.3 Spatial trends....................................................................................................... 111
5.2.3.1 Deep aquifer ............................................................................................... 111
5.2.3.2 Shallow aquifer........................................................................................... 113
5.3 Vertical Relationships Between Aquifers ............................................................ 115
5.3.1 Unconfined and confined systems....................................................................... 115
5.3.2 Saturation Indices................................................................................................ 116
5.3.3 Intra-aquifer relationships ................................................................................... 117
5.3.4 Inter-aquifer mixing trends.................................................................................. 119
5.3.5 Relationship with the bedrock............................................................................. 126
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5.4 Spatial Relationships Within Aquifers.................................................................126
5.4.1 Lateral hydrochemical evolution .........................................................................127
5.4.1.1 Deep aquifer................................................................................................127
5.4.1.2 Shallow aquifer ...........................................................................................130
5.4.2 Seawater intrusion ...............................................................................................132
5.5 Nitrogen in Groundwater ......................................................................................135
5.5.1 Spatial distribution of N.......................................................................................136
5.5.1.1 Shallow aquifer ...........................................................................................136
5.5.1.2 Deep aquifer................................................................................................137
5.5.2 Speciation of N ....................................................................................................139
5.5.2.1 Shallow aquifer ...........................................................................................139
5.5.2.2 Deep aquifer................................................................................................142
5.5.3 Nitrogen transport ................................................................................................144
5.5.3.1 Shallow aquifer ...........................................................................................144
5.5.3.2 Deep aquifer................................................................................................145
5.6 Chapter Summary..................................................................................................147
CHAPTER 6 PHYSICAL RIVER-GROUNDWATER INTERACTIONS .......................149
6.1 Introduction............................................................................................................149
6.1.1 Data availability and preparation.........................................................................150
6.2 Potential for Hydraulic Connection .....................................................................154
6.3 Direction of Interaction .........................................................................................156
6.3.1 Groundwater elevation - river stage relationships ...............................................156
6.3.1.1 Reach A: river mouth to gauge 116001 ......................................................157
6.3.1.2 Reach B: gauge 116001 to Stone River junction ........................................159
6.3.1.3 Reach C: upstream of Stone River junction to Long Pocket.......................160
6.3.1.4 Reach D: Long Pocket to Abergowrie ........................................................161
6.3.1.5 Interpretation of flux direction ....................................................................162
6.3.2 Flow characteristics .............................................................................................164
6.3.2.1 Stream hydrographs ....................................................................................164
6.3.2.2 Flow duration ..............................................................................................166
6.3.2.3 Hydrograph separation................................................................................167
6.4 Implications for N Transport................................................................................170
6.5 Chapter Summary..................................................................................................173
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CHAPTER 7 CHEMICAL RIVER-GROUNDWATER INTERACTIONS..................... 175
7.1 Introduction ........................................................................................................... 175
7.1.1 Factors that influence river chemistry ................................................................. 175
7.1.1.1 Climatic factors .......................................................................................... 177
7.1.1.2 Geologic factors.......................................................................................... 178
7.1.1.3 Biogeochemical factors .............................................................................. 178
7.1.1.4 Mixing of waters......................................................................................... 179
7.1.2 Assessment principles and methods.................................................................... 179
7.1.3 Study site & terminology .................................................................................... 181
7.2 General Hydrochemistry ...................................................................................... 183
7.2.1 Compositional groups ......................................................................................... 184
7.2.2 Stable isotopes..................................................................................................... 185
7.3 Temporal Data ....................................................................................................... 186
7.3.1 Field parameters .................................................................................................. 187
7.3.1.1 Electrical conductivity................................................................................ 188
7.3.1.2 pH ............................................................................................................... 191
7.3.1.3 Temperature................................................................................................ 192
7.3.2 Major ions ........................................................................................................... 193
7.3.2.1 Inter-seasonal trends................................................................................... 193
7.3.2.2 Intra-seasonal trends................................................................................... 196
7.3.2.3 Tributaries of the Herbert River ................................................................. 197
7.3.3 Nitrogen............................................................................................................... 199
7.3.3.1 Particulate vs dissolved N .......................................................................... 199
7.3.3.2 Dissolved organic vs inorganic N............................................................... 201
7.3.3.3 Inorganic species ........................................................................................ 202
7.4 Longitudinal Data.................................................................................................. 202
7.4.1 Salinity ................................................................................................................ 203
7.4.2 Major ions in the tidal zone................................................................................. 205
7.4.3 Processes in the freshwater zone......................................................................... 207
7.4.3.1 Evaporation................................................................................................. 207
7.4.3.2 Overland flow............................................................................................. 209
7.4.3.3 Tributary inflow.......................................................................................... 210
7.4.3.4 Groundwater discharge............................................................................... 212
7.5 Tracing Groundwater ........................................................................................... 214
7.5.1 Radon distribution in groundwater...................................................................... 215
7.5.2 Temporal trends in radon along the river ............................................................ 215
7.5.3 Relative flux of groundwater along the river ...................................................... 217
7.5.3.1 Comparison with hydrochemistry .............................................................. 220
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7.5.4 Uncertainty in groundwater flux..........................................................................222
7.5.4.1 Solute mass balance ....................................................................................222
7.5.4.2 Local baseflow contribution........................................................................224
7.6 Transport of Nitrogen by Groundwater ..............................................................228
7.6.1 Longitudinal trends in DIN..................................................................................229
7.6.1.1 Comparisons with other tracers...................................................................230
7.6.1.2 Comparisons within the dry season.............................................................233
7.6.2 Environmental significance .................................................................................234
7.7 Chapter Summary..................................................................................................237
CHAPTER 8 RESEARCH CONCLUSIONS .......................................................................239
8.1 Introduction............................................................................................................239
8.2 Key Findings...........................................................................................................240
8.2.1 Hydrogeological framework ................................................................................240
8.2.2 River-groundwater interactions ...........................................................................241
8.2.3 The significance of groundwater for river N budgets..........................................242
8.3 Research Approach................................................................................................245
8.3.1 Data collection .....................................................................................................245
8.3.2 Characterising river-aquifer interactions .............................................................246
8.3.3 Characterising nutrient mobility ..........................................................................248
8.4 Implications for the Use of Nutrient Budgets and Models .................................249
8.5 Monitoring and Management Implications for the Tropics...............................250
8.6 Further Research ...................................................................................................251
8.7 Concluding Statement............................................................................................253
REFERENCES ........................................................................................................................255
APPENDIX A LABORATORY ANALYSES AND FIELD DATA....................................271
APPENDIX B RADON SAMPLING PROCEDURE ..........................................................273
APPENDIX C UNCERTAINTY ANALYSIS.......................................................................275
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List of Figures
Figure 1-1 Framework for Integrated Catchment Management ................................................... 2
Figure 1-2 Framework for Integrated Catchment Management illustrating the focus in this study
on surface-groundwater interactions and water resource quality.................................................. 3
Figure 2-1 The basic types of river-groundwater interactions: (a) gaining stream, (b) losing
stream, (c) disconnected stream, and (d) bank storage (after Winter et al., 1998). .................... 13
Figure 2-2 Groundwater and stream channel interactions: (a) parallel-flow and (b) flow-through
dominated streams (after Woessner, 2000, in REM 2002). ........................................................ 14
Figure 2-3 Processes that influence subsurface hydrochemistry (after Back et al. 1993, in
Herczeg and Edmunds 2000). ..................................................................................................... 16
Figure 2-4 The hyporheic zone as the interface between local and regional groundwater flow
systems and surface waters (Winter et al., 1998)........................................................................ 17
Figure 2-5 A conceptual representation of a nitrogen budget..................................................... 19
Figure 2-6 Spatial connections of individual subsystems represented in Figure 2-5.................. 19
Figure 3-1 Location of the Herbert River catchment showing the Herbert River and its major
tributaries .................................................................................................................................... 34
Figure 3-2 Land cover in the lower Herbert River catchment (adapted from Bramley and Muller
1999) ........................................................................................................................................... 39
Figure 3-3 1:250,000 geological map comprising the lower Herbert River catchment (BMR,
1965) ........................................................................................................................................... 40
Figure 3-4 Mean monthly rainfall in the lower Herbert River catchment .................................. 41
Figure 3-5 Cumulative residual rainfall at station 32045 in the lower catchment ...................... 42
Figure 3-6 Daily rainfall versus the cumulative deviation of residual rainfall from the mean at
station 32045............................................................................................................................... 42
Figure 3-7 Location of surface water, groundwater, and rainfall collection sites during the three
sampling periods. ........................................................................................................................ 48
Figure 3-8 The beginning of each sample collection period in 2004-2005, in relation to stream
discharge and rainfall in the lower catchment ............................................................................ 49
Figure 4-1 Bedrock contours in the Herbert River delta............................................................. 56
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Figure 4-2 Cross-sections constructed from lithological logs of bores in the study area ............57
Figure 4-3 Representative lithologic cross-sections in the lower Herbert River catchment........59
Figure 4-4 Fence diagram for the alluvial aquifer system in the lower Herbert River catchment
.....................................................................................................................................................60
Figure 4-5 Digital Elevation Model (DEM) for the lower Herbert River catchment. .................65
Figure 4-6 Historical groundwater elevations for bores sampled automatically and manually. .66
Figure 4-7 Summary of the degree of vertical hydraulic connectivity (strong, good, poor) and
direction of head gradient between the shallowest and deepest aquifers ....................................67
Figure 4-8 Hydrographs for nested monitoring bores screened in the deepest (A-pipe) and
shallowest (B-pipe) water-bearing units in the upper half of the catchment ...............................68
Figure 4-9 Hydrographs for nested monitoring bores screened in the deepest (A-pipe) and
shallowest (B-pipe) water-bearing units in the lower half of the catchment ...............................69
Figure 4-10 Similarity in hydrographic response for selected nested bores displaying an
upwards potential.........................................................................................................................71
Figure 4-11 Hydrographs for nested monitoring bores screened in an intervening sand unit as
well as in the deepest and/or shallowest water-bearing units ......................................................73
Figure 4-12 Cross-correlation of groundwater levels against rainfall at gauge 32091................74
Figure 4-13 Cross-correlation of groundwater levels against stage height at gauge 116001
(shown as a 14-day moving average) ..........................................................................................75
Figure 4-14 Potentiometric surfaces for the deep aquifer system during (a) the dry season of
1976 and (b) consecutive wet season of 1977 .............................................................................78
Figure 4-15 Head difference contours for the deep aquifer system, comparing groundwater
levels during the 1977 wet season with previous dry season levels in November 1976 .............80
Figure 4-16 Watertable maps for the shallow aquifer system during (a) the dry season of 1976
and (b) consecutive wet season of 1977 ......................................................................................82
Figure 4-17 Head difference contours for the shallow aquifer system comparing groundwater
levels during the 1977 wet season with previous dry season levels in November 1976 .............84
Figure 4-18 Contours of groundwater elevation superimposed for the shallow and deep aquifers
during (a) the dry season and (b) the wet season.........................................................................86
Figure 5-1 The nitrogen cycle, modified after Pidwirny (2005) .................................................94
Figure 5-2 Piper diagram for deep and shallow aquifer samples collected during three sampling
periods: May 2004, October 2004, June 2005.............................................................................97
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Figure 5-3 Oxygen-18 and deuterium stable isotope data for deep and shallow aquifer samples
and a coastal rainfall event in May 2004 .................................................................................... 98
Figure 5-4 Piper diagram for (a) HSs and (b) HSd groundwater samples displayed by month of
collection..................................................................................................................................... 99
Figure 5-5 Schoeller plots illustrating the two dominant hydrochemical groups observed in
shallow aquifer samples in June 2005....................................................................................... 100
Figure 5-6 Piper diagram for shallow aquifer samples in June 2005........................................ 101
Figure 5-7 Schoeller plots illustrating the two dominant hydrochemical groups observed in deep
aquifer samples in June 2005.................................................................................................... 102
Figure 5-8 Piper diagram for deep aquifer samples in June 2005............................................. 103
Figure 5-9 TDS vs Cl for all deep and shallow aquifer samples .............................................. 104
Figure 5-10 Bivariate plots of major ions against Cl for deep groundwater samples collected
during the three sampling periods in 2004-2005....................................................................... 105
Figure 5-11 Oxygen-18 and deuterium stable isotope data for HSd samples and a coastal rainfall
event in May 2004 .................................................................................................................... 107
Figure 5-12 Bivariate plots of major ions against Cl for shallow groundwater samples collected
during the three sampling periods in 2004-2005....................................................................... 108
Figure 5-13 Bivariate plot of Br/Cl against Cl for shallow aquifer samples collected during the
three sampling periods in 2004-2005........................................................................................ 109
Figure 5-14 Oxygen-18 and deuterium stable isotope data for HSs samples and a coastal rainfall
event in May 2004 .................................................................................................................... 110
Figure 5-15 Pie charts illustrating the spatial distribution of (a) major anions and (b) major
cations in HSd in June 2005 as a percentage of total meq/L ..................................................... 112
Figure 5-16 Monovalent ions plotted against divalent ions, normalised to Cl for all HSd bores
and seawater.............................................................................................................................. 113
Figure 5-17 Pie charts illustrating the spatial distribution of (a) major anions and (b) major
cations in HSs in June 2005 as a percentage of total meq/L ..................................................... 114
Figure 5-18 Field measurements for shallow and deep aquifer samples collected in May and
October 2004............................................................................................................................. 117
Figure 5-19 Schoeller plots for two screened intervals within the same aquifer based on
available random measurements (1975 - 2005) ........................................................................ 118
Figure 5-20 Plots of saturation indices (logarithmic form) for intervals screened within (a) the
deep aquifer and (b) the shallow aquifer................................................................................... 119
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Figure 5-21 Schoeller plots for groundwater samples from shallow (blue) and deep (black)
nested bores in June 2005..........................................................................................................122
Figure 5-22 Plots of saturation indices (logarithmic form) for selected nested bores based on
samples collected in October 2004............................................................................................124
Figure 5-23 (a) Hydrochemical facies and (b) distribution of total dissolved solids determined at
bores in the deep aquifer in June 2005. .....................................................................................128
Figure 5-24 Stable isotopic values along a flowpath for the deep aquifer based on samples
collected in May 2004. ..............................................................................................................129
Figure 5-25 (a) Hydrochemical facies and (b) distribution of total dissolved solids (b)
determined at bores in the shallow aquifer in June 2005...........................................................131
Figure 5-26 Transects for estimating the theoretical saltwater wedge below freshwater according
to the Ghyben-Herzberg relation ...............................................................................................132
Figure 5-27 Theoretical seawater interface during two periods in 2004 and slotted depths of
bores in HSs and HSd as a function of distance from the coast along three transects................134
Figure 5-28 Distribution of mapped soil types in the study area (Wood et al., 2003) ..............136
Figure 5-29 Spatial distribution of dissolved inorganic nitrogen in bores screened in the (a)
shallow and (b) deep aquifers in June 2005 ..............................................................................138
Figure 5-30 Spatial distribution of (a) NO3- and (b) NH4
+ in bores screened in the shallow
aquifer in June 2005, with selected flowlines depicted .............................................................140
Figure 5-31 Bivariate plots for shallow groundwater samples (2004 samples only), displayed
according to compositional groups............................................................................................141
Figure 5-32 Spatial distribution of (a) NO3- and (b) NH4
+ in bores screened in the deep aquifer
in June 2005...............................................................................................................................143
Figure 5-33 Conceptual diagram summarising the movement of water and N in the alluvial
aquifer system and potentially to the Herbert River..................................................................146
Figure 6-1 Selected QDNRW stream gauges along the lower (116006, 116001) and upper
(116004) Herbert River .............................................................................................................150
Figure 6-2 (a) Surveyed riverbed and (b) estimated river width (June 2005) as a function of
distance from the mouth of the Herbert River ...........................................................................152
Figure 6-3 Derived historical stage height in the Herbert River at Trebonne ...........................153
Figure 6-4 Surveyed topographic features of the lower Herbert River channel ........................154
Figure 6-5 Map showing geographical features which relate to the analyses of groundwater
elevations and river topography/stage along the lower Herbert River ......................................155
xix
Figure 6-6 Comparison of time series groundwater elevation at bore 1160048 (reach A) with the
corresponding surveyed streambed and bank height in the Herbert River. .............................. 156
Figure 6-7 Comparison of historical groundwater elevations in selected (a) shallow bores and
(b) deep bores with corresponding river stage (adjusted) along reach A.................................. 158
Figure 6-8 Comparison of historical groundwater elevations in selected (a) shallow bores and
(b) deep bores with corresponding river stage heights (adjusted) along reach B ..................... 159
Figure 6-9 Comparison of groundwater elevation in shallow (11600068B) and deep
(11600068A) bores with river stage (adjusted) at Lannercost. ................................................. 160
Figure 6-10 Comparison of groundwater elevation at bore 11600070 and stage height at
adjacent gauge 116006.............................................................................................................. 161
Figure 6-11 Daily flow at the two lower Herbert River stream gauges during 1995-2005 ...... 165
Figure 6-12 Flow duration curves for stream gauges in the lower (116001, 116006) and upper
(116004) Herbert River catchment (refer to Figure 6-1 for gauge locations) ........................... 167
Figure 6-13 Baseflow separation using the Lyne-Hollick algorithm at the lower Herbert River
stream gauges............................................................................................................................ 169
Figure 6-14 Time series of calculated baseflow at stream gauges 116001 and 116006 and
observed total flow at gauge 116001 ........................................................................................ 171
Figure 6-15 Calculated daily baseflow as a percentage of observed streamflow at gauge 116001
.................................................................................................................................................. 172
Figure 7-1 Gibbs diagram depicting the key processes that control the chemistry of surface
waters. ....................................................................................................................................... 176
Figure 7-2 Surface water sampling sites during the three collection periods and locations of
rainfall samples ......................................................................................................................... 182
Figure 7-3 Piper diagram for surface water and groundwater samples collected during three
sampling periods: May 2004, October 2004, June 2005........................................................... 183
Figure 7-4 Modified Gibbs diagram with logarithmic plot of TDI against Na/(Na+Ca) for water
samples collected along the lower Herbert River ..................................................................... 184
Figure 7-5 Schoeller plots of (a) samples in the Herbert River and (b) Ca-Mg enriched
groundwaters of the shallow aquifer ......................................................................................... 185
Figure 7-6 Oxygen-18 and deuterium stable isotope data for surface water and groundwater
samples and a coastal rainfall event in May 2004..................................................................... 186
Figure 7-7 Selected QDNRW stream gauges along the lower (116006, 116001) and upper
(116004) Herbert River............................................................................................................. 187
xx
Figure 7-8 Time series water quality data relative to flow at gauge 116001.............................188
Figure 7-9 Electrical conductivity versus flow at gauge 116001 ..............................................188
Figure 7-10 Continuous daily electrical conductivity at gauge 116001 relative to streamflow and
rainfall .......................................................................................................................................189
Figure 7-11 Electrical conductivity versus flow at gauge 116001, depicting theoretical
evaporation lines starting from different flow/EC combinations ..............................................190
Figure 7-12 Continuous daily pH at gauge 116001 relative to streamflow ...191
Figure 7-13 Time series of mean air temperature, mean river temperature and streamflow at
gauge 116001 in the lower Herbert River .................................................................................192
Figure 7-14 Correlation between mean air and river temperatures at gauge 116001 in the lower
Herbert River .............................................................................................................................193
Figure 7-15 Piper plot of historical major ion compositions (1973-2004) of the Herbert River at
three stream gauges ...................................................................................................................194
Figure 7-16 Schoeller plots of historical (a) wet season and (b) dry season water quality samples
collected at gauge 116001 since the 1970’s ..............................................................................194
Figure 7-17 HCO3 and Cl concentrations against streamflow at gauge 116001 during months of
the wet and dry seasons .............................................................................................................195
Figure 7-18 Major ion chemistry for samples collected in the lower Herbert River during the
beginning (May 2004, June 2005) and end (October 2004) of the dry season..........................196
Figure 7-19 Schoeller plots for tributaries of the lower Herbert River during the beginning (May
2004, June 2005) and end (October 2004) of the dry season ....................................................198
Figure 7-20 Saturation indices (logarithmic form) for bores and tributaries in the upper part of
the catchment based on samples collected at the end of the dry season (October 2004). .........199
Figure 7-21 Time series concentrations of (a) total, (b) particulate and (c) soluble N at various
sites along the lower Herbert River ...........................................................................................200
Figure 7-22 Concentrations of dissolved N as inorganic and organic components at (a) upstream
and (b) downstream sampling sites along the lower Herbert River...........................................201
Figure 7-23 Nitrate and ammonium concentrations at (a) upstream and (b) downstream
sampling sites along the lower Herbert River ...........................................................................202
Figure 7-24 Measurements of field electrical conductivity along the lower Herbert River and its
tributaries during the beginning (June) and end (October) of the dry season............................204
Figure 7-25 Longitudinal comparison of major ion and oxide concentrations along the lower
Herbert River during months representing the beginning and end of the dry season................205
xxi
Figure 7-26 Mg/Ca ratio for samples collected along the lower Herbert River during October
2004 and June 2005 .................................................................................................................. 206
Figure 7-27 Salinity, Na/Cl and oxygen-18 (δ18O) along the lower Herbert River .................. 208
Figure 7-28 Longitudinal trends for selected major ions along the lower Herbert River in June
2005 .......................................................................................................................................... 209
Figure 7-29 Radon activities along the freshwater reaches of the lower Herbert River and
sampled tributaries in October 2004 ......................................................................................... 210
Figure 7-30 Longitudinal plots for selected ions along the freshwater reaches of the lower
Herbert River and sampled tributaries in October 2004 ........................................................... 211
Figure 7-31 Longitudinal trends for selected major ions along the lower Herbert River in
October 2004............................................................................................................................. 213
Figure 7-32 Concentration of radon along the lower Herbert River during periods representing
the end (October 2004) and beginning (June 2005) of the dry season...................................... 216
Figure 7-33 Measured radon concentrations and estimated groundwater flux along the lower
Herbert River in October 2004 and at selected reaches in June 2005....................................... 219
Figure 7-34 Ions ratios, estimated groundwater flux and radon along the lower Herbert River in
October 2004............................................................................................................................. 221
Figure 7-35 Speciation of DIN along the lower Herbert River during months representing the
beginning (May) and end (October) of the dry season ............................................................. 229
Figure 7-36 Comparison of NO3- with ion ratios and radon along the lower Herbert River in
October 2004............................................................................................................................. 231
Figure 7-37 Longitudinal plots of NO3- during months representing the beginning (May, June)
and end (October) of the dry season. ........................................................................................ 233
Figure 7-38 Conceptual diagram summarising the movement of water and N between the
aquifers and the lower Herbert River at the end of the dry season ........................................... 236
Figure 8-1 Conceptual diagram summarising the movement of water and N in the alluvial
aquifer system and potentially to the Herbert River ................................................................. 243
Figure 8-2 Conceptual diagram summarising the movement of water and N between the aquifers
and the lower Herbert River at the end of the dry season ......................................................... 244
xxiii
List of Tables
Table 3–1 Available datasets for the lower Herbert River catchment ........................................ 45
Table 3–2 Summary of laboratory and field measurements ....................................................... 46
Table 4–1 Summary of hydraulic characteristics of the water-bearing alluvial stratigraphic units
in the Herbert River valley, after Cox (1979) ............................................................................. 62
Table 5–1 Hydrochemical groups of the alluvial aquifers ........................................................ 110
Table 6–1 Classification system for stream-aquifer interactions relevant to conjunctive use
management (REM, 2002)........................................................................................................ 149
Table 6–2 Features of the QDNRW stream gauging stations in the lower catchment and selected
upper catchment gauges. ........................................................................................................... 151
Table 6–3 Comparison between gauged and derived stage heights on the same day of
measurement during the dry season. ......................................................................................... 152
Table 6–4 Bores and comparison stream gauges at sites along the four river reaches ............. 157
Table 6–5 Comparison of derived river stage at Timrith and Lannercost and actual stage heights
at gauging stations 116001 and 116006 in the lower Herbert River. ........................................ 160
Table 6–6 Summary of the dominant river-groundwater elevation relationships along the lower
Herbert River during wet and dry seasons in the historical record. .......................................... 162
Table 7–1 Distinctive chemical characteristics of processes that influence the chemistry of
surface waters............................................................................................................................ 181
Table 7–2 Required and estimated flow rates in tributaries of the lower Herbert River and
groundwater based on changes in total solute loads ................................................................. 212
Table 7–3 Measured input parameters for modelling radon activities in the river ................... 218
Table 7–4 Best estimates of the percentage of local baseflow that contributes to streamflow
along reach 2 during months in the dry season, based on a solute mass balance. .................... 224
Table 7–5 Calculated values for xi, Δxi and the fraction of error (equation 7-14) during the
beginning (May) and end (October) of the dry season in 2004 ................................................ 227
Table 7–6 Required groundwater discharge to account for observed NO3- concentrations in the
Herbert River in October 2004.................................................................................................. 232
Table 7–7 Concentration range for species of DIN for all samples collected in the lower Herbert
River during selected months of the dry season ....................................................................... 235
1
Chapter 1 Research Context
1.1 INTRODUCTION
Water is a key natural resource that is required for the environment and to support the nation’s
social and economic structures. Sustainable water resources can be defined as ‘those designed
and managed to fully contribute to the objectives of society, now and in the future, while
maintaining their ecological, environmental and hydrological integrity’ (Loucks and Gladwell,
1999). The continued availability of water is open to change, not only through natural variation,
but also through human impacts on the water cycle (ANZECC and ARMCANZ, 1994).
Therefore, the way in which we use and manage water today can influence the availability and
quality of water for all users of the resource in the future.
The notion of water management is concerned with maintaining water quality and quantity in
both surface and sub-surface reservoirs. In many parts of the landscape, sub-surface water
(including groundwater) and surface water (streams, lakes, dams, wetlands and estuaries)
interact such that the quality and quantity of one affects the quality and quantity of the other
(Winter et al., 1998). For example, surface water features can gain water and solutes from
groundwater systems, while in other situations surface waters can be a source of groundwater
recharge and cause changes in groundwater quality. In addition, pumping of groundwater can
affect surface water volumes, or conversely, withdrawal of water from streams can deplete
groundwater reserves. Therefore, double accounting of water resources, based on independent
allocations of surface waters and groundwaters, can have serious consequences for future water
availability.
Water management has traditionally focused on surface water and groundwater as if they were
separate entities rather than connected resources. This can largely be attributed to factors such
as: differing timescales of surface water and groundwater movement in the landscape; lack of
data and methodology; disciplinary differences between surface and groundwater hydrologists;
and legislation. However, increased understanding of the importance of the linkages between
groundwater and surface water is directing more water managers throughout the world to adopt
conjunctive management principles (REM, 2002).
Chapter 1
2
A new awareness of the linkages that exist between groundwater and surface water has been
brought about because of:
• the increased demand for water, resulting in the need to maximise the water available
for use;
• the realisation of the ecological importance of groundwater and the need to consider its
environmental values; and
• the obvious disastrous effects that an ‘out of balance’ system can have in terms of
drawdown, salinity and water quality (REM, 2002).
Although recent developments in water policy in Australia have begun to reflect this increased
awareness, the scientific research and knowledge to underpin national policy, and hence inform
effective management practices, is still in relative infancy. Whilst this thesis is primarily
concerned with the scientific aspects of surface-groundwater interactions, the implications of the
research findings for water policy are also considered.
1.2 INTEGRATED CATCHMENT MANAGEMENT
Integrated Catchment Management (ICM) involves a systems approach to managing a
catchment. It is directed towards achieving a balance between natural environmental values,
economic activity and social concerns. The connections between each element in an ICM
problem can be represented within a framework such as that depicted in Figure 1-1.
POLICY
CLIMATE
LAND USE
WATER RESOURCES
OUTCOMES
LAND MANAGEMENT
PRACTICE
POLICY
CLIMATE
LAND USE
WATER RESOURCES
OUTCOMES
LAND MANAGEMENT
PRACTICE
Figure 1-1 Framework for Integrated Catchment Management
Research Context
3
Considering land use as the start of the ICM loop, then the type of activity carried out on a
parcel of land will govern the land management practices undertaken. For example, in an area of
cropping activity, a landholder has particular management practices with regard to fertiliser
regime: the type, timing and amount applied to their crop. Due to the interconnectedness of land
use and water, land management practices such as crop fertilisation or irrigation will necessarily
have impacts on the quality and quantity of water resources. Impacts on water resources lead to
outcomes which have biophysical, ecological, economic and social implications (Ticehurst et
al., 2007; Letcher et al., 2005). Within such a system, change can be brought about through
voluntary measures or through levers such as policy intervention, at catchment, state or national
level. Such policies for example, can direct how land is used, set rules for how land should be
managed or establish thresholds for the quality and quantity of water in a catchment. Through
the notion of adaptive management, policies can be revised as experience reveals where
amendments are required. An external driver to the framework loop is climate, an overarching
factor in the system, but over which there is no direct control.
The focus of this research is only on one aspect within the framework of ICM: water resource
quality and the interaction between groundwater and surface water (Figure 1-2). Although the
thesis does not deal specifically with the complexities of ICM, the cause and effect relationships
related to water quality are considered within the broader ICM framework.
POLICY
CLIMATE
LAND USE
WATER RESOURCES
OUTCOMES
LAND MANAGEMENT
PRACTICE
fertilisertillagechannelling
croppinggrazingforestryhorticulturenational park
rainfalltemperatureseasons
biophysicalecologicaleconomicsocial
qualityquantity
adaptivemanagement surface water
groundwater
e.g. COAG
POLICY
CLIMATE
LAND USE
WATER RESOURCES
OUTCOMES
LAND MANAGEMENT
PRACTICE
fertilisertillagechannelling
croppinggrazingforestryhorticulturenational park
rainfalltemperatureseasons
biophysicalecologicaleconomicsocial
qualityquantity
adaptivemanagement surface water
groundwater
e.g. COAG
Figure 1-2 Framework for Integrated Catchment Management illustrating the focus in this study on surface-groundwater interactions and water resource quality in an agricultural-based region.
Chapter 1
4
1.3 WATER REFORM IN AUSTRALIA
Australia is the driest inhabited landmass and has the world’s highest rainfall variability
(McMahon et al., 1992); thus, there is an inherent limitation to the water resource-base. Water is
crucial to Australia’s natural and economic wealth. Inefficient and inappropriate water use has
created widespread degradation of Australia’s natural resources, such as a decline in the quality
and quantity of surface and groundwater systems, increased land salinisation, a contraction in
wetlands and diminishing populations of native fish, flora and fauna.
In February 1994, all State and Territory governments in Australia agreed that the management
and regulation of Australia’s water resources required significant policy and institutional
change, as it was recognised that inefficient and inappropriate water use was responsible for
widespread degradation of Australia’s natural resources. This agreement resulted in the
endorsement of a national policy by the Council of Australian Governments (COAG), known as
the COAG Water Reform Framework, to achieve an efficient, economically viable and
environmentally sustainable urban and rural water industry. The framework recognised the
unique characteristics of Australia’s water resources and their economic, social and
environmental importance for Australia’s wellbeing.
1.3.1 Water quality management
An important component of water reform is the management of water quality. In 1992 a
National Water Quality Management Strategy (NWQMS) was introduced by the
Commonwealth, States and Territories and was subsequently included in the COAG Water
Reform Framework (ARMCANZ and ANZECC, 1994). The main policy objective of the
NWQMS is based on the philosophy of ecologically sustainable development: to achieve
sustainable use of the nation’s water resources by protecting and enhancing their quality while
maintaining economic and social development. Guidelines for groundwater quality protection
were included as a module of the NWQMS (ARMCANZ and ANZECC, 1995).
1.3.2 Groundwater policy and conjunctive management
Although the management of groundwater was included in the provisions of the 1994 reforms,
the framework had a focus on surface water resources, and was not explicit about which aspects
applied to surface water and groundwater. It was not until 1996 that the issues surrounding
groundwater management were formally accepted into the reform framework by COAG, and
thus introduced into national water policy. Recommendation 3 drew attention to the problems of
managing interlinked groundwater and surface water resources through different programs, and
recommended that ‘groundwater and surface water resource management should be better
integrated’ (ARMCANZ, 1996). According to the policy paper by the Framework’s Taskforce:
Research Context
5
‘In many situations, groundwater and surface water are interconnected and interchangeable
resources where decisions made in one area affect the other’ (ARMCANZ, 1996). In addition,
Resolution 6 stated that: ‘Progress has been made in some limited areas in meeting
environmental needs of groundwater, but further progress has been constrained by a poor
understanding of the location, extent and processes associated with groundwater/surface water
interactions and associated ecosystems’ (ARMCANZ, 1998).
1.3.3 National Water Initiative
In 2003, COAG agreed to refresh the 1994 water reform agenda and develop a National Water
Initiative (NWI) in order to increase the productivity and efficiency of water use, sustain rural
and urban communities, and to ensure the health of river and groundwater systems. A National
Water Commission was established in 2004 to assess progress in implementing the NWI and to
advise on actions required to better realise the objectives of the Agreement. In relation to
conjunctive water management, one of the key objectives of the intergovernmental agreements
on a NWI is to achieve ‘recognition of the connectivity between surface and groundwater
resources and for connected systems to be managed as a single resource’ (COAG, 2004).
1.4 MOTIVATION
In light of the above policy framework, the interaction between groundwater and surface water
is nationally regarded to be an important water resource management consideration. This thesis
is concerned with the quality aspect of connected water resources and particularly, how the
interaction with groundwater affects in-stream water quality. In regards to water quality, river-
aquifer interactions are important in situations where a stream can be degraded by discharge of
saline or other low-quality groundwater, or where groundwater is polluted by a contaminated
stream. Pollution of surface and groundwater resources by nutrients is of growing concern
worldwide due to potentially harmful effects on both ecosystem and human health. The
movement of water between groundwater and surface water provides a major pathway for
chemical transfer between terrestrial and aquatic systems (Winter et al., 1998).
1.4.1 Nitrogen in surface water and groundwater
Nitrogen (N) inputs in intensive-agricultural catchments have been identified as a major causal
factor in the trends of increased nutrient concentrations in surface, ground and coastal waters
(Heathwaite et al., 2000). In recognition, European directives have been introduced to reduce
water pollution caused or induced by nitrates from agricultural sources and to prevent further
pollution (e.g. Statutory Instrument 1289, 2006; Council Directive 91/676/EEC, 1991). Diffuse
sources of N in agriculture are generally derived from fertilisers and manure. Imbalances in
nutrient budgets, such as for N, have shifted in scale from local to regional and continental
Chapter 1
6
dimensions during the last decade (Oenema et al., 2003). Nitrogen in surface and groundwater
systems is a concern for both human and environmental health. Elevated levels of nitrate in
drinking water can be harmful to humans, especially infants (Lawrence, 1983), as well as
livestock, while nitrogen inputs into surface waters can cause problems of eutrophication.
Nitrogen can be exported from agricultural areas in many forms and through various pathways,
including leaching losses to groundwater. Unlike salt, which is transported conservatively
throughout the landscape, nitrogen has a complex cycle because of the various forms it can take,
depending on the hydrochemical environment. Nutrient mobilisation from non-point sources
depends on the coincidence of source and transport controls (Heathwaite et al., 2000). Land
management factors largely control the magnitude of potential N loss (source control), while the
rate of N loss through leaching (transport control) depends on soil properties and the amount of
water percolating through the soil profile. Nitrogen loss from agricultural land generally occurs
at the catchment scale (Heathwaite et al., 1989). This is especially the case under oxidising
conditions, where nitrate is the dominant species of N, as it is soluble and is hence transported
with the movement of water. The high mobility of N as nitrate in leaching water means that a
significant proportion of the nitrate created by source factors is translated into N loss
(Heathwaite et al., 2000).
Nutrient export from catchments is a concern for Australian river systems for a number of
reasons, one of which is the risk of algal bloom formations. It is now recognised that in addition
to exports from surface runoff, groundwater can be a significant source of nitrogen to
freshwater, estuarine and coastal environments (Lamontagne et al., 2003; Lamontagne et al.,
2002; Linderfelt and Turner, 2001; Smith and Turner, 2001). Therefore, understanding river-
aquifer connectivity is an important aspect of nutrient management in landscapes where
nutrient-bearing surface waters and groundwaters interact.
1.4.2 Tropical Australia and nutrient delivery
Terrestrial runoff of pollutants has led to concerns over the threat to tropical Australian
waterways, wetlands and coastal ecosystems, such as the Great Barrier Reef World Heritage
Area (GBRWHA) (Brodie and Mitchell, 2005; Furnas, 2003). Water borne contaminants have
the potential to be delivered to the GBRWHA via direct farm runoff into streams and rivers;
leaching into groundwater systems with subsequent discharge into rivers, wetlands and near
shore marine waters; and submarine groundwater discharge via paleochannels (Stewart et al.,
2005; Stieglitz, 2005). Runoff incorporating sediment, nutrients and pesticides is increasing,
with loads for most pollutants being many times higher than the natural amount discharged e.g.
150 years ago (Brodie and Mitchell, 2005). The principal land uses in northern Australia
contributing to this pollution are cropping and cattle grazing, with lesser contributions from
industrial areas, mining and urban developments. Furthermore, Brodie and Mitchell (2005) note
Research Context
7
that due to agricultural development, nutrient inputs (of N and P) have changed from
dominantly dissolved organic forms in ‘natural’ systems to dissolved inorganic forms that are
more bioavailable than organic species.
Although remnant forest exists in parts of northern Australian, most catchments are now used
for agricultural or pastoral purposes. Sugarcane is the dominant cultivated crop in northern
Australia, harvested primarily on the coastal plain south of the Daintree, around the Atherton
Tablelands in Queensland, and in the Ord River irrigation area of Western Australia (Brodie and
Mitchell, 2005). Off-site impacts due to nitrogen and other nutrient/chemical losses from
sugarcane production have come under increasing public scrutiny, particularly when they occur
adjacent to sensitive wetlands, marine, and estuarine environments, such as the Great Barrier
Reef (Bohl et al., 2001). It is well documented that high concentrations of N, especially
inorganic nitrate, are found particularly in surface waters draining land under sugarcane (Bartley
et al., 2003; Furnas, 2003; Bramley and Roth, 2002; Bramley and Muller, 1999). In addition to
surface waters, groundwater studies have shown there to be subsurface losses of nitrate beneath
sugarcane growing areas (Bohl et al., 2000b; Weier, 1999). Despite the recognition of nitrate
leaching into both surface and subsurface waters, the role and contribution of nitrate in
groundwater on surface water quality is not well documented.
High nutrient loadings, especially of N and P, can disrupt the nutrient balance and cause
widespread eutrophication in both in-stream and near-shore surface waters. Coral reefs are
particularly susceptible to damage from increased nutrient discharge from terrestrial sources;
reef communities degraded by terrestrial inputs are documented globally (Fabricius, 2005). Due
to the ecological significance of the GBRWHA, this study will focus on one of the thirty-four
GBR coastal catchments, the Herbert River catchment in north Queensland, to explore linkages
between groundwater and surface water systems. The Herbert is an important river system of
the wet-dry tropics because it drains directly into the area of the GBR lagoon considered under
greatest threat from terrestrial runoff (Productivity Commission, 2003). Sugarcane production is
widespread in the catchment; therefore, terrestrial activities can potentially impact on the quality
of water both within the river system and discharged to the marine environment. Further
characteristics of the case study catchment are deferred to Chapter 3.
In general, Australia’s tropical river systems are poorly understood in comparison with
Australia’s temperate freshwater and tropical marine systems (Hamilton and Gehrke, 2005).
Although these tropical catchments are considered to have greater freshwater biodiversity and
aquatic ecosystem health than found in temperate Australia, eutrophication of estuaries and
coastal waters has been documented and is attributed primarily to increased nutrient loading
from rivers. Based on a review of tropical research in Australia, Hamilton and Gehrke (2005)
noted that hydrological linkages between groundwater and surface waters are inadequately
Chapter 1
8
understood in tropical Australia. Therefore, this research aims to contribute to the knowledge
base of river-aquifer interactions in tropical river systems of Australia, including the
significance of groundwater for river nitrogen budgets. Whilst there has been a greater
geographic focus of research centred on catchments draining into the GBR than elsewhere in the
Australian tropics, it is anticipated that the research methodology, implications for monitoring,
and management considerations presented in this thesis, will have greater application than to the
GBR coastal catchments alone.
1.5 SCOPE OF THE THESIS
The overarching aim of the thesis is to characterise the interaction between surface water and
aquifer systems and thereby consider the potential implications of these interactions for river
nitrogen budgets. Emphasis is on investigating surface-subsurface linkages rather than
understanding nutrient sources, processes and loads. Water is a vehicle for mobilising and
transporting nutrients and other dissolved constituents; therefore, knowledge of the dynamics of
water movement, including connectivity between surface and subsurface reservoirs, is central to
understanding the transport of such components. In order to assess large-scale river-
groundwater interactions and the application to nutrient transport, a catchment scale
investigation is undertaken. This scale is considered appropriate for the management of
catchments, rather than individual river-reaches, which also has relevance for water policy. In
addition, potential transferability of the thesis findings to other regions is enhanced by
conducting a broad scale assessment. The selected case study catchment, located in the tropical
climate zone of northeastern Australia, provides a unique opportunity to study baseflow
conditions (largely groundwater contributions) with minimal overland flows. Given the focus on
quality aspects rather than quantity, an extensive water quality sampling program was instigated
in the case study catchment during the dry season, as described in Chapter 3. A range of
hydrogeological, hydrological and hydrochemical analytical tools are applied to existing and
new data in order to address three key questions:
(1) What is the nature of river-aquifer interactions in the Herbert River catchment,
particularly during the dry season?
(2) What is the significance of river-aquifer interactions for the nitrogen budget of the river?
(3) What are the implications of these interactions for nutrient monitoring, management,
and policies relating to water quality at a catchment scale?
Whilst this research primarily focuses on (1) and (2), recommendations for management (3) are
raised in the concluding chapter of the thesis.
Research Context
9
1.6 THESIS OUTLINE
This thesis has been structured into eight chapters. A summary of each chapter is outlined
below:
Chapter 1: Research Context
This chapter provides a brief policy context for this research, outlining the most significant
historical developments in water reform in Australia, and highlights the need to integrate surface
and groundwater management as it relates to water quality. Pollution of connected water
resources by nutrients arising from agricultural activities is discussed in the context of northern
Australian catchments, with particular attention drawn to the significance of the Great Barrier
Reef and terrestrial activities in adjacent catchment areas. This provides a motivation for the
chosen case study catchment and scope of the thesis.
Chapter 2: Connected Water Resources
This chapter reviews current scientific understanding of physical and chemical interactions
between surface and groundwaters in the landscape and summarises methodologies for
investigating interaction processes. Approaches for characterising nutrient transport and fluxes
are also described. The merits of understanding physical processes such as river-aquifer
linkages as an aid to interpreting model outputs are also discussed. This provides a framework
for the methodology employed in this research.
Chapter 3: River-Aquifer Interactions in the Wet Tropics
This chapter discusses the key distinguishing features of tropical climates in relation to
hydrological processes and water quality. In addition, the chapter introduces the characteristics
of the case study catchment and summarises previous water quality studies relating to nitrogen.
The general research approach of the thesis and available datasets are also outlined. Additional
data requirements are highlighted: these underpin the extensive water quality sampling program
undertaken in the catchment. Data collection and details of the sampling methodology are also
described.
Chapter 4: Hydrogeological framework
This chapter is the first of two chapters focused on characterising the hydrogeology in the case
study area. Existing bore logs are interpreted and physical datasets are analysed to conceptualise
the hydrogeology of the alluvial aquifer system.
Chapter 1
10
Chapter 5: Hydrogeochemical framework
This chapter provides a validation and extension of the physical hydrogeological framework,
based primarily on analysis of hydrogeochemical data collected as part of this study. The spatial
distribution and speciation of nitrogen in groundwater are also examined in order to establish
whether there is the potential for nitrogen in groundwater to contribute to the river system.
Chapter 6: Physical River-Groundwater Interactions
This chapter considers physical hydraulic relationships between groundwater and surface water,
in order to assess potential for hydraulic connection and the nature of stream-aquifer interaction
processes along the river. Based on the interaction characteristics, implications for the transport
of nitrogen from groundwater to surface waters are also considered.
Chapter 7: Chemical River-Groundwater Interactions
In this chapter, analyses of an extensive database of hydrochemical data provide a powerful tool
to characterise the hydrochemistry of the river during the wet and dry seasons. In particular,
river-aquifer relationships established in the previous chapter are verified and enhanced through
the application of qualitative and quantitative techniques. The potential for a groundwater
source of nitrate and the environmental significance for in-stream and marine ecosystems are
also raised.
Chapter 8: Conclusions
The final chapter provides a summary of conclusions of the thesis, including the key research
contributions. In light of the conceptual understanding developed throughout the thesis on river-
aquifer interactions, the chapter discusses implications and recommendations for nutrient
monitoring, management, and water policy, particularly in tropical catchments. Suggested
future research is also outlined.
11
Chapter 2 Connected Water Resources
2.1 INTRODUCTION
Groundwaters are fed by rain and surface waters. Groundwaters ultimately discharge to
surface waters or the sea. Surface waters are fed by groundwaters, and feed groundwaters.
Surface and groundwaters form parts of one interlinked system (p. 98, Nevill et al., 2001).
Groundwater and surface water are not isolated components of the hydrological cycle; they
interact in a range of topographic, geologic and climatic landscapes (Sophocleous, 2002; Winter
et al., 1998). Effective management of water requires an understanding of the components of
the hydrological cycle as well as the linkages between those components. One important
connection that has traditionally been overlooked in water resource management in Australia is
the interaction between surface water and groundwater. Direct observation and measurement of
surface-subsurface relationships is complex, as it requires an understanding of processes that
occur over different timescales and how they interact both temporally and spatially. Whereas
water movement on the surface can be quite rapid in response to rainfall events, the movement
of water beneath the land surface can be slow and variable, hence more difficult to predict. As
outlined in Chapter 1, the focus of this thesis is on water quality aspects of river-groundwater
interactions and the significance of groundwater as a vector for nitrogen to the stream. This
chapter reviews current scientific understanding of mechanisms by which surface water and
groundwater interact in the landscape. A conceptual framework for a nitrogen budget is also
presented, as a context for assessing river-aquifer connectivity and the consequent mobilisation
of N. Furthermore, methods for investigating interaction processes and broad approaches for
characterising nutrient transport and fluxes are reviewed. This provides a framework for the
research approach of the thesis.
2.2 MECHANISMS OF INTERACTION
Comprehensive reviews by Sophocleous (2002), Woessner (2000), Winter (1999) and Winter et
al. (1998), outline the key developments in scientific understanding of groundwater-surface
water interactions based on theoretical and field-based studies. Recent research has focussed on
the ecological significance of groundwater-surface water interactions; specifically, in
understanding the biogeochemical processes that occur at the interface. Literature by Boulton et
al. (1998), Dahm et al. (1998) and Brunke and Gonser (1997) provide a review of this emerging
Chapter 2
12
research. The following sections describe the main principles and mechanisms by which surface
water and groundwater interact, with an emphasis on river-groundwater interaction processes.
Whilst the discussion assumes aquifers are simple porous media, it is noted that much of
Australia is underlain by fractured aquifers which are more difficult to characterise and model.
Therefore, many of the conceptual models are oversimplified.
2.2.1 Physical interactions
The movement of surface water and groundwater is largely controlled by the physiography
(topography and geologic framework) of an area, while the sources and losses of water in the
hydrological cycle are controlled by climate (Winter, 1999). These controls in turn influence the
flowpath along which groundwater moves, known as the groundwater flow system, and hence
determine the nature of the interaction with surface water features. Streams, lakes and wetlands
are integral parts of groundwater flow systems whereby fluxes of water and chemicals from and
to groundwater reflect the positions of surface water bodies with respect to different-scale
groundwater flow systems (Winter, 1999).
Groundwater moves along flowpaths of varying length and travel time, from areas of recharge
to areas of discharge. Based on their spatial extent and influence, Tóth (1963) identified three
distinct types of flow systems: local, intermediate and regional. In local flow systems water is
recharged at watertable highs and flows to nearby discharge areas such as streams. Groundwater
in regional flow systems travels a greater distance and often discharges to major rivers, large
lakes or oceans. As local flow systems are the most dynamic and the shallowest flow systems,
they have the greatest interchange with surface water (Winter et al., 1998). In general, areas of
high topographic relief tend to have dominant local flow systems, whereby groundwater moves
in systems of predictable patterns; conversely, areas of nearly flat relief tend to have dominant
intermediate and regional flow systems, in which the scale of groundwater movement is much
larger and less predictable. In nature, a region with irregular topography contains multiple flow
systems of differing sizes and depths that can be superimposed on one another within a
groundwater basin (Sophocleous, 2002).
Surface water features such as lakes, wetlands and streams can have multiple sources of water,
including direct precipitation, surface runoff and groundwater. Although groundwater
interactions with lakes and wetlands are similar to that of river systems, there are subtle
differences that will not be discussed here (refer to Winter, 1999; Winter et al., 1998). Streams
interact with groundwater by: (1) gaining inflow of water from groundwater through the
streambed (gaining stream, Figure 2-1a); (2) losing water to groundwater by outflow through
the streambed (losing stream, Figure 2-1b); or (3) variably gaining or losing water depending
on the season and/or physiographic conditions. A connected stream is in direct contact with the
underlying aquifer via a zone of saturated material, or is separated by a narrow unsaturated zone
Connected Water Resources
13
generally less than twice the stream width (Bouwer and Maddock, 1997). Where these
conditions are not met, disconnection from the groundwater system can result, with losses
occurring to seepage (Figure 2-1c). Bank storage is another type of river-groundwater
interaction whereby a rapid rise in river stage, higher than adjacent groundwater levels, causes
water to move from the river into the streambank (Figure 2-1d). The return of this water to the
stream can be on the order of a few days (in the case of local bank storage) or years (in response
to flooding and aquifer recharge) (Winter et al., 1998).
a
b
c
d
Figure 2-1 The basic types of river-groundwater interactions: (a) gaining stream, (b) losing stream, (c) disconnected stream, and (d) bank storage (after Winter et al., 1998). Arrows show the direction of water movement.
Chapter 2
14
Woessner (2000) additionally defines fluvial plain-groundwater interactions as: parallel-flow
dominated, where groundwater flows parallel to the stream but essentially does not discharge
into it (Figure 2-2a); and flow-through dominated, where a reach both gains and loses water as
groundwater flows through (Figure 2-2b).
a
b
Figure 2-2 Groundwater and stream channel interactions: (a) parallel-flow and (b) flow-through dominated streams. The watertable (between the lightest and intermediate shades), stream (darkest shade), direction of groundwater flow (arrows) and equipotential lines (dashed) are shown (after Woessner, 2000, in REM 2002).
River-aquifer interactions occur at a variety of scales. Whilst groundwater exchange with the
stream occurs by processes of discharge, recharge and flow-through at the fluvial plain scale,
hyporheic exchange occurs at the channel-bed scale, whereby local, shallow surface water
circulation in the underlying sediments creates areas of groundwater recharge and discharge
within river reaches that are characterised as either gaining or losing. Hyporheic exchange itself
occurs at many scales, ranging from centimetres to tens of metres, depending on sediment type,
bed geometry and hydraulic gradients in the adjacent groundwater system (Woessner, 2000;
Winter et al., 1998).
The direction of interaction in hydraulically connected systems depends on the relative
elevations of the watertable and stream-water surface. For instance, in order for groundwater to
discharge into a stream (gaining condition), the altitude of the watertable (groundwater head) in
the vicinity of the stream must be higher than that of the river stage; the converse applies for
surface water discharging to groundwater (losing condition). Therefore, parallel-flow occurs
when the channel stage and groundwater head are equal, while flow-through occurs when the
channel stage is less than the groundwater head on one bank and greater than the head at the
opposite bank (Woessner, 2000). Importantly, whilst the relative hydraulic heads influence the
direction of exchange, the actual river-groundwater flux depends on the hydraulic conductivity
of both the channel sediments and the aquifer (Sophocleous, 2002).
The interaction between surface water and groundwater can vary both spatially and temporally.
While in some environments the river-groundwater relationship may remain fairly constant, in
other environments flow direction may vary along the stream length (gaining and losing in
different reaches), and also change in response to rainfall events or seasonal climatic patterns
(gaining or losing in the same reach at different times). Water that enters a surface water body
Connected Water Resources
15
rapidly in response to a water input event such as rainfall is known as quick flow or event flow.
This water is distinguished from baseflow, or subsurface water (mostly groundwater flow), that
contributes to the stream and maintains streamflow between water input events. Discharge and
recharge of an aquifer has a buffering effect on the flow regimes of rivers (Brunke and Gonser,
1997). For instance, under low precipitation, discharge in many streams is attributed to
baseflow. In contrast, under conditions of high rainfall, increased surface runoff and interflow
lead to higher hydraulic pressures in the lower stream reaches, causing the river to change from
gaining to losing as it infiltrates its banks and recharges groundwater. Australian rivers are in
general fed by surface aquifers, not by overland flow, snowmelt, or direct rainfall. Except for
those areas in Australia with exceptionally high rainfall and surface runoff, the existence of
permanent standing water indicates substantial dependence on groundwater inflows (Nevill et
al., 2001).
In addition to environmental controls, anthropogenic factors such as groundwater pumping and
land use can have significant impacts on the movement of water between groundwater and
surface water resources. This raises an important issue for the management of water availability.
As this thesis is specifically concerned with water quality aspects of connected water resources,
issues associated with site-specific surface water or groundwater extraction will not be
discussed further. On the basis of the physical principles summarised above, the following
section examines factors that affect chemical river-aquifer interactions.
2.2.2 Chemical interactions
The chemistry of surface and subsurface waters is largely influenced by the geology and the
contact time that water has with the geological materials. Microorganisms in the soil, sediments
and water also affect the chemical characteristics of groundwater and surface water. In local
flow systems, groundwater has a relatively short time in contact with aquifer minerals, which
can result in minimal chemical changes prior to surface water discharge (Winter et al., 1998).
However, in the shallow environment groundwaters can be in their most aggressive state or be
exposed to a large store of salts; in addition, being closest to the land surface, shallow
groundwater is the most vulnerable to contamination from anthropogenic sources. Therefore,
the interaction between local flow systems and surface waters is important for water quality
management of connected resources. In contrast, water in deeper flow systems has longer
flowpaths and hence a longer contact time with subsurface materials for products of
geochemical weathering to accumulate. The longer residence time of deep groundwater can also
mean that the effects of point source or diffuse pollution to the groundwater system can take
longer to detect if the groundwater is ultimately discharged to a stream (and can persist for
longer after a change is made). Therefore, in connected systems where shallow and/or deep
Chapter 2
16
groundwaters are connected to a stream, the chemistry of receiving waters can be influenced by
groundwater or surface water having different histories and distinctive chemical signatures.
As water infiltrates through the unsaturated zone of the subsurface and into the saturated zone it
undergoes biogeochemical transformations. Although the most dramatic changes in
hydrochemistry occur in the soil and unsaturated zones, water quality in the saturated zone
progressively evolves along flowpaths towards areas of discharge, recording a chemical
signature that is often closely related to aquifer characteristics and residence time (Herczeg and
Edmunds, 2000). The main biogeochemical reactions that affect the transport of chemicals in
surface water and groundwater include: acid-base reactions; precipitation and dissolution of
minerals; sorption and ion exchange; oxidation-reduction (redox) reactions; biodegradation; and
dissolution and exsolution of gases (Winter et al., 1998). Figure 2-3 summarises the
hydrological cycle and the major processes that influence chemical evolution of water in the
subsurface as it progresses towards surface water features. Many of these processes also directly
influence surface water chemistry, although the relative influence of each process may differ
from that in the subsurface.
Figure 2-3 Processes that influence subsurface hydrochemistry (after Back et al. 1993, in Herczeg and Edmunds 2000).
The hyporheic zone, as defined by Winter et al. (1998), is the subsurface zone where stream
water flows through short segments of its adjacent bed and bank sediments, thereby creating a
mixing zone between subsurface and surface waters (Figure 2-4). Due to this mixing, the
chemical composition of the intervening hyporheic zone may differ significantly from that in
Connected Water Resources
17
both ground and surface waters. The hyporheic zone is typically a region of intensified
biogeochemical activity; these biogeochemical processes can in turn affect the movement of
nutrients and other chemical constituents, including contaminants, in both directions across the
streambed (Sophocleous, 2002; Winter et al., 1998). This has important implications for stream
health and ecological function. For example, microbes and algae attached to sediment in
hyporheic zones can uptake nutrients such as nitrate and hence lower the concentration of
dissolved nitrogen in the stream (Winter et al., 1998). Consequently, the hyporheic zone can be
viewed as a biological filter which can affect the chemistry of groundwater entering surface
water and also that of surface water entering groundwater.
Figure 2-4 The hyporheic zone as the interface between local and regional groundwater flow systems and surface waters (Winter et al., 1998).
The hyporheic zone is one example of an environment that can attenuate nitrogen
concentrations in groundwater. Lamontagne et al. (2001) note that there are two additional
hydrogeological environments: riparian zones and suboxic/anoxic aquifers, which have the
potential to remove a large fraction of nitrogen from polluted groundwater before it reaches
sensitive ecosystems. Issues surrounding the role of riparian zones in nitrogen attenuation have
been discussed by several authors (Rassam et al., 2006; Lamontagne et al., 2005; Malard et al.,
2002). Effective riparian buffers are generally characterised by a shallow impermeable
geological layer, which forces groundwater to move laterally through shallow root zones and
organic-rich deposits (Hill, 1996). Similarly, shallow unconfined aquifers with significant inputs
of surface-derived dissolved organic carbon, or aquifers containing organic matter/reactive
reduced material, can also attenuate nitrate from groundwater travelling through them
(Lamontagne et al., 2001).
The distinction between the three hydrogeological environments for nitrogen attenuation
discussed above is related to scale, ranging from whole of catchment (for reduced aquifers),
floodplain (for riparian zones), and river bed (for hyporheic processes) (Lamontagne et al.,
2001). However, as noted by Winter et al. (1998), an important characteristic of hyporheic
zones is that they represent an area ‘where groundwater that drains much of the subsurface of
landscapes interacts with surface water that drains much of the surface of landscapes’.
Chapter 2
18
Therefore, although hyporheic exchange can be viewed on the one hand as a small-scale
process, it also represents the interaction of surface and subsurface waters influenced by
catchment-scale processes. Woessner (2000) argues that stream-groundwater exchange should
be conceptualised and characterised more holistically at both the fluvial plain and channel-bed
scale, by multidisciplinary researchers, not hydrogeologists alone.
The significance of riparian and hyporheic zones for nitrogen attenuation is not well understood
in tropical and semi-arid river systems, as much of the research has taken place in temperate
regions, generally outside of Australia (Lamontagne et al., 2001). Although the focus of this
study is not specifically on riparian or hyporheic zones, or the complex hydrochemical reactions
at the groundwater-surface water interface, these environments are important controls on what
ultimately reaches the river system. Therefore, it is important to bear in mind the potential role
of these biological systems in river nutrient budgets.
2.3 FRAMEWORK FOR A NUTRIENT BUDGET
The overarching theme of this thesis is concerned with the implications of surface-groundwater
interactions for river nutrient budgets. A nutrient budget for nitrogen can be conceptualised as a
series of interconnected compartments or storages, each of which contributes and/or mobilises
N in its various forms to surface water. As illustrated in Figure 2-5, the key storages are:
atmosphere, rainfall, agriculture, soil, aquifer (shallow and deep), riparian/ hyporheic zone, and
surface water. Whilst point source inputs are not explicitly included, they are potential sources
of N to surface and subsurface storages. Upstream contributions also constitute an important
storage entity as an input into the system. The agriculture storage incorporates vegetation,
including crops; therefore, N can be added to the soil storage through applied fertilisers and/or
organic decay, or removed from the soil storage through uptake by crops and other vegetation.
Similarly, as depicted by the double arrows in Figure 2-5, there is the potential for N exchange
in both directions between other storages, such as between the aquifer storages and surface
water. For the schematic purposes of Figure 2-5, no distinction is drawn between dissolved
inorganic and organic forms of N, with both types represented as dissolved N. In addition, it is
assumed that particulate forms of N are not mobilised to groundwater.
The framework depicted in Figure 2-5 is a description of each subsystem, or node, at a
particular location along a river; the stretch of river between two nodes is referred to as a reach,
which has a corresponding contributing subcatchment area. The sum of each N contribution
(source term) arising from the various storages within a subsystem determines the net load of N
delivered to the downstream node of a particular river reach. The calculated N load for a reach
in turn provides the upstream input load for the adjacent downstream reach (except for the most
upstream node for which the input load is only from the upper catchment area). The river N
Connected Water Resources
19
budget is thus conceptualised as a series of subsystems that are linked spatially by the
contributing N load from the surface water storage in the adjacent upstream subsystem (Figure
2-6). The identification of source strength (of dissolved or particulate N), both within storages
of individual reaches and between river reaches, is important for prioritising management
interventions in a catchment. Within the budget framework the storages of agriculture and
riparian/ hyporheic zone (N attenuation) represent two controls that can be varied in order to
achieve different outcomes (N loads) in the river.
DEEP AQUIFER
UPSTREAM LOADS
SHALLOW AQUIFER
SOIL
AGRICULTURESURFACE
WATER
RAINFALLATMOSPHERE
RIPARIAN / HYPORHEIC
ZONE
dissolved N
particulate N
dissolved N
dissolved N
dissolved N
dissolved N
dissolved N
dissolved N
dissolved N particulate N
particulate N
dissolved N
gaseous N
dissolved N
particulate N
particulate N
particulate N
DEEP AQUIFER
UPSTREAM LOADS
SHALLOW AQUIFER
SOIL
AGRICULTURESURFACE
WATER
RAINFALLATMOSPHERE
RIPARIAN / HYPORHEIC
ZONE
dissolved N
particulate N
dissolved N
dissolved N
dissolved N
dissolved N
dissolved N
dissolved N
dissolved N particulate N
particulate N
dissolved N
gaseous N
dissolved N
particulate N
particulate N
particulate N
Figure 2-5 A conceptual representation of a nitrogen budget. Note that atmospheric losses of N are in the gaseous state.
Upper catchment
N load
DownstreamUpstream Nitrogen transport along the lower Herbert River
Upper catchment
N load
DownstreamUpstream Nitrogen transport along the lower Herbert River
Figure 2-6 Spatial connections of individual subsystems represented in Figure 2-5. Note that adjacent subsystems or nodes are connected by a river reach, which has a corresponding contributing subcatchment area.
Chapter 2
20
This thesis is concerned with the potential movement of N between the surface water and
aquifer storages. Therefore, the nitrogen budget framework provides a context for assessing
river-aquifer connectivity and the consequent transport of N.
2.4 ASSESSMENT METHODS
As outlined in Chapter 1, there are two main analytical components to the thesis: (1)
characterising river-aquifer interactions, and (2) ascertaining the role of groundwater in the
nitrogen budget of river systems. Accordingly, this section describes techniques for
investigating river-groundwater interactions and an overview of approaches used to characterise
nutrient mobility, particularly in surface and subsurface waters. This discussion forms the basis
for the research approach proposed in Chapter 3.
2.4.1 Characterising river-groundwater interactions
There are a variety of techniques for assessing the nature and degree of stream-aquifer
interaction processes. As summarised by Brodie et al. (2007), the range of methods include:
seepage measurement, field observations, ecological indicators, hydrogeological mapping,
geophysics and remote sensing, hydrographic analysis, hydrometric analysis, hydrochemistry
and environmental tracers, artificial tracers, temperature studies, water budgets, and modelling.
These methods can be described in the context of spatial scale, temporal scale, cost, ease of use,
advantages, limitations, and application. As this thesis is concerned with the catchment scale,
this section provides an overview of approaches that are applicable to regional scale studies.
However, it is noted that the findings of large scale studies have implications at the local scale,
such as for water quality protection and ecosystem health. The approaches are broadly grouped
as hydrogeological, hydrometric, hydrological, hydrochemical, GIS-based and modelling.
2.4.1.1 Hydrogeological approaches
Knowledge of the hydrogeology surrounding a surface water feature is critical to characterising
connectivity. This includes an understanding of the factors that control groundwater flow
(Section 2.2.1), such as aquifer geometry, host geology, stratigraphy, hydraulic properties,
geological structures and river morphology (Brodie et al., 2007). Standard techniques for
conceptualising the hydrogeology involve interpretation of borehole (lithological logs and water
levels) and pump test data using a variety of methods. These include construction of cross
sections and three dimensional stratigraphic maps; bore hydrograph analyses; and mapping of
groundwater level contours and flowlines (Domenico and Schwartz, 1990; Fetter, 1988; Heath,
1987; Freeze and Cherry, 1979). Limitations with hydrogeological approaches are that
compilation and analysis of the data can be time consuming and complex; limited borehole data
can also lead to misinterpretation (Brodie et al., 2007).
Connected Water Resources
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2.4.1.2 Hydrometric approaches
Hydrometric methods are based on Darcy’s Law for fluid movement in a porous medium, and
are thus concerned with the hydraulic gradient between groundwater and surface water systems
and the hydraulic conductivity of the intervening aquifer (Brodie et al., 2007). Darcy’s Law can
be applied to estimate the rate and direction of groundwater flux to or from a surface water body
by measuring the head difference between the stream level and groundwater level; the vertical
distance between the measuring point in the aquifer and the streambed; and the vertical
hydraulic conductivity of the material along the vertical flowpath between the groundwater
measuring point and the stream bed. However, the approach can be an oversimplification of the
groundwater flow conditions (Woessner, 2000), in that the flux between the stream and shallow
aquifer are assumed to be entirely vertical. Flownet analysis is another approach for estimating
groundwater seepage rates in the horizontal or vertical plane (Loaiciga and Zekster, 2002).
While this approach can provide a simple and cost-effective way to estimate seepage flux, the
method can not account for spatial variability and local groundwater factors. Quantitative
methods for estimating seepage fluxes also rely on reasonable estimates of the hydraulic
conductivity, which can vary along a flowpath (Brodie et al., 2007).
A qualitative approach for determining the hydraulic gradient is to compare stream and
groundwater levels and hence establish the potential direction of seepage. Furthermore,
comparison of stream and bore hydrographs can indicate temporal changes in the hydraulic
gradient and whether a reach is likely to be gaining or losing at different periods in time
(Section 2.2.1). Groundwater level contours in the vicinity of a stream or lake can also indicate
the relationship between an aquifer and surface water feature (Winter et al., 1998). Whilst these
qualitative methods provide a rapid assessment of potential seepage directions between an
aquifer and a river, their effectiveness relies on having stream level and groundwater level data
at close proximity.
2.4.1.3 Hydrological approaches
Historical streamflow data form the basis of hydrologic methods for characterising surface
water outflow from a catchment. Hydrograph separation techniques aim to separate the stream
hydrograph into slowflow and quickflow components, and thus isolate the low-frequency signal
of a stream. This is useful for characterising the magnitude and timing of groundwater discharge
to a gaining stream. Baseflow separation techniques include digital filtering (Furey and Gupta,
2001; Chapman, 1999; Nathan and McMahon, 1990; Lyne and Hollick, 1979), graphical
methods (McNamara et al., 1997; Linsley et al., 1958), and chemical separation approaches
(Laudon and Slaymaker, 1997; Hooper and Shoemaker, 1986). Strictly speaking, baseflow
separation techniques that are based purely on the observed streamflow are actually estimating
the slowflow component. Whether this is actually baseflow depends on the assumptions used
Chapter 2
22
and whether the filter is appropriate for the system being studied. In this thesis, the output from
the filtering techniques is assumed to be baseflow. Analyses of flow duration curves are another
common method for determining the significance of baseflow contributions to a stream
(Smakhtin, 2001). Whilst these hydrological approaches are useful desktop tools, they rely on
the availability of accurate stream gauging information, which may be of limited spatial extent.
In addition, they do not provide information on the spatial distribution of groundwater input
along the stream.
Water budgets such as river-reach water balances are an additional hydrological approach,
whereby estimates are made of the unaccounted difference in the water balance for a specified
river reach, commonly between two gauging stations (Bratten and Gates, 2003; REM, 2002).
Whilst this method allows the flow volume associated with stream-aquifer interactions to be
inferred, it relies on the accurate measurement of surface water flow, which can have
considerable measurement errors. In addition, the flow volume is a net measurement; therefore,
it does not account for all the other potential gains and losses for the reach (Brodie et al., 2007).
2.4.1.4 Hydrochemical approaches
Dissolved constituents in water have the potential to retain a ‘memory’ of the movement and
interactions of water through surface and subsurface storages. The application of environmental
tracer methods for characterising river-aquifer interactions is based on the principle that surface
waters and groundwaters may have different tracer contents, which allows differentiation of
their sources (REM, 2002). Furthermore, the analysis of groundwater chemistry data in
conjunction with groundwater hydraulic information is useful for building a conceptual model
of the hydrogeological system. There are a variety of tracer techniques to identify gaining and
losing-river reaches and to quantify the groundwater-surface water flux. Commonly used tracers
include field parameters, major ions, stable and radioactive isotopes, trace elements, and
industrial chemicals. Different tracers provide information on different processes; therefore, a
multi-tracer approach is commonly used (Cresswell and Herczeg, 2004; Thayalakumaran et al.,
2004; Cook et al., 2003; Lamontagne et al., 2003; Herczeg et al., 2001). The spatial and
temporal resolution, reliability, and degree of quantification may also vary between tracers;
thus, a combination of techniques provides a more robust process understanding. A major
drawback of hydrochemical approaches is the generally high cost associated with sampling
logistics and laboratory analyses.
2.4.1.5 GIS-based approaches
Geographic information systems (GIS) can be used as a mapping tool for spatial representation
of aspects of the hydrogeology, hydrology and hydrochemistry, and hence aid with the
conceptualisation of processes at a catchment scale. A more sophisticated application of a GIS
is to characterise groundwater-surface water connectivity by integrating geophysical and remote
Connected Water Resources
23
sensing imagery with other spatial datasets. Remote sensing surveys allow for rapid, non-
invasive mapping of landscape parameters that either indicate or control groundwater-surface
water interactions. In addition, they can provide good spatial resolution in the vicinity of surface
water features, including information on changes through time. However, undertaking and
interpreting these surveys can be complex, requiring expensive equipment, technical expertise
and logistical support (Brodie et al., 2007).
2.4.1.6 Modelling approaches
Merritt et al. (2003) review the three basic model types: empirical, conceptual and physically-
based. Empirical models are generally the simplest of the model types; they are primarily based
on the analysis of observations and seek to characterise response from these data.
Computational and data requirements for such models are usually less than for conceptual and
physics-based models, often being capable of being supported by coarse measurements. As a
result, this class of model tends to produce outputs at high spatial and temporal aggregation. In
comparison, conceptual or physics-based models tend to be more dynamic, capturing greater
spatio-temporal resolution. Conceptual models are typically based on the representation of a
catchment as a series of internal storages and thus include a general description of catchment
processes without including the specific details of process interactions. Whilst they tend to be
aggregated, the distinction between conceptual and empirical models is that they still reflect the
hypotheses about the processes governing system behaviour. Physics-based models are based on
the solution of fundamental physical equations and therefore can involve large numbers of
parameters. These models tend to have the greatest spatial and temporal resolution; however,
the physical significance of model outputs due to scale issues is questionable. In addition,
detailed model approaches (i.e. mechanistic, spatially distributed) require large efforts in model
development, calibration and validation for a range of conditions (Wolf et al., 2005). Despite
their relative simplicity, empirical models are frequently used in preference to more complex
models as they can be implemented in situations with limited data and parameter inputs and are
particularly useful as a first step in source identification (Merritt et al., 2003).
Numerical modelling approaches can provide a valuable tool for combining into a consistent
framework the information obtained from the approaches described above. Existing models that
simulate river-aquifer interactions fall into the broad categories of surface water models,
groundwater models, combined models, and fully integrated models (REM, 2002). Numerical
models are useful as a predictive tool for management and policy. However, oversimplified
models may not be adequately robust, while complex models can increase data requirements,
costs and time (Brodie et al., 2007).
Chapter 2
24
2.4.2 Characterising nutrient mobility in water
Approaches for characterising and predicting the movement of nutrients between surface water
and groundwater range from qualitative field-based studies to quantitative nutrient modelling.
With a particular focus on N as the nutrient of interest, this section provides a brief overview of
these contrasting approaches.
2.4.2.1 Field-based studies
A plethora of documented studies have focused on nutrient delivery to and export from surface
waters (Brodie and Mitchell, 2005; Eyre and Pont, 2003; Mitchell et al., 2001). Similarly, there
are numerous studies on nitrate leaching to groundwater, particularly in agricultural catchments
such as below sugarcane (Thorburn et al., 2003; Bohl et al., 2001; Rasiah and Armour, 2001;
Weier, 1999; Verburg et al., 1998; Lawrence, 1983). Other field-based studies consider nutrient
balances in the air-soil-water system of agricultural catchments to determine the efficiency of
fertiliser management practices (Meier et al., 2006; Thorburn, 1999; Bristow et al., 1998; Prove
et al., 1997; Freyney et al., 1994). More recently, there has been heightened research interest in
the interactions between surface and groundwater systems and the implications of this
connectivity for N movement, particularly from groundwater to rivers or coastal waters.
Investigations range from purely groundwater N studies that infer implications for transport to
surface waters (Ahern et al., 2006; Merrill and Benning, 2006; Rasiah et al., 2003) to more
comprehensive studies of N cycling between groundwater and adjacent surface waters during
different time periods (Lamontagne et al., 2005; Lamontagne et al., 2002). Characterising the
hydrogeochemical environment using tracers in addition to N, have also been the focus, or a
component of, studies concerned with N cycling (Grimaldi et al., 2004; Thayalakumaran et al.,
2004; Korom, 1992).
As discussed in Section 2.2.1, surface water-groundwater interactions occur at various spatial
scales (Woessner, 2000). Furthermore, key biogeochemical reactions involving N, such as
nitrate reduction, occur in distinct hydrogeological environments that differ in scale
(Lamontagne et al., 2001). Accordingly, studies concerning N transformations and movement
between surface and groundwaters have been represented as regional (catchment), floodplain
(riparian zone) or riverbed (hyporheic zone) scale investigations, or a combination. The scale of
investigation depends largely on the underlying research questions, for example, whether there
is an interest in small scale biogeochemical interactions at the surface-groundwater interface
(Hefting et al., 2006; Rassam et al., 2006; Young and Briggs, 2005; McKergow et al., 2004;
Lamontagne et al., 2003; Malard et al., 2002; Boulton et al., 1998; Cirmo and McDonnell,
1997), or whether large scale mechanisms such as discharge of regional groundwater and
nutrients to surface waters are relevant (Linderfelt and Turner, 2001). Exchanges between near-
channel and in-channel water are critical to stream restoration and riparian management efforts
Connected Water Resources
25
(Sophocleous, 2002). Hence, given the ecological implications of N cycling at the surface
water-groundwater interface, there has in general been greater research attention towards reach
scale rather than catchment scale investigations.
2.4.2.2 Nutrient modelling
Due to the capacity of a catchment to store N in soil, vegetation and groundwater, it may take
considerable time for the full impacts of anthropogenic activity to be reflected in river nitrate
quality. Therefore, models are considered essential for predicting how changes in deposition,
land use, management and climate will affect N loading to rivers (Whitehead et al., 1998).
Efforts to model diffuse nutrient losses within catchments are intensifying, with the types of
model used varying from simple lumped and GIS representations to fully distributed models
(Neal and Heathwaite, 2005). There is an extensive body of literature on different nutrient
models that have been used for a variety of applications. As reviewed by Wolf et al. (2005),
existing nutrient models differ with respect to: (1) modelling approach (e.g. mechanistic,
empirical, equilibrium); (2) input data aggregation (e.g. spatially distributed, lumped); (3)
calculation time step (e.g. hourly, yearly, average annual); and (4) spatial unit or scale of
modelling (e.g. plot, catchment). According to Wade et al (2005), ‘whilst purely empirical
approaches may fit the observed patterns, only dynamic, process-based models can also
represent the likely future changes in the driving factors, the catchment N stores and the
hydrological pathways, and therefore predict the change in the catchment hydrology and stream
water nitrate concentrations’.
For some studies, there may be a requirement for integrating multiple variables and modelling
scenarios (such as potential impacts) at different spatial and temporal scales. Detailed numerical
modelling may have a role in these situations. In contrast, nutrient budgets, or mass balance
approaches, summarise nutrient inputs and outputs of a defined system over a defined period of
time. Hence, budget approaches, which are an example of an empirical type of model, are often
static or produce outputs at large temporal scales such as annually. Budgets are a valuable tool
for summarising large amounts of information in transparent and easily understood input-output
diagrams and are flexible in that they allow for approximations and revisions where there are
gaps in data. However, there are various sources of uncertainty in the budget approach due to
biases (e.g. personal, sampling, measurement, data choice and manipulation) and errors related
to spatial and temporal variability (Oenema et al., 2003). The conceptual framework presented
in Figure 2-5 illustrates the input-output structure of a nutrient budget. Input data for nutrient
budgets can be classified according to: (1) type (primary, estimate or assumptions); (2) source
(field/ laboratory measurement, observation, computation); and (3) frequency (continuous,
seasonal, annually). Nutrient budgets are commonly based on a combination of data types which
are derived from various sources and are collected at different frequencies. Ultimately, the
Chapter 2
26
purpose of the study defines the budgeting approach, scale, data acquisition strategy and
required accuracy and precision of the nutrient budget (Oenema et al., 2003).
2.4.3 Implications for the research approach
Many of the techniques for characterising groundwater-surface water interactions (Section
2.4.1) are standard methods historically used by separate disciplines; however, it is the
combination of approaches applied to a connected water resource problem that provides a
powerful toolbox to qualitatively or quantitatively assess surface water-groundwater linkages.
Therefore, whilst there is no method, formula or computer model that can be universally
applied, different datasets and techniques can be explored at different stages in the assessment
to inform the various aspects of the investigation (REM, 2002). Factors to consider in the choice
of method(s) include: the water resource question and objectives; availability of data and other
resources; location and characteristics of the study area; assessment time; and scale/type of
output required. In light of the above considerations, a range of hydrogeological (Chapters 4 and
5), hydrometric and hydrological (Chapter 6), and hydrochemical (Chapters 5 and 7) techniques
are regarded as appropriate for the objectives of this study, with use of a GIS as a mapping tool.
The chosen techniques complement one another such that together they provide a more
complete picture of groundwater-surface water interactions, not provided by any one method.
Based on standard reference texts and applications in other studies, the general principles
underlying the methods employed in the thesis are outlined in the relevant analytical chapters.
As outlined in Section 2.4.2, field-based and modelling approaches represent the two broad
categories for investigating the mobility of nutrients in the environment. Nutrient modelling
approaches are concerned with how much N is ultimately delivered to the river; they do not
explicitly require a detailed understanding of the underlying hydrological or biogeochemical
processes for N gains/losses from source to sink. Whilst a nutrient budget, such as the
conceptual nitrogen budget (Figure 2-5 and Figure 2-6), is a simple approach for identifying
source strength within and between reaches, of interest to this study are the physical processes
that underpin the transport of N to surface waters, particularly in relation to river-aquifer
interactions. Due to the complex transformation reactions that are associated with N transport in
the environment, characterising the dynamics of water movement between river and aquifer
storages is a first step to understanding the mobility of dissolved N between them. Measured
concentrations of N in groundwater and along the river are hence interpreted in light of this
process understanding. Thus, a broad scale field-based approach is considered appropriate for
investigating the significance of groundwater N fluxes to the river, which is founded on an
understanding of the hydrology, hydrogeology, hydrometrics and hydrochemistry of the river-
aquifer system. This is the basis for the research approach, as outlined further in Chapter 3
(Section 3.4). Whilst it is considered that completely parameterising a nutrient budget or other
Connected Water Resources
27
models is outside the scope of this thesis, the interpretation of outputs from these approaches
can be enhanced with reference to a conceptual framework based on process understanding of
the system.
2.5 CHAPTER SUMMARY
Chapter 1 provided a context for the importance of considering integrated water resources,
whereby groundwater and surface water are managed as a single resource. Based on a review of
studies concerning groundwater-surface water interactions, it was also highlighted that the role
and significance of groundwater for river nitrogen budgets is not well understood. Therefore,
two key analytical components to this research were identified: (1) characterisation of river-
aquifer interactions and (2) assessment of the potential role of groundwater for the nitrogen
budget of river systems. This chapter introduced the key physical concepts surrounding
connected water resources and highlighted the implications of surface-groundwater interactions
for water quality. Assessment tools and methods available to characterise these interactions
were also examined: many of these techniques are utilised in this thesis. In addition, approaches
for understanding nutrient transport and fluxes were reviewed, which ranged from field-based
studies at various spatial scales to complex numerical modelling at different spatial and
temporal resolutions. Given the focus on process understanding and the extensive data
requirements which could not be met with existing data, a modelling approach was considered
beyond the scope of the thesis. However, it was noted that the interpretation of model outputs
can be improved with process knowledge of key components such as river-aquifer linkages.
Therefore, based on considerations such as the key research questions, data availability,
characteristics of the study area and the scale of the study, it was established that hydrochemical
techniques, combined with hydrogeological, hydrometric and hydrological methods, are
appropriate for characterising river-aquifer interactions at a catchment scale. The implications
for N cycling will hence be examined with regards to process understanding of how water
moves between surface and subsurface systems.
The following chapter reviews tropical hydrological processes and issues concerning water
quality. The case study area is presented for the research, with previous N studies outlined. The
broad research approach is described in the context of the overarching research questions of the
thesis. Based on the characteristics of the case study catchment and existing datasets, the
methodology for further data collection is also presented.
29
Chapter 3 River-Aquifer Interactions in the Wet Tropics
3.1 INTRODUCTION
Since the 1960’s there has been considerable progress in process hydrology research in the
temperate latitudes, while the humid tropics have received comparatively less attention. As a
result, findings from temperate regions have tended to be transferred to tropical environments
(Bonell and Balek, 1993). In the Australian context, hydrology studies have concentrated on
sub-humid and semi-arid climates and landscapes (Fleming, 1993). Due to world-wide concerns
regarding the impact of tropical forest clearance, international research specifically related to
hydrology in the tropics has generally focused on the environmental and human impacts of
deforestation and land cover change in large humid catchments of South East Asia, Africa and
South America (Bonell and Bruijnzeel, 2005; Bonell et al., 1993). In comparison, hydrological
studies in the Australian tropics have largely been undertaken in agricultural catchments that no
longer preserve large areas of native vegetation, where the issues addressed are largely
concerned with agricultural impacts on water resources (Rasiah et al., 2003; Thorburn et al.,
2003). Despite the different foci of overseas and Australian studies, the international literature
provides insight into key features that differentiate hydrological conditions in the tropics from
those in temperate climates. The review that follows draws on the available literature to shed
light on river-aquifer interactions and nutrient transport, specifically in tropical climates. This
subsequently leads to a discussion of the case study catchment in the wet/dry tropics of north
Queensland.
3.1.1 Climate
Geographically, the tropical regions are approximately bounded by the Tropics of Cancer
(23o27’N) and Capricorn (23o27’S). The humid tropic regions are defined as areas where the
mean temperature of the coldest month is above 18oC, the duration of the wet season exceeds
4½ months and a wet month has on average 1000 mm of rainfall (Chang and Lau, 1983). Under
this definition, approximately 12% of Australia can be classified as humid tropics, which
includes the wet-dry tropical sub-region of northern Australia (Fleming, 1993; Stewart, 1993).
The remainder of tropical Australia is considered to be in the dry tropics, where the duration of
the wet season is less than 4½ months. A great amount of solar energy in the tropics creates a
climate without harsh winters; therefore, solar energy influences the hydrological cycle more
directly in the tropics than in other climatic zones (Latrubesse et al., 2005). In tropical areas,
Chapter 3
30
rainfall is considered to be the main factor that determines the seasons and is also the most
variable element of tropical climate (Callaghan and Bonell, 2005; Latrubesse et al., 2005).
According to Bonell and Balek (1993), the magnitude of rainfall is one of the driving forces in
differentiating between the response of hydrological processes in the humid tropics compared
with processes in higher latitudes.
Rainfall in tropical northern Australia is distinctly seasonal. Due to the geographic location, the
Australian tropics are subject to monsoonal influences and tropical storms and cyclones
(Fleming, 1993). The highest rainfall areas are associated with the Great Dividing Range in the
east, exceeding 3200 mm annually along the wet tropical coast (Stewart, 1993).
3.1.2 Features of streamflow and groundwater
Tropical rivers, in general, show high but variable peak discharges during the rainy season and
periods of low flow when rainfall decreases. These rivers can be grouped into two main types:
(1) rivers with well defined high and low discharges corresponding to unimodal rainy periods;
and (2) rivers with two flood peaks per year, consistent with bimodal rainy periods. Tropical
rivers in Australia tend to have highly seasonal flow regimes, with much of their annual flow
between November to May (Hamilton and Gehrke, 2005). Individual rivers in northern
Australian catchments can have multiple major flows each year (such as the Tully), one major
annual flow (such as the Herbert) or major flows that are separated by several years (such as the
Burdekin). In both the wet and dry tropical catchments of Australia, the two distinct flow
conditions are relatively well separated in time, with only short periods dominated by high
flows compared to the otherwise dominant low flow conditions (Brodie and Mitchell, 2005).
Higher rainfall intensities in the tropics lead to differences in runoff generation (especially
during tropical storm events) compared to non-tropical areas (Bonell and Balek, 1993). Based
on research in northeast tropical Queensland, Bonell et al. (1998) found that ‘at low rainfall
intensities, precipitation is routed to streamflow via the groundwater/soil system. However, at
high precipitation rates, the flow capacity of these pathways is exceeded and alternative
pathways involving rapid flow are invoked which effectively short-circuit the former pathways’.
Hence, while vertical movement of soil water prevails between rain events and during most
small events, under heavy and prolonged rain the horizontal transfer of water develops to
produce lateral flow (Douglas and Guyot, 2005). This is in contrast to other climatic and
geographic regions where the event hydrograph consists of pre-event or ‘old’ water
(groundwater/soil water), rather than event or ‘new’ water from rainfall/overland flow. It is
suggested that any shift in the delicate balance between rainfall intensity, soil hydraulic
properties and topography can provide widely different runoff processes, both within and
between various tropical forests (Bonell and Balek, 1993).
River-Aquifer Interactions in the Wet Tropics
31
In parallel with differences in surface hydrological processes, there are fundamental differences
in groundwater behaviour with regard to mechanisms of recharge and discharge under humid
and sub-humid conditions. Aquifers in the humid tropics tend to fill up rapidly in the wet
season, with the watertable virtually reaching the land surface; further excess rainfall is rejected
because of the absence of storage space, which leads to overland flow (Bonell and Balek, 1993).
In addition, due to a shallow watertable, the volume of aquifer discharge may increase rapidly
following groundwater recharge from excess rainfall, and hence contribute to the peak runoff
response of tropical catchments (Foster and Chilton, 1993). Callaghan and Bonell (2005)
indicate that groundwater remains a neglected area of research in the humid tropics and
highlight the need for better coupling of surface water-groundwater interactions in future
assessments. Geographically, large areas of the humid tropics are underlain by aquifers of
varying geological character. Most of the groundwater system types are characterised by
shallow watertables directly connected to surface waters, which makes groundwater highly
vulnerable to pollution from a range of human activities (Foster and Chilton, 1993). This is
exacerbated by the high rates of precipitation in the humid tropics which cause rapid leaching of
pollutants from urban, agricultural and industrial wastes. According to Bonell (2005), despite
recognition of the connectivity between the surface hydrology with the hydrogeology, globally
there are only a few studies that have parameterised these groundwater systems.
3.1.3 Water quality and nutrients
There have been numerous studies concerned with the physical processes of erosion and
sedimentation in the humid tropics (Rose, 1993); however, little effort has been devoted to
specifically understanding the transport of dissolved constituents through surface and
groundwater pathways in tropical climates. Based on a review of water quality issues in the
humid tropics, Roche (1993) concluded that managing the microbiological aspects of water
quality control is the major issue in terms of direct effects on human health. Bonell and Balek
(1993) note that ‘there is a need for environmental isotope and other conservative tracer studies
in the humid tropics… for assisting studies in water-borne nutrient cycling’. With regard to
water quality, temperature is one of the parameters that differs the most between the temperate
and tropical zones, which has a direct influence on the physical, chemical and biological
properties of an aquatic medium (Roche, 1993). Although the magnitude of water quality
problems, particularly in relation to human health, is perhaps less in northern Australia
compared to other tropical countries, there are similar causes of water pollution. For example,
one of the most urgent water quality problems in the humid tropics is associated with the change
in salt and nutrient cycles and contents, because of land and water resources management.
Tropical systems have an important role in sediment and nutrient transfer to the oceans and
coastal areas (Latrubesse et al., 2005). In relation to northern Australian conditions, discharge of
Chapter 3
32
terrestrial material to the coast occurs predominantly during the major river floods generally
associated with cyclonic rainfall events between November and May (Furnas and Mitchell,
2001). Given the seasonal differences in discharge from monsoonal catchments, it is important
to distinguish between water quality in flow event conditions (flood pulse) from baseflow
conditions (Brodie and Mitchell, 2005). Flow events in northern Australian rivers are
characteristically short and energetic, with water residence times in the river of about one week.
Therefore, water quality measurements during flow events provide information on catchment
contaminant loads (suspended sediment, nutrients and pesticide residue) discharged to
downstream environments. In contrast, nutrient concentrations measured during baseflow
conditions indicate the water quality status which persists for much of the year, which
influences the health of in-stream ecosystems (Brodie and Mitchell, 2005). According to Harris
(2001), the baseflow period is when in-stream interactions between macrobiota, microbiota and
water chemistry have adequate time to fully progress; hence, there is a tight coupling between
water chemistry and water biology.
Given the focus on nitrogen (N) in this thesis, it is pertinent to discuss N inputs into tropical
rivers. Based on a review of studies in northern Australia, Brodie and Mitchell (2005) found that
waters draining pristine rainforest and woodlands have moderate concentrations of dissolved
organic nitrogen (DON), low to moderate concentrations of particulate N and low
concentrations of dissolved inorganic nitrogen (DIN); the dominant form of N being DON.
Similarly, in savannah woodlands and grasslands with low grazing intensity, N speciation is
dominated by DON. However, the authors note that with land clearing for agricultural and
urban development, N concentrations in receiving waters have increased and the form of N has
changed from organic (DON) to inorganic (nitrate, ammonia) forms that are more bioavailable
than DON. Furthermore, it has been suggested that although the concentrations of DIN are
generally low under ‘natural’ conditions, occasional high concentrations of nitrate may be
associated with groundwater discharge to the stream after the main peak flow (Brodie and
Mitchell, 2005). Furnas (2003) observed that in most of the rivers of the Great Barrier Reef
(GBR) catchment, nitrate concentrations are generally highest during wet season flood events,
usually during the first flow or flood event of the season (first flush) when large amounts of
water wash across and through catchment soils. However, Furnas (2003) concluded that in some
rivers such as the Herbert, characterised by having some rainfall throughout the year, the highest
nitrate concentrations occur at the end of the dry season when high nitrate groundwater inputs
make a larger relative contribution to water in the river. Other comparisons between the wet and
dry tropical rivers of the GBR catchment in regards to nutrient export characteristics are
summarised in Furnas (2003).
River-Aquifer Interactions in the Wet Tropics
33
In a review of water and N balances, Bristow et al. (1998) found that deep drainage rates are
higher in both natural and agricultural systems in the Wet Tropics of north Queensland
compared to other regions of Australia. It is suggested that this characteristic, coupled with high
inputs of N in fertilisers, results in considerable N loss below the rootzone. In addition, Bristow
et al. (1998) propose that the intensity of the N cycle in the humid tropics is driven by constant
high temperatures and rainfall which enable year-round biomass production together with high
rates of decomposition and hence nutrient release; this high rate of nutrient release increases the
potential for leaching through tropical soils. Therefore, the role of N in groundwater is
potentially an important consideration in the wet/dry tropics of northern Australia.
3.2 SELECTION OF CASE STUDY AREA
The Herbert River in north Queensland (Figure 3-1) represents a useful study area in the tropics
to develop linkages between land-based activities, stream water quality and potential impacts on
the near-shore marine environment. The catchment is one of thirty-one that drain into the Great
Barrier Reef Marine Park, a marine ecosystem that is recognised internationally for its unique
biological and physical features (Johnson et al., 2000). In addition to its environmental value,
the Park has economic significance, supporting a 4 billion dollar tourism industry. It is believed
that diffuse pollution from cropping and grazing lands in adjacent catchments pose a significant
threat to the Reef. The Herbert is one of the Wet Tropics catchments that drains directly into the
area of the Great Barrier Reef lagoon considered under greatest threat from terrestrial runoff
(Productivity Commission, 2003).
Chapter 3
34
Figure 3-1 Location of the Herbert River catchment showing the Herbert River and its major tributaries. The lower catchment target area for this study is circled.
3.2.1 Catchment water quality issues
Concerns over sediment, nutrient and contaminant export to the Great Barrier Reef World
Heritage Area (GBRWHA) have driven much research into water quality in the Herbert River
catchment. The major water quality issues relate to the transport of nutrients and sediment; it
has been suggested that land management can be improved in the catchment to minimise off-
site exports (Bramley and Roth, 2002). Phosphorus (P) and nitrogen (N) are the nutrients of
most concern. Research has shown that P tends to be largely bound to sediment particles,
whereas the N budget is dominated by dissolved N from both runoff and subsurface flows
(Bartley et al., 2003). Given that N inputs into the system are correlated with fertiliser
application (Bramley and Roth, 2002), N is a link between land use/management and water
resource quality in both surface and subsurface resources. According to Mitchell et al. (1997)
and Brodie et al. (2001), excess N derived from the major land uses in the catchment has the
potential to impact on in-stream and near-shore water quality.
It is well documented that high concentrations of N, especially inorganic nitrate, are found
particularly in surface waters draining land under sugarcane (Bartley et al., 2003; Furnas, 2003;
Bramley and Roth, 2002; Bramley and Muller, 1999). Groundwater studies have also shown
there to be subsurface losses of nitrate beneath sugarcane growing areas (Weier, 1999). Given
an average fertiliser input of 160 kg N/ha/yr and an average output of roughly 80 kg N/ha/yr in
River-Aquifer Interactions in the Wet Tropics
35
millable cane, is suggested that about 50 percent of the N applied in the sugarcane industry can
potentially be lost through denitrification, surface runoff and/or leaching to groundwater (Bohl
et al., 2001). Under current management practices of minimal soil erosion (e.g. minimum tillage
and green cane harvesting/trash blanketing), losses of nitrogen from cane land are attributed to
fertiliser as the major source of nutrients, rather than a sediment source (Brodie and Mitchell,
2005). Based on surface water studies, it is considered that the tendency for nitrate
concentrations to be elevated in the Herbert River during low flow periods is due to inputs of
high-nitrate groundwater (Furnas, 2003; Furnas and Mitchell, 2000). Furthermore, it has been
proposed that the quantity of nitrate introduced into the Herbert River is comparable to the
estimated loss of fertiliser N to groundwaters (Furnas and Mitchell, 2000). Based on a study of
radium isotopes in riverine muds in the Herbert River estuary, Brunskill (2000) also suggested
that groundwater inputs are important during medium to low river discharge periods. Whilst
these studies raise the possibility of groundwater discharge to the river, the interaction of
groundwater and surface water in the lower catchment has not been investigated in any detail. In
addition, despite concerns over N levels found in the Herbert River and in some of the shallow
aquifers, no previous work has explicitly assessed the role of subsurface N on surface water
quality in the catchment.
The target area for this study is the lower Herbert River catchment, due to the occurrence of
sugarcane farming which is considered to be the major contributor of N to surface waters
(Bartley et al., 2003; Bramley and Muller, 1999). In addition, several key datasets from
previous studies and ongoing monitoring are available for the lower catchment (Section 3.5).
3.2.1.1 Previous N studies on surface water
Water quality studies previously undertaken in the lower catchment have generally been
concerned with both N and P. Below is a summary of key findings from past research relating
specifically to N, for both surface water and groundwater resources.
Two major water quality studies in the lower catchment were previously undertaken by the
Australian Institute of Marine Science (AIMS) in collaboration with the Bureau of Sugar
Experimental Stations (BSES) (Furnas et al., 1995), and the Commonwealth Scientific and
Industrial Research Organisation (CSIRO) (Bramley and Muller, 1999). The AIMS study was
primarily concerned with determining riverine export of nutrients and fine sediment to the Great
Barrier Reef lagoon. The CSIRO sampling programme involved extensive water and soil
sampling to resolve source areas for the nutrients and sediment leaving the catchment.
Sampling by AIMS took place during 1989-1994 at three sites along the lower Herbert River.
Key findings from the study were that:
Chapter 3
36
• the concentrations of DIN and particulate N in river source waters are likely to be
similar throughout the catchment, or that chemical processes stabilise concentrations in
river and soil waters;
• the concentration of DON decreases downstream, indicating dilution with low-DON
water on the floodplain or in-stream consumption (bacterial mineralisation) of DON in
the lower reaches of the river;
• the concentration of nitrate is generally higher at the downstream sampling site than that
measured upstream, particularly during low flow periods;
• most of the nitrate exported is considered to come from a floodplain source, with
agricultural fertilisers the most likely source.
Of direct relevance to the current research is that according to Furnas (2003), the tendency for
nitrate concentrations to be elevated during low flow conditions and diluted during floods
suggests that inputs of high-nitrate groundwater are responsible for the higher concentrations;
dilution occurs because groundwater inputs from aquifers are relatively constant and not closely
coupled to surface runoff.
The CSIRO water quality sampling was undertaken during October 1992 – May 1995 at 33
surface water sites (reduced to 19 by the end of the study) around the lower catchment: the sites
were selected to reflect the major land uses, soil types and subcatchments. The CSIRO data
showed that:
• N concentrations tend to be greater downstream than upstream, indicating an export of
nutrients from land draining into the Herbert River between these areas;
• the concentration of N in streams draining land under sugarcane tends to be greater than
in streams draining other land uses (grazing and forestry);
• peak wet season events dominate the annual riverine flux of nutrients.
In addition, Bramley and Roth (2002) found that with respect to total nitrogen in the lower
catchment, approximately 31%, 9% and 3% of samples collected from streams predominantly
draining cane land, grazing and forestry, respectively, were above the ‘interim trigger levels for
assessing possible risk of adverse effects due to nutrients’ (ANZECC and ARMCANZ, 2000).
Bramley and Johnson (1996) further suggested that nutrient concentrations in the Herbert were
generally below ANZECC (2000) target levels for the protection of freshwater ecosystems,
except during high flow conditions when the trigger values were exceeded; hence, the authors
concluded that nutrient loss in the catchment is event based and insignificant outside the wet
season months. Although nutrient losses from intensively managed cane lands exceeded those
River-Aquifer Interactions in the Wet Tropics
37
from other land uses, Bramley and Johnson (1996) also noted that these losses might be
expected to occur in these areas irrespective of crop type due to strongly seasonal climate and
high rainfall intensities. Nonetheless, Bramley and Roth (2002) concluded that land
management can be improved in the catchment to minimise this off-site export of nutrients.
A recent modelling study for the entire Herbert River catchment (Bartley et al., 2003) using a
sediment budget, SedNet (Prosser et al., 2001), combined with a nutrient budget component,
ANNEX (Young et al., 2001), was undertaken to spatially identify the major sediment and
nutrient sources. With regard to N, the study showed that dissolved forms of N predominate and
that the N budget is dominated by losses from canelands on the floodplain, followed by open
forests and cultivated areas of the middle and upper catchment. Nutrient scenario results
suggested that reduced fertiliser application rates on cane lands (reduced from 200 kg/ha/yr to
130 kg/ha/yr) could produce a decrease of 27% in dissolved inorganic N levels and a 10%
decrease in the overall N budget.
3.2.1.2 Previous N studies on subsurface water
Although the majority of water quality research in the catchment has focussed on surface water
rather than groundwater, the studies summarised below of Bohl et al. (2000a), Bohl et al.
(2000b) and Bohl et al. (2001), which are specific to the lower Herbert River catchment,
provide some important background information for this research in regards to subsurface N
losses under different soil types. A further study by Thorburn et al. (1999) modelled the impact
of trash retention on soil nitrogen, using the lower catchment as a case study site. The
magnitude of N losses to groundwater from sugarcane have previously been investigated for
various soils and areas in the Australian sugar industry such as: the freely drained basaltic soils
in the South Johnstone catchment (Rasiah et al., 2003; Hunter and Walton, 1997; Reghenzani et
al., 1996); under irrigated cane in the Bundaberg and Burdekin delta regions (Stewart et al.,
2005; Kuhanesan et al., 1998; Verburg et al., 1998); and in several coastal areas of northeast
Australia (Thorburn et al., 2003). Bristow et al (1998) summarised research relating to water
and nitrogen balances in natural and agricultural systems in the Wet Tropics of north
Queensland; recent modelling/experimental studies by Meier et al. (2006) and Thorburn et al.
(2005) add to this body of research.
Rasiah and Armour (2001) noted that nitrate leached below the crop-root zone may be adsorbed
onto soil, move laterally to discharge into streams and rivers, enter deep groundwater, and/or
denitrify in the profile. Furthermore, the amount of N adsorbed in soil depends on: anion
exchange capacity, net negative charge, ionic strength of the soil bulk solution, nitrate
concentration, competition with other anions, pH2O, pH, anionic composition of the soil
solution, cation exchange capacity, and leaching. Given the range of factors that potentially
affect N leaching, the transferability of findings from one study site to another in the Wet
Chapter 3
38
Tropics is questionable. For this reason only literature specific to the lower Herbert River
catchment is summarised below.
Based on field sampling over two wet seasons during 1997-1999, Bohl et al. (2000b)
determined the components of the water and nitrogen balance on a range of soils under
sugarcane in the Ripple Creek area in the lower catchment. Gaseous losses (such as
denitrification) or losses via surface runoff were found to be the major loss pathways for N
under wet conditions (up to 40% of the amount of applied nitrogen), rather than nitrate leaching
via subsurface flow (groundwater and interflow). However, on highly permeable soils, such as
the sandy soils of the riverbank, N losses to groundwater from fertiliser were found to be up to
45% of that applied. Large variations in N loss were observed between bore sites as well as
between sampling years, with the highest losses to groundwater under plant cane (compared to
ratoon) due to the less developed root system. Bohl et al. (2000b) suggested that the
distribution, size and intensity of rainfall after fertiliser application are important factors in
dictating N losses.
Bohl et al. (2001) expanded on the previous work by Bohl et al. (2000b) to examine the spatial
distribution of N leaching losses based on pedological (mapped soil types) and hydrological
features. The study found that the more freely draining soils of the alluvial fans and sandy river
banks had the highest leaching losses, while the lowest losses were estimated for the heavy soils
on the plain. The authors concluded that, in general, the risk of nitrate leaching is comparatively
low, with the exception of pedohydrological units characterised by high internal permeability
and high drainage in relation to landscape position.
3.3 CATCHMENT CHARACTERISTICS
The Herbert River catchment (Figure 3-1) can be divided into four distinct physiographic
sections (Bartley et al., 2003). The upper part is characterised by extensive cattle production and
other minor agricultural activities such as horticulture and dairy. The central part of the
catchment, comprising the deep Herbert River gorge, is predominantly Wet Tropics World
Heritage Area, State Forest and timber reserves. The lower catchment comprises the Herbert
River floodplain and the southern coastal section, containing a network of streams that drain
directly to the coast. Sugarcane farming dominates the banks of the major rivers in the lower
catchment, while the remaining area supports cattle grazing, native vegetation and plantation
forestry (Figure 3-2). Agricultural and pastoral production are the largest users of land in the
entire catchment, with less than 1% of the area allocated to industrial and urban uses (Johnson et
al., 2000).
River-Aquifer Interactions in the Wet Tropics
39
Figure 3-2 Land cover in the lower Herbert River catchment (adapted from Bramley and Muller 1999). Note that the Herbert River runs through the sugarcane growing areas.
Land cover and land use in the catchment has changed substantially since European settlement
in the early 1860’s, particularly in the lower catchment (Johnson et al., 2000, 1999). Prior to
settlement, natural vegetation in the lower catchment was dominated by grassland, riparian
forests and freshwater wetlands. However, extensive land use change for grazing and
particularly sugarcane production has since led to significant losses in riparian and wetland
areas, which prior to clearing, are considered to have provided buffer strips protecting coastal
river systems, estuaries and shorelines (Johnson et al., 1999).
The geology of the region comprises undifferentiated Paleozoic quartzite, Carboniferous acid
volcanics, Carboniferous granites, Cainozoic basalts and Quaternary alluvium and beach sands
(Figure 3-3). A detailed account of the geology of the Ingham district in the lower catchment is
provided in Rienks et al. (2000). Soils are closely related to geology and in the lower catchment
are heavily influenced by fluvial processes. Soils are mostly coarse textured in the upper
catchment and fine textured clay/loam soils in the lower catchment (Johnson and Murray,
1997). Wilson and Baker (1990) defined three major geomorphological units for the Herbert
River floodplain: (1) soils of the alluvial fans derived from granite and acid volcanic rocks; (2)
heavier textured (duplex) soils of low permeability of the alluvial plain characterised by perched
watertables (waterlogging); and (3) sandy, freely drained soils of the riverbank bordering the
Herbert River.
Figure 3-3 1:250,000 geological map comprising the lower Herbert River catchment (BMR, 1965). The town of Ingham is marked for reference (refer to Figure 3-1).
INGHAM■
River-Aquifer Interactions in the Wet Tropics
41
Based on detailed mapping in the sugarcane areas of the lower catchment, 24 soil types have
been identified, which can be broken down into seven broad categories based on similarity of
parent materials (Wood et al., 2003).
3.3.1 Climate
The catchment is characteristically tropical, experiencing warm humid summers and mild dry
winters. Mean annual rainfall is 1370 mm and ranges from greater that 3000 mm in the north-
east of the lower catchment, to 750 mm in the extreme west of the upper catchment.
Approximately 74% of the mean annual rainfall occurs from December to March (Johnson and
Murray, 1997). The marked seasonal rainfall pattern is illustrated in Figure 3-4 for five rainfall
stations in the lower catchment. The upland areas to the northwest and southwest have lower
average monthly rainfall compared to closer to the coast, particularly during the wet season. The
cumulative residual rainfall curve provides a measure of the accumulated deficit or surplus of
rainfall at a particular time, relative to average rainfall. The curve can be interpreted at different
scales: a positive slope indicates a cumulative period of above average monthly rainfall while a
negative slope indicates the reverse. As illustrated in Figure 3-5, the early 1900’s to 1980’s was
an extended period of above average monthly rainfall, while the following period to 2004
represents a cumulative deficit. Fluctuations in the residual mass curve are also apparent within
the gross wet and dry trends, representing shorter duration rainfall surplus and deficit periods
(Figure 3-6).
%U
%U
%U
%U
%U
32091
32043
32045
32023 32031
10 0 10 KmN
%U Rainfall stations
Streams
Herbert R
0
100
200
300
400
500
1 2 3 4 5 6 7 8 9 10 11 12
Month
Ave
rage
rain
fall
(mm
)
3202332031320433204532091
%U
%U
%U
%U
%U
32091
32043
32045
32023 32031
10 0 10 KmN
%U Rainfall stations
Streams
Herbert R
0
100
200
300
400
500
1 2 3 4 5 6 7 8 9 10 11 12
Month
Ave
rage
rain
fall
(mm
)
3202332031320433204532091
Figure 3-4 Mean monthly rainfall in the lower Herbert River catchment. Source: BoM
Chapter 3
42
-6000
-4000
-2000
0
2000
4000
6000
1900
1904
1909
1914
1919
1924
1929
1934
1939
1944
1949
1954
1959
1964
1969
1974
1979
1984
1989
1994
1999
2004
Cum
ulat
ive
resi
dual
rain
fall
(mm
)
Figure 3-5 Cumulative residual rainfall at station 32045 in the lower catchment. The arrows depict periods of above (blue) and below (red) average monthly rainfall based on records since 1900. By definition, zero residual rainfall represents the overall average monthly rainfall over the averaging period.
0
50
100
150
200
250
300
350
400
450
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Dai
ly ra
infa
ll (m
m)
0
1000
2000
3000
4000
5000
Res
idua
l rai
nfal
l (m
m)
rainfallresidual
Figure 3-6 Daily rainfall versus the cumulative deviation of residual rainfall from the mean at station 32045. Years are labelled as the approximate start of the wet season (November). Source: BoM (rainfall)
3.3.2 Rivers and aquifers
The entire catchment drains an area of approximately 10,000 km2 to the Coral Sea and is the
largest of the river systems in the sub-humid to humid tropical region of northeast Australia
(Johnson and Murray, 1997). The main drainage systems in the catchment are the Herbert, Wild
and Stone Rivers. Mean annual flow in the Herbert River is 3440 GL, with the highest monthly
flows occurring from November to May. Mean annual runoff for the catchment is 4991 GL (493
River-Aquifer Interactions in the Wet Tropics
43
mm), with a runoff to rainfall ratio of around 37%. Flooding of the Herbert and Stone Rivers is
confined to the wet season, generally from December to March. Tropical cyclones and
associated heavy rainfall are usually responsible for major flooding (Johnson and Murray,
1997). Further physical characteristics, as well as chemical trends in the Herbert River, are
provided in Chapters 6 and 7.
In the upper and middle catchment, groundwater supplies are available from alluvial, fractured
basalt and other fractured rock aquifers, with the majority of aquifers in fractured material.
Groundwater quality in the upper and mid-sections of the catchment is considered to be good.
Investigations of the sub-surface hydrology have only been carried out in detail for the lower
catchment, which represents an alluvial aquifer system (Cox, 1979). Reinterpretation of the
hydrogeology as part of this study is provided in Chapter 4. The shallowest sandy aquifer is of
greatest economic significance as it is very permeable; however, pumping rates are limited
because of the shallow depth. The aquifer is mainly used to supply town water and for domestic
purposes, with only minimal use for irrigating crops. Although the quality of groundwater in the
lower catchment is considered to be good, high nitrate concentrations and high salinities have
been found in isolated areas. The hydrogeochemistry of the alluvial aquifer system is examined
in detail in Chapter 5. The threat of salt water intrusion on the coastal fringes has traditionally
been the main consideration for groundwater management in the area, as groundwater extraction
in excess of recharge may cause groundwater levels to decline, with a subsequent reversal in the
groundwater flow gradient. Although seawater intrusion into the coastal aquifers can cause
significant deterioration in water quality, formal licensing and allocation systems have not been
adopted in the catchment because the volume of groundwater extracted is significantly less than
the long term yield (Johnson and Murray, 1997).
3.4 RESEARCH APPROACH
As outlined in Chapter 1, specific questions that this research will address include:
(1) What is the nature of river-aquifer interactions in the lower Herbert River catchment,
particularly during the dry season?
(2) What is the significance of river-aquifer interactions for the nitrogen budget of the river?
(3) What are the implications of these interactions for nutrient monitoring, management,
and policies relating to water quality at a catchment scale?
Water is a vehicle for mobilising and transporting nutrients and other dissolved constituents;
therefore, knowledge of the dynamics of water movement, including river-groundwater
connectivity, is central to understanding the transport of such components. In addition, an
Chapter 3
44
understanding of groundwater dynamics is an essential element of characterising the
relationships between surface and subsurface waters. In light of these considerations, including
the research aims, the thesis has been divided into four analytical chapters. Chapters 4 and 5
examine the hydrogeology of the alluvial aquifer system; Chapters 6 and 7 specifically
investigate river-groundwater interactions, given the hydrogeological framework. Physical
datasets are derived from existing sources, and provide a background for understanding the
hydrogeology and hydrology in the catchment. Verification and extension of concepts is
accomplished through analysis of an extensive database of hydrochemical information which
was collected for the purposes of this research. Hence, a conceptual understanding of river-
aquifer interactions in the catchment is developed through the course of the thesis. Drawing on
the conceptual framework for river-groundwater interactions, the implications for N transport in
groundwater and potentially to the river are also progressed. Establishing the significance of
groundwater as a vector for dissolved inorganic forms of N to the river is the culmination of the
core analytical components of the thesis. The environmental significance of the key research
outcomes for in-stream and marine ecosystems provides a context for recommendations
regarding nutrient monitoring, management and water policy.
Whilst the broad research approach outlined above is driven by the research questions, specifics
of the methodology are governed by the available datasets and resources to collect additional
data. Attributes of the case study catchment necessarily influence the design of the data
collection program, as discussed in the following section. In reality, the research approach and
methodology are interrelated, whereby decisions about each are the result of an iterative
process.
3.5 DATA COLLECTION
3.5.1 Existing data and applicability
As discussed in Section 3.2.1.1, river water monitoring in the lower catchment has previously
been undertaken by the research agencies of AIMS and CSIRO (Bramley and Muller, 1999;
Furnas et al., 1995). These extensive water sampling programs had a focus on surface water
quality and therefore complementary groundwater data was not collected. Additionally, water
quality samples in these studies were collected upstream of one of the Herbert River tributaries
that drains an intensive sugarcane farming area (Ripple Creek), and were not collected in the
lower tidal reaches of the river. Water quality data is collected by the Queensland Department of
Natural Resources and Water (QDNRW) at least twice a year at each of their monitoring bores
(approximately 70 bore locations, some of which have multiple pipes) and also at two stream
gauges in the lower catchment. Determining the correlation between in-stream water chemistry
and adjacent groundwater sites is difficult with few surface water monitoring sites and when the
River-Aquifer Interactions in the Wet Tropics
45
timing of water quality sampling does not necessarily coincide for surface water and
groundwater. In addition, QDNRW groundwater samples are not routinely tested for other
dissolved species of N such as nitrite and ammonium, which are potentially important
components of total DIN. The reliability of existing groundwater quality data is also uncertain.
Whilst previous groundwater studies in a subcatchment of the study area (Bohl et al., 2001;
Bohl et al., 2000a; Bohl et al., 2000b) provide valuable background information, all
groundwater samples in these studies were collected from within the upper 10 cm of the aquifer.
Although the concentration of N at this depth is considered to provide an estimate of the
nitrogen which had passed through the profile after leaving the root zone (Bohl et al., 2000b),
similar nitrogen samples were not collected for the deeper aquifer which may also be an
important nitrogen store. In addition, the studies were performed during the wet season only,
were of limited spatial extent, and did not explicitly look at interactions with surface water
resources.
Existing water quality data accessed for this study are summarised in Table 3-1. Although
incomplete for the aims of this study, the available data is particularly useful for observing
temporal variations in water chemistry in the Herbert River. In addition to water quality
information, supplementary data specific to the lower catchment was obtained from numerous
sources (Table 3-1).
Table 3–1 Available datasets for the lower Herbert River catchment
Data type Source
time series river flow
time series stage height
surveyed river profiles
river water chemistry
groundwater chemistry
bore logs
time series groundwater levels
hydrogeological mapping
time series rainfall
GIS data
QDNRW1
QDNRW
HSC2
QDNRW; Bramley and Muller (1999)
QDNRW
QDNRW
QDNRW
Cox (1979)
BoM3
HRIC4; CSIRO5; QDNRW 1 Queensland Department of Natural Resources and Water 2 Hinchinbrook Shire Council 3 Bureau of Meteorology 4 Herbert Resource Information Centre 5 CSIRO Land and Water
Chapter 3
46
To address the concerns outlined above, a sampling program was designed to collect consistent
river and groundwater samples in the lower Herbert River catchment (Dixon-Jain et al., 2005).
Key considerations included: types of samples required; scale and spatial distribution of
sampling sites; timing and logistics of sample collection; and technical aspects of sample
preparation and analysis.
3.5.2 Sample types for this study
Based on the aims of the study, water quality samples were collected from surface waters and
groundwater, and of rainfall, for laboratory analysis of major ions, stable isotopes of water and
dissolved N species. In addition, radon samples were collected from surface waters and
groundwaters. Field measurements of water quality parameters and other physical
characteristics were also recorded. Table 3-2 summarises the primary reasons for analysing
samples for environmental tracers and measuring other parameters. Further details on the
principles of environmental tracers and ion, isotope, and nitrogen chemistry, are provided in
Chapters 5 and 7. Field and laboratory data generated during the sampling programs are
summarised in Appendix A.
Table 3–2 Summary of laboratory and field measurements
Measurements Description Purpose
major ions standard cations, anions water sources, evolution, mixing
nutrients TON+, nitrite, ammonium* hotspots, transformations, transport
stable isotopes δ18O, δ 2H water sources, mixing, evaporation
radon 222Rn groundwater discharge
other lab TDS#, alkalinity, pH, EC@ hydrochemical environment, water sources, evolution
field water quality pH, EC, T, Eh^, DO' hydrochemical environment, stable water samples (groundwater), detect hydrochemical change
other field water column depth, river width, flow, depth to groundwater, GPS location, sampling method, date and time
groundwater discharge calculations (radon), water level elevation calculations, mapping
+ Total Oxidised Nitrogen (dissolved) * Nitrate concentration is calculated as the difference TDN - nitrite # Total Dissolved Solids (evaporation to dryness) @ Electrical Conductivity at 25 oC (specific electrical conductance) ^ Redox potential ' Dissolved oxygen
River-Aquifer Interactions in the Wet Tropics
47
As described below in Section 3.5.4, three sampling trips were undertaken. Major ion and
nutrient data were collected during each field campaign; stable isotopes of water were only
analysed during May 2004, while radon samples were collected in October 2004 and June 2005.
The spatial coverage of river and groundwater samples was greatest during the final sampling
program, largely due to resolution of logistical issues through increased familiarity with the
area. Despite differences in the types and distribution of samples, it is shown in subsequent
chapters that data from each trip has its merits for different aspects of the research.
3.5.3 Site selection
An important consideration for data collection is the scale of the required output, as this affects
the spatial distribution of sampling sites. This thesis is concerned with processes at the
catchment scale; therefore, sampling was undertaken at an appropriate level of detail in order to
detect major hydrochemical changes in the river and groundwater. The density of groundwater
sites (and hence the aquifers sampled) was largely determined by the availability of QDNRW
monitoring bores. Groundwater was sampled from bores as close as possible, and on both sides,
of the Herbert River. Where there were obvious gaps in spatial distribution, samples were also
collected from selected private bores (owned by landholders and the Hinchinbrook Shire
Council). Note that all groundwater samples were from the alluvial aquifer system.
Surface water samples were collected along the entire length of the lower Herbert River, at
intervals determined by access, the location of tributaries, and the distribution of groundwater
sites. Where possible, samples were collected immediately downstream of the entry point of a
tributary: in some locations it was also feasible to sample upstream of tributary inflows. Nash’s
Crossing, within Yamanie National Park, represents the upstream extent of sampling in the
river: this location was chosen as it is situated above the region of sugarcane farming. Hence,
the chemistry of the river at Nash’s Crossing reflects inputs from the upper catchment. At the
downstream extent, samples were collected towards the river mouth in order to examine the
hydrochemistry downstream of the main tributaries and to assess the quality of water potentially
entering the marine environment. Note that samples within the tidal zone were collected further
downstream than in previous studies. In addition to the Herbert River, sampling was undertaken
in selected tributaries, depending on accessibility (refer to Section 3.5.5). A map of river and
groundwater sampling sites is provided in Figure 3-7.
Chapter 3
48
U
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INGHAM
Surface water samples#S
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5 0 5 Km
Figure 3-7 Location of surface water, groundwater, and rainfall collection sites during the three sampling periods. The two QDNRW stream gauges (116001, 116006) and major town of Ingham are also indicated.
3.5.4 Timing of sampling
The tropical climate zone of Australia provides a unique opportunity to study baseflow without
(or with minimal) surface water flows. Therefore, given the specific interest of this thesis on
groundwater contributions to surface waters, water quality sampling was purposely undertaken
during the dry season, when surface runoff is minimal and baseflow dominates streamflow. Due
to the more stable water quality conditions in the dry season, compared with the rapid changes
in water quality parameters during an event, the sampling frequency required for collecting
water quality data during low flow conditions is much less than during event flows (Brodie and
Mitchell, 2005). In the lower catchment, baseflow (low flow) conditions persist for much of the
year; hence, it was considered that grab-samples, during months representing the beginning and
end of the dry season, would be adequate for comparing differences in river-aquifer connectivity
relationships between the extremes of the dry season.
Note that baseflow is comprised of numerous potential sources: groundwater discharge; bank
discharge; unsaturated zone flow (interflow); delayed surface water (e.g. wetlands, lakes); and
delayed groundwater (e.g. perched aquifers) (Evans, 2005). As groundwater discharge is
River-Aquifer Interactions in the Wet Tropics
49
generally the major process, it is assumed for the purpose of this study that all baseflow, unless
otherwise stated, represents discharge from groundwater sources.
Three sampling trips were undertaken during the dry seasons in 2004-2005 (Figure 3-8). Water
levels in the aquifers of the catchment are at their highest at the cessation of the wet season,
generally by the end of April. Therefore, the start of May 2004 (11/5-23/5) was selected to
conduct an initial sampling program, to capture water quality during conditions of high aquifer
recharge and potentially high physical river-groundwater connectivity. To complement this
sampling, a subsequent collection was undertaken in October-September 2004 (23/10-4/11),
before the start of the 2004-05 wet season. This period was characterised by low groundwater
levels and low river stage. A final field session was undertaken in May-June 2005 (29/5-11/6) to
repeat sampling procedures and hence verify earlier results. Note that for simplicity, the last two
sampling periods are henceforth referred to as October 2004 and June 2005.
0
25
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Jan
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ar 0
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October2004
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0
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ar 0
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m)
0.1
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1000
Stre
am d
isch
arge
(GL/
day)
rainfall stream discharge
October2004
May2004
June2005
Figure 3-8 The beginning of each sample collection period in 2004-2005, in relation to stream discharge and rainfall in the lower catchment. Source: QDNRW (stream discharge); BoM (rainfall)
Chapter 3
50
3.5.5 Logistics, materials and methods
3.5.5.1 Sampling technique
Access to the rivers in the lower catchment was hampered by thick riparian rainforest as well as
hazards associated with wildlife. Consequently, river samples were collected either from
bridges, in the middle of the river by boat, or from the riverbank. Different water collection
devices were used in these different scenarios. For bridge samples, water was collected by
hauling a bucket attached to a rope over the bridge; in the case of low crossings or riverbanks,
an extendable pole was used with an attached sample bottle or by use of a small submersible
pump (Amazon) attached to the pole. Samples were collected in the middle of the river or in the
deepest water, upstream of the bridge/crossing and away from bridge support structures. This
was to ensure that water collected was not influenced by any accumulated debris. Samples from
a boat were collected in the middle of the river using a plastic scoop attached to a short pole or
the submersible pump. Where there was adequate control of the sampling depth, sampling was
from a depth of around 10 cm below the stream surface. The majority of groundwater samples
were collected from QDNRW monitoring bores using a Grundfos MP1 submersible pump.
Bores were purged to remove stagnant water before obtaining a representative sample. At least
three casing volumes of water were removed and pH, temperature and electrical conductivity
(EC) were allowed to stabilise before sampling (MDBC, 1997). Other groundwater samples
were collected from domestic bores equipped with a pump.
Basic field measurements of pH, temperature, EC, redox potential and dissolved oxygen were
determined at all sampling locations using probes connected to a field meter, either directly
from the water source (e.g. in the river or flowing through a hose attached to a pump) or in a
bucket of unfiltered sample immediately after collection. All water collection vessels were pre-
rinsed with the new sample three times to decontaminate between samples. Sample bottles were
also pre-rinsed three times with unfiltered sample, followed by a final rinse with filtrate before
filling. In order to check the accuracy of field sampling procedures and laboratory analyses, a
set of duplicate and spiked1 duplicate samples were collected for major ion and nutrient analyses
at the beginning, middle and end of each field program.
3.5.5.2 Sample preparation and preservation
Samples were either filtered on site or collected in glass bottles and filtered later on the same
day. All filtration equipment was rinsed with deionised water between samples, followed by
rinsing with filtrate. River and groundwater was filtered through a 0.45 μm membrane filter
(nitrate-free for N samples).
1 Containing a known concentration of major ions or nitrate (for N samples)
River-Aquifer Interactions in the Wet Tropics
51
Samples were preserved in distinct bottles2 for analysis of different chemical parameters,
including major ions, nutrients, stable isotopes and radon. The procedure used for radon
sampling is outlined in Appendix B. Cation samples were acidified to < pH 2 with 1 mL nitric
acid. All samples were refrigerated after collection, with nutrient samples frozen at the end of
each day (and transported frozen to the laboratory).
3.5.5.3 Analytical techniques
Nutrient analyses were performed by flow injection colorimetric methods to determine the
concentrations of dissolved total oxidised nitrogen, nitrite and ammonium3. The concentration
of dissolved nitrate was calculated as the difference between dissolved total oxidised nitrogen
and nitrite. Analyses of major ions were undertaken by ICP-OES (Inductively Coupled
Plasma Optical Emission Spectrometry) for cation and ion chromatography for anion
determinations4. Radon was counted in the laboratory5 by liquid scintillation, on a LKB Wallac
Quantulus counter using the pulse shape analysis program to discriminate alpha and beta decay
(Herczeg et al., 1994). Note that corrections were made for radioactive decay that occurred
between the time of sampling and the time of analysis. Alkalinity, total dissolved solids (TDS),
pH and EC determinations were performed in the laboratory. Alkalinity of the samples was
determined by potentiometric titration; TDS was measured by evaporation to dryness at 180°C.
3.6 CHAPTER SUMMARY
This chapter progressed the theme of groundwater-surface water interactions which was
introduced in Chapters 1 and 2, with a focus on tropical hydrology. A review of research in
tropical systems indicates that there has been far greater emphasis worldwide, including in
Australia, on hydrology in temperate climates. While temperate and tropical regions share many
similarities, there are some fundamental differences related to rainfall patterns and other
climatic factors that affect both stream hydrology and groundwater characteristics. These
differences in turn affect the way in which surface and subsurface waters interact. River systems
in tropical catchments have distinct high and low flow periods that are largely influenced by
seasonal rainfall patterns. Groundwaters are commonly recharged to levels close to the land
surface during the wet season, and thus can contribute to the peak runoff response of rivers.
2 Sample bottled were pre-rinsed with unfiltered sample, followed by a final rinse with filtrate before filling 3 Nutrient analyses were undertaken by ECOWISE Environmental and The Murray Darling Freshwater Research Centre 4 Major ions analyses were performed by The Bureau of Rural Sciences and CSIRO Land and Water Laboratories, South Australia 5 CSIRO Land and Water Laboratories, South Australia
Chapter 3
52
Given the shallow depth of the watertable, groundwater is vulnerable to leaching of pollutants
from the land surface. This in turn has the potential to impact on the quality of surface water
resources.
The water quality status of tropical rivers can vary markedly between seasons due to differences
in the residence time of water and flow pathways. While sediment and nutrient transport during
flow events provide an indication of catchment contaminant loads, water quality measurements
during low flow conditions reflect more typical conditions that persist for much of the year.
Therefore, it might be expected that seasonal variations in stream water quality are more
pronounced in tropical catchments. The potential importance of groundwater as a vector for
dissolved species, such as nitrate, to surface waters was also highlighted.
The chosen case study area in the wet/dry tropics of north Queensland was introduced,
including a discussion of water quality issues and a summary of previous research related to N
in the lower Herbert River catchment. Nitrogen in surface and subsurface waters is associated
with fertiliser inputs onto cane land, thus N is a link between land management practices and
water resource quality. Despite concern over nitrate concentrations found in both the Herbert
River and in the shallow aquifer systems, the explicit role of groundwater as a vector for N has
not previously been explored. Therefore, this thesis will characterise river-groundwater
connectivity in the lower catchment and hence elucidate the significance of subsurface N for
surface water quality.
The final part of the chapter outlined the general research approach of the thesis and discussed
data availability and additional requirements. Details of the sampling methodology employed
for this research were also outlined. Key considerations for data collection included: types of
samples required; scale and spatial distribution of sampling sites; timing and logistics of sample
collection; and technical aspects of sample preparation and analysis. This chapter concludes the
introductory chapters of the thesis. The following chapters (4-7) have an emphasis on
examining the dynamics of water and nutrient movement through surface and subsurface
waters, drawing on a range of physical and chemical-based techniques. Whilst data analysed is
specific to the lower Herbert River catchment, the methodology is generic.
53
Chapter 4 Hydrogeological Framework
4.1 INTRODUCTION
Central to this thesis is the role of groundwater as a possible transport vector for dissolved
forms of nitrogen to surface waters. A first step to understanding this role is the development of
a hydrogeological framework to describe the way in which water, and its dissolved constituents,
are transported through the subsurface (Section 2.4.1.1). This chapter is the first of two chapters
centered on characterising the hydrogeology of the alluvial aquifer system in the case study
area. Whilst this chapter considers the physical aspects, Chapter 5 validates and extends the
physical model through the analysis of hydrogeochemical data. Key questions that the current
chapter addresses include:
• what is the spatial extent of the aquifers?;
• what is the nature of vertical hydraulic connection between the aquifers?;
• how does groundwater move laterally through the lower catchment?
Whilst subsequent chapters (6 and 7) are specifically aimed at characterising river-aquifer
connectivity, the relationship between surface water and groundwater is briefly examined in this
chapter as evidence of potential for interaction. The analyses presented in this chapter provide a
framework to examine the speciation, concentrations, and spatial variability of dissolved
nitrogen in groundwater, and hence to explore the prospect of a subsurface source of nitrogen to
the river.
4.1.1 Key concepts and definitions
Detailed discussions of the fundamental physical principles of hydrogeology are provided in
classic textbooks such as Freeze and Cherry (1979) and Heath (1987). The aim here is to
summarise the main aspects that are of direct relevance to the analyses presented in this chapter.
Darcy’s Law is fundamental to hydrogeology as it describes the factors controlling the
movement of groundwater. According to Darcy’s Law, Q = KA × dh/dl where Q is the rate of
groundwater flow (volume per unit of time), K is the hydraulic conductivity (dependent on the
size and arrangement of pore spaces and on the dynamic characteristics of the fluid), A is the
cross-sectional area (at right angles to the flow direction, through which flow Q occurs), and
dh/dl is the hydraulic gradient (the change in head per unit distance). An aquifer can be defined
Chapter 4
54
as a saturated permeable geologic unit that can transmit significant quantities of water under
ordinary hydraulic gradients (Freeze and Cherry, 1979). This is in contrast to a confining layer,
which is a less-permeable geological unit or layer that retards the movement of water in and out
of an adjacent aquifer. Aquifers tend to have relatively high hydraulic conductivities, and are
usually comprised of unconsolidated sands and gravels, permeable sedimentary rocks and
heavily fractured crystalline rocks. In contrast, confining beds possess a very low hydraulic
conductivity and are typically comprised of clays, shales and dense crystalline rocks.
Aquifers can be confined or unconfined: this important distinction affects hydraulic behaviour.
A confined aquifer has a confining bed above and below, and is fully saturated with respect to
water in the pore spaces of the aquifer material. As water in a confined aquifer is pressurised,
the measured water level in a well tapping this type of aquifer will be above the top of the
aquifer, but not necessarily above the land surface. If the latter, the well is considered to be a
free-flowing artesian well. In an unconfined aquifer, the watertable forms the upper boundary of
the saturated zone. As the aquifer is not fully saturated, the saturated thickness can vary over
time, for example, with changes in recharge or groundwater use.
The measured level at which water stands in a well tapping either a confined or unconfined
aquifer is referred to as the hydraulic head, which is the sum of the pressure head, elevation
head, and velocity head. Given that groundwater movement is generally slow, the velocity head
is often disregarded. Moreover, for a watertable aquifer, the pressure head at the watertable is
zero because it is equal to atmospheric pressure, and thus the hydraulic head is equal to the
elevation head at that point. Measurement of depth to groundwater and subtraction from the
elevation of the top of the well casing provides a measure of total head. A groundwater surface
map can be generated by contouring water level elevations (referenced to a common datum such
as sea level): for a confined system this represents the potentiometric surface, while for an
unconfined aquifer this represents the watertable surface. As groundwater flows in the direction
of decreasing head, these surfaces provide an indication of the direction of groundwater flow.
As per Darcy’s Law, the rate of flow depends on the hydraulic gradient and the permeability.
The hydraulic properties of an aquifer can be determined through pumping tests and other
methods such as laboratory measurements of aquifer materials.
Recharge to a groundwater system represents the addition of water to the body of water already
stored in the ground. As the volume of water stored is increased, it is reasonable to expect a rise
in water level in the area receiving recharge. This change in water level will only occur when
the recharge water reaches the watertable, taking a few minutes to hours in a shallow
groundwater system within a permeable unsaturated zone, or months to years if the watertable is
overlain by relatively impermeable material or is very deep (Armstrong and Narayan, 1998).
Therefore, the response of an aquifer to recharge processes provides key information about
Hydrogeological Framework
55
aquifer properties in the unsaturated zone. In conceptualising groundwater flow, it is important
to be explicit about the type of recharge model being considered and the associated
assumptions. For example, a one dimensional recharge model assumes that recharge operates
largely in the vertical direction, while a three-dimensional model considers both vertical and
lateral movement of water, including overland flow, shallow throughflow and deeper
groundwater redistribution (Hatton, 1998). Although recharge by rainfall is only one of many
possible causes of water level rise (e.g. water-related activities such as recovery from pumping
in a nearby well or the application of irrigation water can also influence water levels), the
following analyses assume that all rises in the watertable are due to natural recharge events.
This assumption is reasonable given that irrigation is not a common practice in the case study
catchment.
Groundwater moves from areas of recharge to discharge sites under the influence of a hydraulic
gradient. Therefore, an understanding of discharge processes, in addition to recharge, is
important in formulating a conceptual model for groundwater flow. This is particularly relevant
when considering the transport, potential residence time, and ultimate fate of dissolved
constituents that are mobilised in groundwater, as is central to this thesis for dissolved species
of nitrogen.
4.2 GEOLOGIC CHARACTERISATION
The following section describes the depositional environment for the unconsolidated alluvial
sediment of the Herbert River valley, and provides a lithostratigraphic interpretation for the
sedimentary sequences identified. This forms the basis for the defined lateral extent of the study
area, together with identifying confined and unconfined layers and their spatial (vertical and
lateral) relationships.
4.2.1 General depositional environment
Of particular interest to this study is the depositional environment of the sediments that
comprise the Herbert River valley, as this affects the composition, characteristics and spatial
extent of the alluvial aquifers. According to Rienks et al. (2000), sediment was deposited during
the Quaternary Period (i.e. less than 2 million years ago) above a predominantly granitic
basement. The largest structural unit in the area is the Ingham Batholith, which comprises
Carboniferous-Permian igneous rocks of the Kennedy Province. The province represents at least
two episodes of granitoid emplacement and possibly three of felsic volcanism (Rienks et al.,
2000). Granitic and volcanic landforms also crop out above the deltaic sediments (Figure 3-3).
Sedimentation occurred from both terrestrial and marine deposition, with five broad
depositional settings recognised: alluvial-plain, estuarine-plain, strand-plain, dune-field and
marine (Holmes et al., 1991). A bedrock contour map clearly delineates a palaeochannel of the
Chapter 4
56
Herbert River that was probably formed in response to a relative fall in sea level (Figure 4-1).
Importantly, the map also shows that there is a bedrock-high centered on Trebonne Creek,
which defines the southern margin of the palaeovalley. The trend of the Creek and
palaeochannel of the Herbert River approximately parallel the Palmerville Fault, which is
considered to be a major crustal structure of north Queensland (De Keyser, 1963). Major
movement of the fault is thought to predate the late Carboniferous rocks comprising the Ingham
Batholith. As a result, the fault is difficult to recognise on the ground in the study area (Rienks
et al., 2000).
streams
# bores
? inferred fault
bedrock contours
(59)
53
(40)
(27)
(29)29(17)
(27)(23)
+24+19
+12+19+17+6+13 +2
+81 13
1517
+6+7
+1724
22
(38)
(34)25
(33)51
2930(36)
3438
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54(48)
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(66)65
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-10
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0
-2 0
-90-80
-30
-50
Mt Cordelia
-70
5 0 5 Km
-30
-40
Figure 4-1 Bedrock contours in the Herbert River delta referenced to the Australian Height Datum (AHD). Depths to bedrock (negative numbers unless shown with a ‘+’) were converted to AHD by subtraction from the bore elevation, surveyed or estimated from a DEM (enclosed in parentheses). The approximate location of the Palmerville Fault and Mt Cordelia (approximately to scale) are also shown. Source: QDNRW groundwater database (depths to bedrock and surveyed bore elevations); CSIRO (DEM).
4.2.2 Lithostratigraphic interpretation
In order to identify the main lithostratigraphic units of the deltaic sediments, lithologic cross-
sections have been constructed from logs of the bores described by Cox (1979), as well as
selected QDNRW water bores in the study area (Figure 4-2). Bores with complete logs, known
screened intervals and surveyed elevations were selected for the lithostratigraphic interpretation.
The lithologic descriptions have been simplified such that units are distinguished as mud, sand,
sand and gravel, or bedrock (weathered granite). Although sand units vary in their proportion of
Hydrogeological Framework
57
mud, they are nonetheless grouped together as one type based on the dominance of sand.
Delineation of the main lithostratigraphic units is important for understanding the spatial extent
and vertical relationships between the alluvial aquifers. Selected cross-sections and a fence
diagram are displayed in Figure 4-3 and Figure 4-4: bores have been projected onto the section
lines shown in Figure 4-2. Surveyed riverbed and bank heights (estimated from river profiles
provided by the Hinchinbrook Shire Council) and major outcrops are also included on the cross-
sections. For convenience bores are referred to by the last two or three digits in the QDNRW
numbering system, omitting the 116000 prefix.
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%U QDNRW bores#S Cox's bores
River sections$T
5 0 5 Km
Mt Cordelia
Outcrop
Herbert R
Stone R
Trebonne Ck
Palm Ck
boundary ofstudy area
Figure 4-2 Cross-sections constructed from lithological logs of bores in the study area. Outcrops, including Mt Cordelia are shown approximately to scale. Source: bore logs from Cox (1979) and QDNRW groundwater database; river cross-sections from the Hinchinbrook Shire Council.
Interpretation of the lithologic sections is complex because the identified sedimentary units
show considerable interfingering. However, within individual bores, sequences of gravel-sand-
mud can be identified, which are interpreted to represent discrete sedimentary cycles. At least
four sedimentary cycles are recognised, with a unit of mud defining the top of each cycle. In
bores where sedimentary cycles appear to be incomplete, this could either be due to erosion and
reworking or due to a transition in deposition. As depicted in Figure 4-3 and Figure 4-4, lateral
correlation of some units is possible based on the lithologic descriptions and relative depths (in
Chapter 4
58
m AHD); however, where there is insufficient evidence to indicate spatial continuity, units are
represented as lenses.
Interpretation of the cross-sections and fence diagram indicates that from the upper part of the
valley to the coast there is a dominant sedimentary unit that overlies the bedrock in the
palaeochannels. This sand and gravel unit varies in thickness from approximately 15 m to 80 m
at the coast. Lenses of sand containing varying degrees of mud are also present within the unit.
Based on lateral comparisons with adjacent bores, the sand and gravel unit is depicted as
continuous along the section H-I-J-K (Figure 4-4). However, radiating outward from this
section, particularly south of Trebonne Creek, spatial correlations are less apparent. This may
reflect a difference in depositional environment on either side of the Trebonne Creek bedrock-
high (Figure 4-1 and Figure 4-2). Sediment deposited along the middle to upper Stone River is
also lithologically different to that of the Herbert River valley (cross-section J-J', Figure 4-4).
Thick mud layers dominate the sediment in these areas and the sand/gravel units are generally
not as thick or laterally continuous. This is consistent with a different source of sediment and/or
a different depositional environment.
The basal sand/gravel layer in the Herbert River valley is generally overlain by a unit of mud of
varying thickness. In the upper parts of the valley the mud contains a thin lens of sand; further
down-valley additional sand/gravel lenses are apparent to the east of bore 67 (Figure 4-4), that
interfinger in the vicinity of bores 52 and 53 (cross-section C-C', Figure 4-3). It is unclear how
laterally continuous these lenses are to the south of the study area (e.g. between bores 54 and 55
in cross-section C-C', Figure 4-3). Towards the coast there is considerable thickening of the
sedimentary strata compared to the upper reaches of the valley; a change in composition to sand
and sand/gravel dominated units with minor intervening muds is also evident. A distinctive
organic-rich mud is found towards the top of the sequence in the coastal bores (cross-section A-
A', Figure 4-3). It has previously been suggested that this represents a marine mud (Cox, 1979).
Figure 4-3 Representative lithologic cross-sections in the lower Herbert River catchment (continued next page). Refer to Figure 4-2 for cross-section locations.
Hydrogeological Framework
59
cont.
Figure 4-3 Representative lithologic cross-sections in the lower Herbert River catchment (refer to Figure 4-2 for cross-section locations).
Figure 4-4 Fence diagram for the alluvial aquifer system in the lower Herbert River catchment (refer to Figure 4-2 for cross-section locations).
Hydrogeological Framework
61
Based on the litholostratigraphic analysis presented in this research, the sediments of the
Herbert River valley are predominantly comprised of interbedded mud, sand and sand/gravel
sequences. Whilst sedimentary sequences corresponding to different cycles of deposition can be
identified in individual bores, the lateral discontinuity of these layers and interfingering nature
of the strata make it difficult to define distinct aquifer units. In contrast to this interpretation,
Cox (1979) considered that there were four lithologically distinct sand units, as well as two mud
units. From a hydrogeological perspective, the characterisation of discrete aquifer units is only
important if hydraulically significant at the scale of interest. Hydraulic characteristics and water
levels in nested boreholes are analysed below in order to determine the hydraulic significance of
the observed lithological relationships.
4.2.2.1 Relationships with the river
The locations of river cross-sections available along the Herbert River and other tributaries in
the lower catchment are shown in Figure 4-2. The historical minimum riverbed elevation and
maximum riverbank elevation were estimated from surveyed river profiles and included on the
lithologic cross-sections. Figure 4-3 and Figure 4-4 illustrate the depth at which the Herbert
River and the underlying lithology physically intersect. In the upland areas in the northwest of
the study area, the sedimentary profile has been deeply incised by the Herbert River to a depth
of at least 20 m (e.g. cross-section H-I, Figure 4-4). Hence, the riverbed intersects the basal sand
and gravel unit. Further downstream, the depth of incision gradually declines, such that the base
of the riverbed penetrates one of the upper sand lenses, to a depth of approximately 10 m (e.g.
cross-section J-E, Figure 4-4). These observations have implications for the connectivity of
groundwater with surface water, as discussed in detail in Chapters 6 and 7.
4.2.3 Boundary of the study area
It was noted above that the observed lithology of bores located along the southern extent of the
lower catchment (along the upper Stone River and approximately south of Trebonne Creek) was
different to that observed for bores further north, making lateral comparisons difficult (refer to
the most southern bores in cross-sections B-B', C-C', D-D' and J-J', Figure 4-4). In addition,
lithological logs for some of the southern bores are absent from the database or poorly
described. For these reasons, the southern boundary of the study area is defined by Trebonne
Creek and the middle Stone River valley in the hydraulic analyses that follow. Disregarding the
southern/southwestern bores is considered reasonable given the emphasis in this thesis on
groundwater interactions with the Herbert River. The approximate extent of the study area for
the purposes of this research is depicted in Figure 4-2.
Chapter 4
62
4.3 HYDRAULIC PROPERTIES OF THE AQUIFERS
No pumping tests were performed as part of this research. However various tests were
previously carried out by Calvert (1959) and Cox (1979) in the alluvial sediments of the Herbert
River valley. The tests of Calvert varied from the testing of properly screened bores to testing of
open bores cased without screens; recovery readings were used to determine transmissivity. In
contrast, the tests performed by Cox were more comprehensive, comprising measurements of
drawdown and recovery over a period of 24 hours (constant discharge tests) to determine the
aquifer characteristics. Cox defined four alluvial aquifers, referred to as S1, S2, S3 and S4 (from
deepest to shallowest). The results from the pumping tests are summarised in Table 4-1 based
on Cox’s terminology; Cox’s interpretation of the data are summarised below. In order to
distinguish aquifer units based on hydraulic properties, the existing pumping test data are also
assessed in the context of the lithostratigraphic interpretation presented as part of the current
study (Section 4.2.2).
Table 4–1 Summary of hydraulic characteristics of the water-bearing alluvial stratigraphic units in the Herbert River valley, after Cox (1979). A summary of transmissivity values from Calvert (1959) is also provided.
Stratigraphic unit
Aquifer Number of testing sites+
Hydraulic conductivity*
(m/day)
Transmissivity (m2/day)
Cox
Transmissivity (m2/day) Calvert
shallow S4 2 61 - 86 400 - 578 1634-6530
intervening S3 1 2.9 25 170-6114
intervening S2 1 13.9# 117 6-63
basal S1 4 0.5 - 50 50 - 1675 64
+ Cox’s tests * Calculated values by Cox # Results considered by Cox to be of low credibility
4.3.1 Summary of Cox’s interpretation
Based on the pumping tests, Cox concluded that the deepest aquifer (which coincides with the
basal sand and gravel unit described in Section 4.2.2) is mainly confined, but semi-confined in
the upper parts of the valley. The hydraulic characteristics of the aquifer were considered to be
generally uniform, except for in the upper part of the valley where the permeability of the
sediments is higher. The S2 aquifer was interpreted to be confined, although discontinuous and
of limited areal extent. Cox also considered the S3 aquifer to be confined as there was no
evidence of leakage from overlying or underlying units; the transmissivity was observed to be
Hydrogeological Framework
63
variable. Cox suggested that the S4 aquifer represented an unconfined shallow aquifer of
consistently high to very high transmissivity.
Cox summarised that units S1 and S3 were more permeable on their western margins, with
permeability in each unit decreasing towards the coast due to the increasing mud content to the
east. In terms of vertical variation through the sedimentary sequence, it was noted that with the
exception of the up-valley area, permeability in the S1 unit was low and a little higher in unit
S2. In unit S3 the permeability was substantially higher again, with isolated areas of very high
values. The upper S4 aquifer was considered to be very permeable, although infiltration was
slow due to the veneer of fine grained material covering the more permeable sand.
4.3.2 Interpretation based on lithostratigraphy
In the absence of lithological logs for several bores for which pumping tests were performed, it
is not possible to reinterpret some of the test results. Of the bores with known lithologies, the
results indicate that the hydraulic properties of the aquifer sediments are quite variable, even
within the stratigraphic units defined by Cox. This is most likely a reflection of the complex
stratigraphy in the Herbert River valley typical of fluvial environments. Based on the
lithostratigraphic interpretation presented in Section 4.2.2, the S1 bores of Cox are screened
within the basal sand and gravel layer, while the other bores are screened in the overlying
sand/gravel units (S2-S4). From the available pumping test data it is difficult to delineate
hydraulic characteristics which typify a particular stratigraphic unit; however, in general, bores
screened in the shallower sand/gravel layers have a higher hydraulic conductivity and hence
transmissivity compared to the deeper screened intervals. The exception is in the upper part of
the valley, where the transmissivity in the deep aquifer is high. This is in agreement with Cox
(Section 4.3.1 and Table 4-1). Only two of Cox’s pumping tests were at nested sites; the results
for these tests also highlight the contrast in hydraulic properties between the deepest and
shallowest sedimentary units. However, given the limited number of pumping tests over the
study area and lack of conclusive data, it is considered that further evidence is required in order
to confirm or eliminate the existence of separate confined/unconfined aquifers. Analysis of
water level data (Sections 4.4 and 4.5) and hydrogeochemistry (Chapter 5) will assist in the
characterisation of the alluvial aquifer system in the study area.
4.4 VERTICAL FLOW IN THE SUBSURFACE
Having explored the lithostratigraphic relationships of the sediments in Section 4.2, the aim of
the following sections is to conceptualise how water moves through the subsurface. Water level
data are interpreted in order to describe recharge and discharge processes in the alluvial aquifers
of the Herbert River valley and establish the potential routes along which groundwater flows
between these areas. Bore hydrographs and water level contours are examined in order to
Chapter 4
64
characterise the vertical and lateral connectivity between aquifers. The resulting conceptual
framework for recharge, discharge and flow through the aquifer system is validated and
enhanced in Chapter 5 through the interpretation of hydrogeochemical data.
The key characteristics related to vertical flow in the subsurface include: (i) the degree of
vertical aquifer connectivity; (ii) direction of hydraulic gradient between the aquifers; and (iii)
recharge-discharge dynamics. In order to examine these attributes, bore hydrographs are plotted
for nested monitoring bores in the study area based on the historical record of groundwater level
measurements. Bore hydrographs reflect local groundwater behaviour, which will vary
throughout an aquifer in response to local variations in hydraulic properties and distance to
recharge/discharge sites. Importantly, the analyses of hydrographs provide a basis for
identifying discrete aquifers, based on hydraulic behaviour as opposed to lithostratigraphic
units, which may not necessarily correspond to individual aquifers. Nevertheless, water level
behaviour between nested bores has been interpreted within the lithostratigraphic framework,
using the limited spatial extent of monitoring bores. An additional limitation with the bore
hydrograph analysis is that, whilst groundwater is found at various depths in the stratigraphic
profile, the majority of nested bores are screened at only two of those depths in any location:
generally the shallowest and deepest intervals. Therefore, the interpretation is biased towards
the behaviour between the upper and basal sandy units. The analyses in this section build on the
hydraulic characteristics evident from the pumping test data (Section 4.3).
4.4.1 Data preparation
Groundwater level (head) data were extracted from the QDNRW database for all registered
bores (approximately 200) in the study area (Figure 4-2). Of these bores, only around 150 have
historical water level data (of varying length of record), with stratigraphic and screened depth
information at 100 of these bores. Water level records date back to 1967, however some of these
early bores were abandoned and the longest records are generally from 1976 to the present.
Monitoring of water levels averaged four times per year in the 1970’s, but has since declined to
about twice per year: once before the wet season (October-November) and once soon after the
wet season (March-April-May). A limited number of bores have also been monitored daily
using data loggers.
In the database, depth to groundwater and elevation of the bore (in m AHD) are recorded from
the top of the bore casing, known as the reference point. In the absence of an elevation record
for a particular bore, an estimate was made using a 100 m digital elevation model (DEM) grid
(Error! Reference source not found.). Groundwater elevations (in m AHD) were obtained by
subtracting the depth to groundwater from the bore elevation. Screened intervals have mostly
been recorded in the database as depth to the bottom of the slots: in general the slotted interval
extends 3 m above this level (QDNRW technical staff, pers. comm.). Nested monitoring bores
Hydrogeological Framework
65
comprise multiple pipes (within the same bore) that are screened in different water-bearing
units. The deepest pipe is denoted as the ‘A’ pipe, while ‘B’ and ‘C’ refer to progressively
shallower pipes.
Figure 4-5 Digital Elevation Model (DEM) for the lower Herbert River catchment. Cox’s bores are shown for reference. Source: CSIRO.
Many groundwater monitoring networks similar to that in the study area are read manually and
infrequently (monthly/seasonally), with the objective of measuring the water levels when they
are at their highest and lowest. Although infrequent monitoring does not guarantee to pick the
true maximum and minimum levels, it does provide an idea of the general range of fluctuations
(Armstrong and Narayan, 1998). Illustrated in Figure 4-6 are bore hydrographs for shallow and
deep bores based on groundwater levels measured by an automatic recorder (daily records but
with a significant number of gaps) and manually (from piezometers screened in the same
intervals and located at the same site as each of the recorders) to highlight the difference in bore
hydrographs obtained by the two methods of measurement. Clearly, the high temporal
resolution of the continuous data is useful for analysing dynamic processes such as the rainfall
response of the aquifers, particularly in the wet season. However, these bores are limited in
spatial extent compared to the majority of bores in the catchment which have historically been
sampled manually. It is considered that due to the strong seasonal contrast in groundwater
elevations, this lower temporal resolution data is still adequate for establishing inter-aquifer
relationships in the area.
Chapter 4
66
6
8
10
12
14
16
1978
1979
1980
1981
1982
Gro
undw
ater
ele
vatio
n (m
AH
D)
0
100
200
300
Rai
nfal
l (m
m)
61B 80A 79A 61A rainfall 32091
#S
#S
#S#S
78A, 69A
79A, 61A80A, 61B
Herbert R
75A, 46B
4
6
8
1978
1979
1980
1981
1982
Gro
undw
ater
ele
vatio
n (m
AH
D)
0
100
200
300Ra
infa
ll (m
m)
46B 75A rainfall 32045
12
14
16
18
20
22
24
1978
1979
1980
1981
1982
1983
1984
1985
1986
Gro
undw
ater
ele
vatio
n (m
AHD
)
0
100
200
300
Rain
fall
(mm
)
69A 78A rainfall 32091
Figure 4-6 Historical groundwater elevations for bores sampled automatically (labelled) and manually (dots). Groundwater elevations are from the shallowest (80A, 61B, 75A, 46B) and deepest (79A, 61A, 78A, 69A) water-bearing units. Daily rainfall is from sites closest (up-valley) to the monitoring bores. Source: QDNRW (historical groundwater depths) and BoM (rainfall). Refer to inset map for general bore locations in the catchment.
shallow
shallow
deep
deep
automatic
automatic
automatic
Hydrogeological Framework
67
4.4.2 Inter-aquifer connectivity
Nested bores that have identical hydrographs can be considered to have strong vertical
connectivity. Conversely, where there is head separation between nested intervals, retarded
vertical flow between semi-confined units (in-phase hydrographs) or poor vertical connection
due to intervening confining layers (different hydrograph patterns) is suggested. Note that
similar hydrograph patterns may also occur for reasons other than connectivity; therefore,
reference to bore logs is also important where possible. The vertical head gradient (potential
direction of vertical flow) is deduced by comparing the relative groundwater elevations between
screened intervals. A downwards pressure results if the groundwater elevation for the shallower
screened interval is greater than that of the deeper interval, reflecting a positive head gradient.
Conversely, a negative head gradient indicates an upwards potential. The hydrographs of 12
nested bores screened in the shallowest and deepest water-bearing stratigraphic units are
analysed below based on visual inspection. These nested bores were selected for analysis on the
basis of the length of historical record (at least 10 years) and the presence of bore logs and
screened intervals at known depths. A map of bores referred to in this section is provided in
Figure 4-7, including a spatial representation of the degree and direction of vertical hydraulic
connection between the shallowest and deepest aquifers. Inter-aquifer relationships are further
examined in Chapter 5 through the analysis of hydrochemical data. Hydrographs representing
the main patterns observed in the study area are provided in Figure 4-8 and Figure 4-9.
#S
#S
#S
#S
#S
#S
#S
#S
#
#S#S
#S
#S68
62
6160
53 46 (74)
38
36
48
51 (101)
49
Herbert R47
(54)
#S
#S
Good (downwards potential)#S
Strong (indeterminant)#S
Vertical hydraulic connectivity
Poor (downwards potential)
Poor (upwards potential)
Strong (upwards potential)#S
5 0 5 10 KmN
Figure 4-7 Summary of the degree of vertical hydraulic connectivity (strong, good, poor) and direction of head gradient between the shallowest and deepest aquifers. Note that the locations of bores 54, 101 and 74, screened in an intervening sand unit and the shallowest and/or deepest aquifers, are also shown.
Chapter 4
68
8
12
16
20
24
2819
7619
77
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)68A (-21.5 m AHD) 68B (13.3 m AHD)
#S#S
#S68
6160
Herbert R
6
10
14
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)
61A (-36.1 m AHD) 61B (8.3 m AHD)
11
12
13
14
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)
60A (-27.7 m AHD) 60B (-4.2 m AHD)
Figure 4-8 Hydrographs for nested monitoring bores screened in the deepest (A-pipe) and shallowest (B-pipe) water-bearing units in the upper half of the catchment. The elevation of the base of the screened interval in each pipe is also indicated in parentheses. Source: QDNRW (historical groundwater depths). Refer to inset map for general bore locations in the study area.
shallow
deep
deep
shallow
shallow
deep
Hydrogeological Framework
69
4
6
8
10
12
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)
53A (-76.8 m AHD) 53B (3.5 m AHD)
#S
#S
#S53
48
36
Herbert R
0
1
2
3
4
5
6
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AH
D)
48A (-52.6 m AHD) 48B (-2.2 m AHD)
0.0
0.5
1.0
1.5
2.0
2.5
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)
36A (-92.7 m AHD) 36B (-5.9 m AHD)
Figure 4-9 Hydrographs for nested monitoring bores screened in the deepest (A-pipe) and shallowest (B-pipe) water-bearing units in the lower half of the catchment. The elevation of the base of the screened interval in each pipe is also indicated in parentheses. Source: QDNRW (historical groundwater depths). Refer to inset map for general bore locations in the study area.
shallow
deep
shallow
deep
shallow
deep
Chapter 4
70
Based on the available data, nested bores in the upper half of the catchment display in-phase
hydrograph patterns (Figure 4-8). The similarity in patterns between nested intervals indicates
good vertical connectivity; however, the head separation at bores 68 and 61 suggests that the
screened intervals are separated by a semi-confining layer and/or have significantly different
hydraulic properties such that vertical flow, although not inhibited, is retarded. The positive
head gradient indicates a downwards potential between the shallow and deep intervals during
both wet and dry periods. Bore logs illustrate the presence of a mud unit of approximately 8 m
in thickness underlying the shallow interval in these bores (Figure 4-4). In addition, the
pumping test results (Section 4.3) indicated that the transmissivity of the shallowest
stratigraphic unit far exceeds that of the deeper unit. Therefore, the observed head separation is
considered to be the result of an increase in storage in the shallow aquifer due to the slight
impediment of downward flow resulting from the lower permeability of the underlying
sediments. The magnitude of head separation declines from the northwest towards the centre of
the catchment, such that at bore 60 the nested hydrographs are near-coincident and indicate
strong vertical connection. These bore hydrograph interpretations are supported by
hydrochemical evidence presented in Chapter 5.
Nested hydrographs in the lower half of the catchment display different trends from bores
screened in the similar shallow and deep sandy units up-valley (Figure 4-9). South of the
Herbert River, in the vicinity of bores 53 and 46 (not shown), the hydrograph of the deeper
interval represents a subdued trend of the hydrograph for the shallower interval. The relative
groundwater elevations indicate a downwards head gradient. The broad similarity in
hydrographic response between the shallow and deep intervals suggests poor vertical
connectivity. Retardation of vertical flow is not unexpected in this area given the presence of a
thick mud unit (e.g. 30 m thick at bore 53) in the middle of the sedimentary profile (e.g. cross-
section C, Figure 4-3). The significance of the observed poor vertical connection is that recharge
to the deep aquifer is likely to be from lateral sources rather than via the shallow aquifer.
The hydrographs of nested bores to the north and northeast are characteristically non-parallel
and sometimes coincide, in contrast to the bores further up-valley which display parallel
hydrographs and a consistent seasonal head separation. Importantly, some bores displaying this
type of nested hydrograph pattern have a dominantly downwards head gradient (e.g. bore 51,
Figure 4-7), while further east the dominant direction of flow is upwards (e.g. bores 48, 49, 36,
38, Figure 4-7). The non-parallel hydrograph patterns between the shallowest and deepest
screened intervals at bore 51 indicate poor downwards vertical connectivity. This is supported
by bore log information, which highlights a 26 m thick mud unit below the shallow aquifer
(cross-section C-C', Figure 4-3). Bore logs illustrate the presence of mud units between 9-12 m
in thickness at bores 48 and 49, and 4-7 m at bores 36 and 38 below the shallowest interval
(cross-sections B-B' and A-A', Figure 4-3).
Hydrogeological Framework
71
Although these nested bores generally display a head separation, as a consequence of the
intervening muds, there is similarity in hydrographic response between screened intervals
(Figure 4-10). However, whilst the hydraulic information is consistent with good upwards
vertical connection, it is shown in Chapter 5 from hydrochemical evidence, that there is a degree
of confinement between the deep and shallow aquifers at bores 48 and 49 and hence poor
connectivity. Strong upwards vertical connectivity is however demonstrated using
hydrochemistry at bores 36 and 38, indicative of a large component of vertical discharge from
HSd to HSs (Section 5.3.4, Chapter 5).
1
2
3
4
5
6
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Ele
vatio
n (m
AH
D)
A-p
ipe
0
1
2
3
4
5
648A (-52.6 m AHD) 48B (-2.2 m AHD)
Ele
vatio
n (m
AH
D)
B-p
ipe
1.5
2.0
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)A-
pipe
0.0
0.5
1.0
1.5
2.036A (-92.7 m AHD) 36B (-5.9 m AHD)
Ele
vatio
n (m
AHD
)B-
pipe
Figure 4-10 Similarity in hydrographic response for selected nested bores displaying an upwards potential. Note the different vertical scales for the A and B pipes in each bore.
shallow
deep
shallow
deep
Chapter 4
72
As depicted in Figure 4-4, east of cross-section J-J' there are additional sand units between the
shallowest and deepest aquifers. Three nested bores in the study area, bores 54, 46 and 101, are
screened in one of these intervening sands, in addition to the shallowest and/or deepest aquifers
(cross-sections B-B' and C-C', Figure 4-3). Hydrograph analysis of bore 54 (A and B pipes)
indicates that there is a slight impediment to vertical flow between the nested sand units, due to
separation by a 2 m thick mud unit. However, the similarity in hydrograph patterns indicates a
high degree of vertical connectivity between the two upper sand units (Figure 4-11). Bores 46
(with co-located bore 74) and 101 are screened in the two lower sand units, as well as in the
shallowest aquifer. Note that data from bore 101 is analysed rather than co-located bore 51 (bore
101 is the replacement of 51), as 51 is not screened in an intervening sand unit. Analysis of bore
46 highlights the close similarity in hydrographic response between the deepest and intervening
sand units (compare A pipe and bore 74 in Figure 4-11). Whilst these units are separated by at
least 15 m of mud, the head separation is less than 40 cm. Thus, it is considered that vertical
connectivity is high, particularly in comparison to the shallowest water-bearing unit (B-pipe),
which has a different hydrograph pattern (compare A and B pipes in Figure 4-11). Similarly,
analysis of bore 101 highlights the difference in hydrograph pattern between the shallowest unit
(C-pipe) compared to the two lower sand units (A and B pipes). These deeper units display
almost identical hydrographs (maximum of 10 cm head separation); however, it is noted that
there is a consistent upwards pressure between the sand units due to a thin intervening mud
layer (Figure 4-11).
Based on hydrograph analysis of the limited nested bores screened in the intervening sand units,
it is concluded that there is limited retardation of vertical flow between the water-bearing sand
units in the lower stratigraphic layers. This is consistent with their being overlap in the
hydraulic properties of Cox’s S2 and S1 aquifers (Table 4-1). In comparison, the degree of
confinement is larger between the two upper sand units, consistent with the differences in
hydraulic properties between Cox’s S3 and S4 aquifers. However, as noted in Section 4.2.2, the
characterisation of discrete aquifers is only important at the scale of interest. Therefore, in terms
of large scale processes, the greatest hydraulic differences are observed between the shallowest
and deepest water bearing units; the intervening sand units are hence considered as components
of the shallowest (bore 54) or deepest (bores 46 and 101) aquifers. This simplification is also
supported by hydrochemical evidence presented in Chapter 5 (Section 5.3.3).
Hydrogeological Framework
73
7
8
9
10
11
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)
54A (-15 m AHD) 54B (3.7 m AHD)
shallow
intermediate
#S
#S
#S
101 (51)
54
Herbert R
46 (74)
2
4
6
8
1976
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
Elev
atio
n (m
AHD
)
46A (-89 m AHD) 74 (-48 m AHD) 46B (-0.3 m AHD)
shallow
intermediate
deep
2
3
4
5
6
1993
1995
1997
1999
2001
2003
Ele
vatio
n (m
AHD
)
101A (-64 m AHD) 101B (-36 m AHD) 101C (1 m AHD)
intermediate
shallow
deep
Figure 4-11 Hydrographs for nested monitoring bores screened in an intervening sand unit (green) as well as in the deepest (blue) and/or shallowest (red) water-bearing units. The elevation of the base of the screened interval in each pipe is also indicated in parentheses. Source: QDNRW (historical groundwater depths). Refer to inset map for general bore locations in the study area. Note that bore 101 is the replacement of bore 51, while bores 46 and 74 are co-located.
Chapter 4
74
4.4.3 Rainfall response
Visual comparison of available bore hydrographs with historical rainfall indicates that
groundwater elevations display a distinct seasonal pattern of maximum head during the wet
season declining to a minimum during the dry season. Hydrographs for automatically sampled
bores (daily data but with gaps) screened in the shallow interval, represented by bores 80A and
75A, display a rapid response to rainfall (Figure 4-6). Given the shallow depth to water, rapid
recharge response and highly fluctuating water level, the shallow screened unit is considered to
represent an unconfined watertable aquifer. Whilst a similar hydrograph pattern is noted for the
basal aquifer, represented by bores 78A and 79A (Figure 4-6), there is less fluctuation in water
levels compared to the watertable aquifer. This observation is consistent with a longer time for
rainfall to recharge the deeper aquifer. The more subdued hydrograph pattern of the deep
aquifer is particularly evident when comparing the hydrographs of co-located deep bore 79A
with shallow bore 80A (Figure 4-6). Note that this difference in hydrographic response between
aquifers was not apparent when comparing the nested hydrographs of an infrequently sampled
bore at the same location (A and B pipes of bore 61, Figure 4-8). This highlights the importance
of continuous data for establishing recharge dynamics.
Cross-correlation analysis (Croke, in prep.) of groundwater levels in co-located bores 79A and
80A (screened at -36.1 and 8.3 m AHD, respectively) with rainfall clearly shows that there is a
rapid groundwater level response (zero lag) to rainfall in the shallow aquifer, consistent with a
shallow watertable aquifer (Figure 4-12).
-0.2
0
0.2
0.4
0.6
0.8
1
-20 -10 0 10 20
Lag (daily time step)
Corr
elat
ion
coef
ficie
nt
autocorrelation of rainfall
cross-correlation of groundwaterlevel (deep aquifer) with rainfall
cross-correlation of groundwaterlevel (shallow aquifer) with rainfall
rapid groundwater level response at
zero lag
rapid drainage from shallow aquifer
Figure 4-12 Cross-correlation of groundwater levels against rainfall at gauge 32091. The autocorrelation of rainfall provides a reference for the lag in groundwater level response. The deep and shallow aquifers are represented by co-located bores 79A and 80A, respectively (refer to Figure 4-6 for bore hydrographs).
Hydrogeological Framework
75
However, there is almost no correlation between groundwater levels in the deep aquifer and
rainfall at zero lag. This suggests that unlike the shallow aquifer, the deep aquifer does not
instantaneously respond to rainfall events. Furthermore, the analysis indicates that drainage
from the shallow aquifer after a rainfall event is rapid (around 2 days).
Stage height data (at gauge 116001, approximately 7 km down-valley from the bores), used as a
surrogate for rainfall, gives a better correlation with groundwater levels at non-zero lag. This is
possibly due to the smoothing effect of catchment response in a stream compared to much
higher frequencies and gaps in the rainfall signal. Cross-correlation analysis of stream stage and
groundwater levels highlights a strong wet season-dry season signal (Figure 4-13). In addition
to the instant rainfall response observed in the shallow aquifer (Figure 4-12), a delayed response
is evident in both aquifers relative to streamflow, as shown by the cross-correlation peaks for
each aquifer at approximately 20 and 40 days for the shallow and deep aquifers, respectively
(Figure 4-13). This reflects delayed recharge to the aquifers, such as from lateral flow. Whilst
the effect of a streamflow event persists in both aquifers for longer than in the river, which
drains more rapidly, there is greater persistence in the basal aquifer, as shown by broadening of
the deep aquifer trend after the peak response. This indicates a lag in drainage of the deep
aquifer compared to the shallow aquifer, consistent with differences in aquifer properties.
-0.6
-0.4
-0.2
0
0.2
0.4
0.6
0.8
1
-500 -400 -300 -200 -100 0 100 200 300 400 500
Lag (daily time step)
Cor
rela
tion
coef
ficie
nt
autocorrelation of river stage
cross-correlation of groundwater level(shallow aquifer) with river stage
cross-correlation of groundwater level(deep aquifer) with river stage
peak at ~40 days
peak at ~20 days
aquifer drainage
Figure 4-13 Cross-correlation of groundwater levels against stage height at gauge 116001 (shown as a 14-day moving average). The autocorrelation of stream stage provides a reference for the lag in groundwater level response. Deep and shallow aquifers are represented by bores 79A and 80A, respectively (refer to Figure 4-6 for bore hydrographs).
Chapter 4
76
4.4.4 Classification of the aquifers
Based on the interpretation of hydraulic characteristics (pumping tests and bore hydrographs) as
well as lithostratigraphic units, it is concluded that the alluvial subsurface of the lower Herbert
River valley can be conceptualised as a two-aquifer system. The shallow, or upper aquifer,
represents an unconfined watertable aquifer, while the deep, or basal aquifer, is semi-confined.
The alluvial aquifers are bounded at depth by weathered granitic bedrock. Although Cox (1979)
distinguished additional confined units, it is considered that based on the available hydraulic
information, the intervening water-bearing units can be lumped together with the upper and
basal aquifers. Thus, a simplified interpretation is considered reasonable. In the west, the
shallow aquifer extends to a depth of approximately 15 m below the surface, while in the east,
the depth varies between 10 m to at least 35 m below ground level. Although the alluvial aquifer
system is conceptualised to have two aquifers, the degree of vertical hydraulic connection varies
spatially. Given that classification of the aquifers is based on a hydro-stratigraphic
interpretation, the aquifers distinguished as part of this study are henceforth referred to as HSs
(shallow aquifer) and HSd (deep aquifer).
4.5 LATERAL FLOW IN THE SUBSURFACE
Potentiometric surfaces and watertable maps have been constructed in this section to establish
an understanding of lateral groundwater flow through the HSd and HSs aquifer systems. These
analyses build on the recharge-discharge relationships that were developed in the previous
section based on bore hydrograph interpretation. In addition, groundwater level differences
between consecutive dry and wet periods are contoured to illustrate the seasonal response of the
aquifers to rainfall and hence identify potential areas of recharge. Based on the availability of
groundwater level measurements across a number of bores, representative months during 1976-
77 have been selected for the analysis. The wet season in 1977 corresponds to one of the
historically largest flooding events in the catchment (Johnson and Murray, 1997) and therefore
the recharge response is more clearly evident during this period.
Hydrogeological Framework
77
4.5.1 Deep aquifer system
4.5.1.1 Flow pattern
Potentiometric surfaces show very similar patterns during consecutive wet and dry seasons
(Figure 4-14). Although the potentiometric surface in the deep aquifer system is higher in the
wet season than the dry, the general direction of lateral groundwater flow throughout the year is
from the upper parts of the Herbert River and Stone River valleys towards the coast
(groundwater flow is perpendicular to the contours). The potentiometric surfaces also indicate
that in the dry season there is a preferential flowpath in the vicinity of bore 48 (Figure 4-2) that
has a northeasterly trend. Notably, groundwater flow is towards the Herbert River along a large
section of the river: this has implications for the relationship between groundwater and surface
water systems, as discussed in Chapters 6 and 7. The preferential flowpath depicted in Figure
4-14 (flowline 1) that parallels the western half of the Herbert River, departs from the river in
the middle of the catchment and thereafter follows an eastward trend towards the coast. The
geological control imposed on this flowpath is clearly evident with reference to the bedrock
contour map (Figure 4-1); in particular, the movement of groundwater through the palaeovalley.
In addition, a flow divide is apparent in the south, such that the direction of groundwater flow in
the deep aquifer (e.g. flowline 2) is away from the Herbert River; this may indicate that
groundwater flow is also affected by the bedrock high (Figure 4-1).
4.5.1.2 Recharge-discharge characteristics
Based on the analysis of changes in piezometric level between the wet season and the end of the
previous dry season, there is a major recharge zone for the deep aquifer system in the upper,
northwestern part of the catchment (Figure 4-15a). This zone, in the vicinity of bores 69 and 70
(Figure 4-2), is where the greatest rise in piezometric level is observed in the study area (up to 9
m by March 1977). Less recharge is observed in the southwest, along the upper Stone River
valley (1.8 m by March 1977), which influences water levels in bores along the Stone River and
further to the east. Recharge also occurs in the northeast (1-3 m rise, shown in Figure 4-15a), in
the vicinity of bores 48 and 49 (Figure 4-2). In the months following the peak of the 1977 wet
season, the dominant recharge front shifts further down the Herbert River valley as water levels
decline and ultimately approach their dry season values. The volume of recharge to the deep
aquifer has previously been estimated at 540 megalitres (Cox, 1979).
Chapter 4
78
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#
#
##
#
#
#
#
#
#
#
#
#
#
## 28 24 20 16
1.8
2.5
2.2
3.4
5.0
5.4
3.6
3.7
2.0
1.0
3.2
5.9
6.0
5.67.6
8.0
17.3
10.1
12.0
12.5
19.318.0
24.0
10.5
15.426.131.9
261022 1826 14
1632
N
Stream
Groundwater elevationcontour
5 0 5 Km
November 1976
# Groundwater elevation
Outcrop
a
#
#
#
#
#
#
#
#
#
#
#
#
##
#
#
#
#
#
#
#
#
#
#
## 32 28
1.9
2.5
2.6
3.6
5.7
4.0
3.9
3.0
3.9
3.8
6.4
6.3
7.1
9.9
17.8
10.8
12.9
13.020.0
19.0
25.8
11.5
14.9
24.434.5
35.2
261022 1826 14
24
N
Stream
Groundwater elevationcontour
5 0 5 Km
March 1977
# Groundwater elevation
Outcrop
b
Figure 4-14 Potentiometric surfaces (m AHD) for the deep aquifer system during (a) the dry season of 1976 and (b) consecutive wet season of 1977. The direction of groundwater flow is shown by selected flowlines, with flowlines 1 and 2 labelled. Source: QDNRW (groundwater depths)
1
2
1
2
Hydrogeological Framework
79
The association of the northwest and southwest recharge areas with the topographically elevated
parts of the catchment (Figure 4-5) suggests that lateral recharge to the deep aquifer is sourced
from rainfall towards the catchment boundary (Cardwell Range and Seaview Range, Figure
4-15b) during the wet season. For the recharge zone in the northwest, this is consistent with the
observed correlation between rainfall and groundwater response illustrated in Figure 4-6 at bore
78. In addition, a local recharge area associated with the Mount Leach Range to the north of the
river is evident (Figure 4-15b). Strong vertical connectivity between the aquifers was also
observed in this area (Section 4.4.2), consistent with vertical leakage to the deep aquifer. For the
identified recharge zone in the northeast, the upwards head gradient suggests that recharge to
the deep aquifer is unlikely to be from direct rainfall; however, a combination of recharge from
the southwest and west via lateral flow is plausible (refer to flowlines in Figure 4-14a).
Visual inspection of the potentiometric surfaces for the deep aquifer highlights a decrease in the
hydraulic gradient (increase in spacing between equipotential lines) nearer to the coast (Figure
4-14), indicative of lower hydraulic pressure and hence faster lateral movement of groundwater
due to an increase in hydraulic conductivity. This is consistent with lithological evidence at the
coast, whereby thinning (or absence) of mud units is observed at depth (cross-section A-A',
Figure 4-3). Given the direction of lateral groundwater flow and reversal in vertical head
gradient towards the coast, these observations are consistent with groundwater discharge. Whilst
discharge from the deep aquifer is ultimately out to sea, it was noted in Section 4.4.2 that a
component of vertical leakage from the deep to the shallow aquifer is also plausible towards the
coast. Groundwater pumping is generally from the shallow rather than the deep aquifer and
therefore is unlikely to contribute to losses from the groundwater system. However, given the
deep incision of the Herbert River into HSd in the northwest (Section 4.2.2.1) and the
groundwater flow pattern, this raises the possibility of groundwater discharge from the deep
aquifer to the river. This relationship is explicitly examined in Chapters 6 and 7.
Chapter 4
80
#
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#
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#
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##
#
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##
0.1
0.1
0.3
0.2
0.5
0.4
0.3
0.2
1.0
3.0
0.7
0.5
0.20.7
1.5
2.3
0.50.7
0.9
3.5
4.4
9.08.53.3
0.89 7
3
1.52 1 0.5
68
21
N
Stream
Head differencecontour
5 0 5 Km
March 1977-November 1976
# Water level difference
Outcrop
a
N10 0 10 20 Km
Catchment boundaryStreamHead difference contoursOutcropCardwell Range
Seaview R
ange
RECHARGEZONE
RECHARGEZONE
LOCALRECHARGE
DISCHARGEZONE
#
Mount LeachRange
b
Figure 4-15 Head difference contours (m AHD) for the deep aquifer system, comparing groundwater levels during the 1977 wet season with previous dry season levels in November 1976. An enlarged area of the catchment is shown in (b): the main recharge areas associated with the mountain ranges and the coastal discharge zone are highlighted.
Hydrogeological Framework
81
4.5.2 Shallow aquifer system
Watertable maps have been constructed for the shallow aquifer system for selected months
during the dry and wet seasons in 1976-1977 (Figure 4-16). Due the scarcity of monitoring
bores, data from private wells were used in conjunction with monitoring bore data. Manual data
selection from private wells were based on comparisons with groundwater levels recorded in
surrounding monitoring bores. Manual selection was necessary because of the absence of
information on the total depth of some wells: groundwater levels that did not appear consistent
with surrounding bore measurements were discarded. Therefore, in the first instance, contours
were based on water levels in bores with known screened depths, with gaps filled in using
private well data. The addition of the private well data resulted in more detailed flow patterns
downstream of the junction of the Stone River with the Herbert River. Although the watertable
contours are based on a degree of interpretation, the flow patterns observed in Figure 4-16 are
plausible based on the corresponding maps constructed for the deep aquifer system, particularly
where there is good vertical connectivity (compare with Figure 4-14).
4.5.2.1 Flow pattern
Although the watertable is higher in the wet season compared to the dry, the groundwater flow
pattern in the shallow aquifer system is essentially the same in both seasons (Figure 4-16). In
general, it is observed that groundwater flows either from the west towards the Herbert River
and thereafter to the coast, or from the west to the east/southeast towards the coast. As for the
deep aquifer system, a preferential flowpath coincident with the Herbert River is evident;
however, in the centre of the catchment the flowpath is deflected to the northeast towards bore
48 (Figure 4-2).
4.5.2.2 Recharge-discharge characteristics
As for the deep aquifer system, water level difference maps have been constructed for the
shallow aquifer system to compare groundwater levels in the 1976 dry season (November) to
the subsequent wet season in 1977. Comparison of wet season versus dry season groundwater
levels indicates a major recharge area in the northwest: from November 1976 to February 1977
water levels in bore 68B increased by 6.9 m (Figure 4-17a), and by 7.2 m in March (not shown).
In the months following the peak of the wet season, groundwater levels in the main recharge
area declined, while levels in bores further down valley increased in response to lateral
movement of the northwesterly recharge front. Dry season water levels were approached again
by September (not shown).
Chapter 4
82
#
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24612 1018 8
5.5
3.0 2.1
1.01.0
0.5
1.9
1.4
5.84.4
0.8
0.72.8
3.3
9.2
9.2
8.4
8.7
9.9
7.08.7
7.2
6.7
3.9
3.6
1.0
1.1
3.9
0.6
6.9
16.0 10.1
11.510.2
17.6
15.3
13.012.4
12.8
14.3
1416
N
Stream
Groundwater elevationcontour
5 0 5 Km
November 1976
# Groundwater elevation
Outcrop
a
#
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# #
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24612 1018 81416
7.0
4.4 3.2
2.52.3
1.6
3.0
2.8
6.95.5
3.8
2.44.5
4.9
9.0
8.19.7
8.7
8.9
5.22.5
2.55.3
9.7
17.8 10.8
10.210.5
10.8
13.2
12.1
24.8
12.4
20.114.5
15.7
16.1
16.1
20222426
N
Stream
Groundwater elevationcontour
5 0 5 Km
March 1977
# Groundwater elevation
Outcrop
b
Figure 4-16 Watertable maps (m AHD) for the shallow aquifer system during (a) the dry season of 1976 and (b) consecutive wet season of 1977. Groundwater elevations are from both monitoring bores and private wells. The direction of groundwater flow is shown by selected flowlines. Source: QDNRW (groundwater depths)
Hydrogeological Framework
83
The watertable difference contours also indicate potential small areas of local recharge to the
east (adjacent to the river), whereby water levels in some bores increased by around 3 m from
November 1976 to February 1977 (Figure 4-17a). This is in contrast to bores recharged from the
northwest that attained maximum increases in water levels from March onwards, suggesting that
different recharge mechanisms (with different timings) operate in the two areas. The cross-
correlation analysis presented in Section 4.4.3 also demonstrated that there is a rapid and a
delayed groundwater level response to rainfall events in the shallow aquifer.
It was established in Section 4.4.3 that the shallow aquifer is a watertable aquifer that is
recharged directly by rainfall. However, the analysis of watertable difference contours indicates
that lateral recharge from the northwest is also an important source of recharge, at least in some
parts of the aquifer. Therefore, there are at least two major recharge sources to HSs, with the
importance of each source varying spatially. Similar to the deep aquifer, there is also evidence
of a local recharge area in the north (Figure 4-17b). In addition, it has previously been suggested
by Herbert (1994) that the northeasterly trending flowline (near bore 48, Figure 4-16) is a
preferential flowpath for the inflow of seawater to the shallow aquifer. As discussed in Section
4.4.2, the upwards vertical head gradient towards the coast also indicates the potential for
leakage of deep groundwater to the shallow aquifer.
The groundwater flow pattern, in combination with the incision depth of the river bed into the
shallow aquifer (Section 4.2.2.1), indicates the potential for exchange of water between HSs and
the Herbert River. For instance, the shape of the contour lines, pointing upstream (concave)
where they cross the western half of the river, is indicative of a gaining stream due to the
discharge of shallow groundwater (Winter et al., 1998). In contrast, in the eastern half of the
river, watertable elevation contours point slightly downstream (convex) in the vicinity of the
stream, such as where the 4-10 m watertable contours cross the Herbert River in the wet season
(Figure 4-16b). This may indicate a losing river reach at that time of year and hence the
possibility of the river recharging the shallow aquifer. Further analyses are provided in Chapters
6 and 7 to examine river-aquifer interactions. Given the inter-aquifer connectivity relationships
established in Section 4.4.2 (Figure 4-7), vertical leakage from HSs to HSd is likely in the west.
Other plausible groundwater losses from the shallow aquifer system could result from
evaporation, discharge to sea, and pumping for town water supplies.
Chapter 4
84
#
#
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# #
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2.0 1.7
1.0 0.91.41.3
0.7
1.0
1.2
1.4
0.8
3.1
1.41.7
1.5
1.6
1.0
0.5
1.30.2
1.70.9
1.9
6.9
1.9
1.10.9
2.0
1.3
5.2
1.8
1.8
1.5 1.4
1.5
1.5
2.33.4
1.3
246
23
1
1
N
Stream
Head differencecontour
5 0 5 Km
February 1977-November 1976
# Water level difference
Outcrop
a
N10 0 10 20 Km
Catchment boundaryStreamHead difference contoursOutcropCardwell Range
Seaview R
ange
RECHARGEZONE
RECHARGEZONE
LOCALRECHARGE
DISCHARGEZONE
#
Mount LeachRange
DIFFUSERECHARGE
b
Figure 4-17 Head difference contours (m AHD) for the shallow aquifer system comparing groundwater levels during the 1977 wet season with previous dry season levels in November 1976. An enlarged area of the catchment is shown in (b), depicting recharge areas associated with the mountain ranges, diffuse rainfall sources, and the coastal discharge zone.
Hydrogeological Framework
85
4.5.3 Relationships between aquifers
Superposition of groundwater elevation contours for HSs and HSd during the same time periods
(Figure 4-18) illustrates that that there is a dominant preferential pathway for groundwater flow
from the northwest and parallel to the Herbert River in both aquifer systems. As noted above,
this flowpath follows the major palaeovalley delineated by the bedrock (Figure 4-1). The
concave shape of the groundwater elevation contours, where they cross the western half of the
river, suggests that the river acts as a groundwater sink for both aquifers. However, in the east,
where the main flowpath diverges from the river, a subtle difference between the aquifers is
observed: groundwater elevation contours for HSs are slightly convex in the vicinity of the river,
consistent with discharge from the river to the shallow aquifer, while those of HSd are more
concave. Further discussion of river-aquifer relationships is deferred to Chapters 6 and 7. In
general, the flowpaths in each aquifer follow a similar trajectory from recharge to discharge
areas, consistent with good vertical connectivity and hence a semi-confined alluvial aquifer
system. However, minor deviations between the HSd and HSs flowlines in the east (south of the
river) are probably due to the presence of localised thick mud sequences which restrict vertical
hydraulic connection, as shown by the area of poor vertical connectivity in Figure 4-7.
In addition to illustrating the direction of lateral groundwater movement in each aquifer, Figure
4-18 depicts the areas (shaded) where the direction of vertical flow is from the shallow aquifer
to the deep aquifer in representative wet and dry seasons. To derive these areas, groundwater
elevation contours between each aquifer were compared at the same period in time, to delineate
where the elevation of HSs was equal to HSd and hence where the elevation of HSs exceeded
HSd. This approach is similar to that used to compare bore hydrographs at nested sites to
determine the potential direction of flow between aquifers (Section 4.4). Note that not all areas
outside of the shaded area represent an upward head gradient between aquifer units, for
example, the northwestern extent of the shaded area is constrained by the available water level
data for the shallow aquifer (refer to Figure 4-16). Comparison of dry and wet season maps
clearly shows that the area of downwards hydraulic head expands to the east during the wet
season. Therefore, there is a seasonal switch in the vertical head gradient in bores towards the
coast and a resultant shift in the extent of potential discharge from HSd to HSs.
Chapter 4
86
24612 1018 81416
2
610
1620242832 30 26
N
Stream
HSd groundwaterelevation contour
Outcrop
5 0 5 Km
November 1976
#
#
# #
#
##
#
elevation HSs = HSd#
#
#
#
#
# # # #
#
#
#
elevation HSs > HSd
HSs groundwaterelevation contour
HSs flowlineHSd flowline
24612 1018 81416202224262610
24263034
N
Stream
HSd groundwaterelevation contour
Outcrop
HSs groundwaterelevation contour
5 0 5 Km
March 1977
# elevation HSs = HSdelevation HSs > HSd
#
#
#
#
#
#
#
##
#
#
## # # # #
##
# #
#
HSs flowlineHSd flowline
Figure 4-18 Contours of groundwater elevation superimposed for the shallow (HSs) and deep aquifers (HSd) during (a) the dry season and (b) the wet season. The approximate area where the groundwater elevation is greater in the shallow aquifer compared to the deep aquifer at the same period in time (i.e. the region of downwards hydraulic head) is shaded.
Hydrogeological Framework
87
4.6 CHAPTER SUMMARY
Lithological bore logs and hydraulic data were analysed in this chapter in order to conceptualise
the hydrogeology of the lower Herbert River catchment. Two alluvial aquifers are distinguished:
HSs, a shallow unconfined aquifer, and HSd, a semi-confined deep aquifer. Although the
aquifers are distinct, the extent of vertical hydraulic connection varies spatially. On balance, it is
considered that there is generally good vertical connectivity between HSs and HSd, with the
exception of an area in the middle of the catchment, where there is poor connection.
Groundwater in each aquifer follows approximately the same flowpaths: from the upland
recharge areas in the northwest and southwest towards (and parallel to) the Herbert River, and
then eastward to the coastal discharge area. A component of vertical discharge from the deep
aquifer to the shallow is also evident towards the coast, where the head gradient reverses from a
downwards to an upwards potential. In addition to lateral recharge at the catchment boundary, a
local recharge zone is evident in the middle of the catchment, to the north of the river. Direct
rainfall is also an important component of recharge to the shallow aquifer, with vertical leakage
to the deep aquifer likely in some areas. Importantly, the analyses in this chapter highlight the
relationship between the aquifers and the Herbert River. Whilst the direction of groundwater
flow in the deep aquifer is towards the river along much of its length, the depth of incision of
the river into HSd decreases down the valley. Based on the available information, potential
groundwater discharge from the deep aquifer to the river is restricted to the upper reaches of the
valley, approximately west of the junction with the Stone River. In the east, where incision of
the river bed is only shallow, the movement of deep groundwater is through the sediments of the
palaeochannel of the Herbert River. In contrast, there is the potential for direct exchange of
water between HSs and the Herbert River along its entire length. Available evidence suggests
that the river acts as a sink for shallow groundwater in the west and perhaps a source in the east.
Given the vertical hydraulic connection between HSd and HSs over a large part of the study area,
dissolved constituents that penetrate the groundwater system may be found in the two aquifers.
Therefore, in reaches where one or both of the aquifers interact with the Herbert River there is
the potential for groundwater to influence in-stream water quality and vice versa. In addition,
groundwater discharged to the river or directly to the sea has the potential to impact on water
quality in the marine environment. The following chapter examines hydrogeochemical data to
verify and develop the hydrogeological framework, including analysis of the distribution of
nitrogen in the subsurface. Subsequent chapters evaluate the physical and chemical relationships
between groundwater and surface water and assess the potential for a groundwater source of
nitrogen to the Herbert River.
89
Chapter 5 Hydrogeochemical Framework
5.1 INTRODUCTION
A conceptual framework for the movement of water in the alluvial aquifer system was
developed in Chapter 4 from existing lithologic and hydraulic information. Based on available
evidence, it was concluded that the subsurface can be conceptualised as a two-aquifer system
comprising a shallow unconfined aquifer (HSs) and semi-confined deep aquifer (HSd). The
nature of hydraulic connection (i.e. degree and direction) between the aquifers varies spatially.
Groundwater in each aquifer follows approximately the same flowpath, including flow towards
a section of the lower Herbert River enroute to the coastal discharge zone. Hydrogeochemical
data are examined in this chapter to verify and further develop these concepts. In accordance
with the interpretation based on hydraulics, hydrochemical analyses in the following sections
consider groundwater samples as representative of one of the two aquifer systems. Field data
collected for this study (Table 3-2) as well as existing sources are analysed (Table 3-1). Details
of the sampling program were outlined in Chapter 3 (Section 3.5).
There are numerous techniques used to analyse hydrochemical data, which take the form of
qualitative (graphical) and /or quantitative (numerical) procedures (Herczeg and Edmunds,
2000; Mazor, 1991; Hem, 1985; Freeze and Cherry, 1979; Zaporozec, 1972). The chosen
techniques are a function of the research questions and the nature of the available data (Section
2.4.1.4). This chapter specifically considers groundwater chemistry, while surface water
chemistry is introduced in Chapter 7 in the context of river-aquifer interactions. Given the
overarching interest in this thesis in the mobility of dissolved N, hydrochemical data relating to
species of N are examined in groundwater in this chapter and in surface waters in Chapter 7. A
conceptual diagram summarising water and N movement within the alluvial aquifer system is
presented at the end of this chapter (Figure 5-33). Whilst this chapter does not aim to provide a
comprehensive hydrogeochemical analysis, hydrochemical data is specifically analysed to:
• examine hydrochemical signatures within the lower catchment aquifers;
• establish the degree of vertical mixing between aquifers;
• examine lateral hydrochemical relationships within each aquifer; and
• explore the distribution and speciation of N within the hydrogeologic framework.
Chapter 5
90
Therefore, hydrochemistry is used as a tool to inform physical processes. Furthermore, by
characterising the chemistry of the groundwater, this provides a hydrochemical signature for the
groundwater component in surface waters, as discussed in Chapter 7.
5.1.1 General principles
5.1.1.1 Environmental tracers
Environmental tracers constitute natural or anthropogenic isotopes, elements or compounds that
are widely-distributed in the near-surface environment of the earth, such that variations in their
abundance can be used to infer environmental processes (Cook et al., 2002). Useful tracers
include (1) common dissolved constituents such as major cations and anions; (2) stable isotopes
of oxygen and hydrogen; (3) radioactive isotopes such as tritium; and (4) physical properties
such as water temperature (Winter et al., 1998). As most of these constituents are found in a
dissolved state in water, they have the potential to track the evolution of water as it moves
through the landscape. Tracers fall into the categories of conservative (inert) and non-
conservative (reactive) types, each type having their relative merits for particular applications.
Within each type there are: timescale, fingerprinting, reaction, pollution, and artificial tracers
(Cook, 2005). Central to this research is the application of pollution tracers, particularly
inorganic species of nitrogen, as a way of linking land-based agricultural activities to water
quality outcomes in both surface water and groundwater resources. Also of interest are
fingerprinting tracers which rely on the fact that where different water sources have different
chemical compositions, measurement of the chemistry of a water sample can provide
information on its origin.
A particular tracer can be conservative or non-conservative depending on the chemical
environment; for example, nitrate may be relatively inert under aerobic conditions and reactive
under reducing conditions. Thus, in contrast to Cl, which is generally considered to be
conservative, the concentration and speciation of nitrogen along a flowpath can vary
dramatically due to chemical interactions (Section 5.1.1.4). Therefore, it is important to
characterise the hydrochemical environment in order to inform the choice of environmental
tracer for the research problem, and to assist with interpretation of the results.
5.1.1.2 Ion chemistry
The solutes present in groundwater are primarily derived from rainfall and weathering,
including water-rock interaction. Water that infiltrates through the soil and unsaturated zones is
altered by evapotranspiration and the production of carbon dioxide from decay of organic matter
and plant respiration. The decay of organic matter also consumes dissolved oxygen and
produces water that is low or deficient in oxygen. The formation of carbon dioxide leads to the
production of H2CO3 which promotes mineral-water reactions. Other organic acids produced in
Hydrogeochemical Framework
91
the soil zone also aid mineral dissolution and hence affect the chemistry of water prior to its
passage to the saturated zone. Water that infiltrates into the saturated zone evolves
geochemically towards areas of discharge; chemical modification reflects time-dependent
processes such as water-rock/mineral interactions, dissolution-precipitation, and mixing.
Therefore, analysis of the hydrochemistry of groundwater along its flowpaths provides insight
into chemical processes that have influenced its lateral hydrochemical development.
In general, the bicarbonate ion is the dominant anion in recharge areas, derived from soil zone
CO2 and mineral weathering, especially dissolution of calcite and dolomite. Therefore, the
presence of bicarbonate-type waters is often indicative of proximity to a rainfall-recharge zone.
In addition, as groundwater moves along flowpaths there is generally an increase in total
dissolved solids (TDS) and the concentration of major ions. Hence, shallow groundwater is
typically lower in dissolved solids in recharge areas than in discharge areas (Freeze and Cherry,
1979). According to Chebotarev (1955), groundwater changes chemically towards the
composition of seawater, with anion evolution from bicarbonate (HCO3- ) to sulfate (SO4
2-) and
ultimately to chloride (Cl-) with increasing age and distance along the flowpath. Back (1960)
also related the chemical composition of groundwater to its flowpath; specifically, Ca-Mg-
HCO3 facies are predominantly found in recharge areas, whilst Na-Cl-SO4 facies are more
commonly found in discharge areas. Hydrochemical changes along the flowpath are attributed
to the processes of ion exchange and sulfate-reducing bacteria. An additional hydrochemical
evolution sequence is that of electrochemical evolution, concerning the tendency for the redox
potential of groundwater to decrease as water moves along its flowpath (Germanov et al., 1958).
Of particular relevance to this study is that in a closed system, the oxidation of organic matter
accompanied by consumption of O2 is followed by reduction of NO3-. Thus, in some
groundwater systems NO3- occurs at shallow depth and diminishes in concentration as the water
moves deeper into the flow system (Freeze and Cherry, 1979). It is important to note that the
hydrochemical evolution sequences described above are broad generalisations and each system
must be examined for peculiarities and local variability.
The notion of conservative and non-conservative tracers was discussed above (Section 5.1.1.1).
Chloride is generally assumed to behave conservatively in groundwater due to its high solubility
and inability to precipitate in minerals except at very high concentrations. This means that the
mobility of Cl is very similar to water molecules, noting that an ‘enriched’ Cl signature relative
to rainfall is carried to groundwater after leaving the zone of evapotranspiration. Relationships
between major ions (including non-conservative types) and chloride (as an inert reference ion)
are commonly used in hydrogeochemical investigations to differentiate compositional groups,
explore mixing relationships and calculate groundwater recharge (Cresswell and Herczeg, 2004;
Herczeg and Edmunds, 2000; Mazor, 1991; Hem, 1985; Zaporozec, 1972). Caution, however,
must be applied for groundwaters known to be influenced by seawater intrusion or in areas
Chapter 5
92
where Cl-bearing evaporite minerals are present, as in such situations Cl is no longer
conservative. Variations in the concentrations and ratios of non-conservative ions provide
information on geochemical reactions along a flowpath, and hence indicate maturity and relative
residence time of groundwater down-gradient.
5.1.1.3 Isotope chemistry
In addition to major and minor ion chemistry, isotopes (stable and radioactive) provide
information on hydrogeological characteristics of aquifers including origins, age and rate of
recharge, and aquifer interactions. Given the large number of isotopes and variety of
applications, the following discussion provides a brief introduction to stable isotopes of water
and selected radioactive isotopes. The application of 222Rn, a radioactive isotope useful for
examining river-groundwater interactions, is described in Chapter 7. Comprehensive reviews of
the principles and applications of isotopes in hydrogeology are provided in Clark and Fritz
(1997) and Gonfiantini et al. (1998).
Isotopes are atoms of the same element but with different masses, due to different numbers of
neutrons in the nucleus. Isotopes may be either stable (i.e. they do not change over geological
time under ambient conditions), such as 18O and 16O, or may be radioactive (i.e. are unstable and
decay by emitting a high-energy fragment, or fragments, until a stable isotope is created), such
as 14C. If the rate at which radio-isotopes decay is known, they can be used to measure process
rates and date events. Stable isotopes are generally measured as the ratio of the two most
abundant isotopes. The isotopes of the water molecule are particularly useful for hydrological
studies as they trace the actual movement of water (Cook and Herczeg, 1998). For a water
molecule, the oxygen isotopic ratio is represented by 18O/16O, while that for hydrogen is 2H/1H.
Measuring the ratio of heavy to light isotopes relative to a standard reference material is easier
and more accurate than determining absolute concentrations. Stable isotope ratios are generally
expressed using the delta (δ) notation:
δ (o/oo) ×−
=std
stdsample
RRR
1000
where R is the isotope ratio and δ is expressed as parts per thousand (o/oo). For δ18O and δ2H, R = 18O/16O or 2H/1H and the standard (std) is Standard Mean Ocean Water (SMOW). Thus,
seawater is assigned the δ values for oxygen and hydrogen equal to 0 o/oo (although local
variations are possible). Given this definition of δ, water samples that are relatively depleted in
the heavy stable isotopes are represented by negative values.
The different masses of isotopes result in different rates of mobility and reactions during
physical and chemical processes such as evaporation or water-rock exchange, a process known
Hydrogeochemical Framework
93
as fractionation. Fractionation processes affect the relative abundance of the isotopes of a given
element, which is central to their utility in hydrochemical investigations.
These processes and their implications for the stable isotope composition of rainfall can be
summarised as (Gat, 1980):
• the elevation effect: depletion of heavy isotopes in higher altitude rainfall;
• the amount effect: depletion of heavy isotopes with increasing quantity of rainfall;
• the continental effect: depletion of heavy isotopes with increasing distance from the coast; and
• the temperature effect: depletion of heavy isotopes at cooler temperatures.
In general, stable isotopes of oxygen and deuterium of most fresh waters around the world
conform to the relationship approximated by δ2H = 8δ18O + 10 (Craig, 1961), known as the
Global Meteoric Water Line (GMWL). While the GMWL is a useful reference for comparing
stable isotopic data, δ18O and δ2H for a given region may reflect different meteoric conditions
and be described by a local meteoric water line (LMWL) of a slightly different slope and
intercept (Ingraham, 1998). Hence, the LMWL may be the more relevant reference line for
detailed stable isotopic studies. Isotope fractionation occurs during evaporation of water such
that the residual water phase becomes relatively enriched in heavy isotopes (more positive δ18O
and δ2H). Thus, an evaporation line is an additional trend that may be observed in a collection of
water samples, depicted in a δ2H- δ18O plot as a series of points to the right of the local water
line with a slope determined by temperature and humidity (Mazor, 1991).
Radioactive isotopes play an important role for establishing the relative age of groundwater
(Gonfiantini et al., 1998). The naturally-occurring radio-isotope of tritium, 3H, is another
isotope of water that is useful for hydrochemical investigations. Large increases in the
concentration of 3H in rainfall occurred in the 1950’s and 1960’s due to thermonuclear weapons
testing. Prior to testing, atmospheric levels were 5-10 tritium units (TU), where one TU
represents one atom of 3H in 1018 atoms of hydrogen. In the southern hemisphere mean annual
tritium concentrations exceeded 50 TU in the early 1960’s before declining to natural levels
(Cook and Herczeg, 1998). The half-life of tritium is 12.43 years and thus can only be used to
study systems in which the residence or transit time of groundwater is a few to 100 years
(Gonfiantini et al., 1998). The most useful environmental radio-isotopes for determining
groundwater age in confined aquifers are those with a half-life ranging from 103 to 106 years.
For example, carbon-14 has a half-life of 5730 years, while chlorine-36 has a much longer half-
life of 301 000 ± 2000 years. Further discussion of the theory and application of these isotopes
can be found in Cook and Herczeg (1998) and Gonfiantini et al. (1998).
Chapter 5
94
5.1.1.4 Nitrogen chemistry
Nitrogen can exist in many forms. In order of decreasing oxidation state, the dissolved species
of interest to the soil/water environment include: nitrate (NO3-), nitrite (NO2
-), nitrogen gas (N2),
ammonium (NH4+), and ammonia (NH3). Dissolved organic N (DON) may also be present in
freshwaters. Transformation of N compounds can occur through several mechanisms: fixation,
ammonification, synthesis, nitrification, and denitrification (Canter, 1997). Transformations that
occur between the different forms of N in the biosphere are summarised by the nitrogen cycle
(Figure 5-1).
Leaching to groundwater
(N2)
Nitrification
Mineralisation
Nitrification
LightningFixation
BiologicalFixation
IndustrialFixation
Denitrification
PlantConsumption
Adsorption
Leaching to surface water
INTERACTION
Leaching to groundwater
(N2)
Nitrification
Mineralisation
Nitrification
LightningFixation
BiologicalFixation
IndustrialFixation
Denitrification
PlantConsumption
AdsorptionDenitrification
PlantConsumption
Adsorption
Leaching to surface water
INTERACTION
Figure 5-1 The nitrogen cycle, modified after Pidwirny (2005). The interaction between groundwater and surface water and the potential transport of N between them is also indicated.
Given an inorganic source of N, such as fertiliser, the key transformation reactions of direct
relevance to this study include:
• Nitrification: biological oxidation of ammonium ions to nitrite and then nitrate,
involving specific chemoautotrophic bacteria and an inorganic carbon source
i.e. NH4+ + O2 → NO2
- + O2 → NO3- (5-1)
• Denitrification: biological reduction of nitrate to nitrogen gas, involving heterotrophic
bacteria and an organic carbon source
i.e. NO3- + organic carbon → NO2
- + organic carbon → N2(g) + CO2 + H2O (5-2)
Hydrogeochemical Framework
95
Whilst nitrification leads to the formation of nitrate that is readily leached to soils and
groundwater, denitrification produces nitrogen gases that can escape from the aqueous
environment. Changes in environmental conditions that can affect N transformations include
redox potential, temperature, pH, microbial population, and the concentrations of electron
donors (e.g. dissolved organic carbon) and electron acceptors (e.g. oxygen, nitrate)
(Thayalakumaran et al., 2004; Canter, 1997).
5.1.2 Methods of interpretation
Groundwater investigations can produce large volumes of chemical data. Therefore, effective
methods for organising, classifying and presenting the data are important for assisting with
analysis and hydrochemical interpretation. Methods utilised specifically in this chapter are
summarised below. In this chapter, ionic concentrations are generally expressed in the units of
milliequivalents per litre (meq/L), while concentrations of other chemical species are expressed
in milligrams per litre (mg/L).
Trilinear plots, such as the Piper diagram (Piper, 1944), provide a useful tool for summarising a
large number of chemical analyses in a single plot. The Piper diagram (e.g. Figure 5-2)
combines two trilinear plots, each representing the relative proportions (in meq/L) of the major
cations and anions. In addition to comparing relationships between cations and anions, Piper
diagrams can also highlight distinct chemical groups, trends along flowpaths and mixing
between waters. Fingerprint diagrams, such as the Schoeller plot (Schoeller, 1955), also have
the ability to represent numerous major ion analyses on one diagram (e.g. Figure 5-5); unlike
trilinear diagrams they have the advantage of representing actual ion concentrations rather than
normalised values. This feature is particularly useful when exploring chemical similarities and
differences between aquifers and hence ascertaining the degree of inter-aquifer groundwater
interactions (Section 5.3).
Bivariate plots allow interpretation of the relationships between two variables. In particular,
expressing the concentrations of ions as ratios of one ion to another, or of one constituent to the
total concentration, can be helpful for highlighting similarities and differences between waters
(Hem, 1985). Ion/chloride bivariate plots (e.g. Figure 5-10) are commonly used to show the
relationships between major or minor ions relative to the conservative Cl ion. Compositional
clusters and mixing relationships can also be ascertained from these types of graphical analyses.
Stable isotope plots comparing δ18O and δ2H (e.g. Figure 5-11) are another example of a
bivariate plot.
The graphical techniques discussed above lack the ability to easily represent the geographical
distribution of the hydrochemistry. For this reason, a combination of methods is desirable in any
hydrogeochemical investigation. To explore the spatial hydrochemical evolution of groundwater
Chapter 5
96
along flowpaths, water quality maps (e.g. Figure 5-23) can be constructed to illustrate the
distribution of particular ions or their ratios, TDS, water types, isotope compositions and other
measured/calculated parameters at each sampling location. The determination of recharge and
discharge areas is an important aspect of hydrogeological studies and therefore hydrochemical
maps are a useful tool for conceptualising flow within the groundwater system. Inclusion of
other geographical attributes such as surface topography, flowpaths and surface drainage can
also aid with spatial interpretation.
In addition to the above qualitative methods for analysing hydrochemical data, semi-
quantitative/quantitative techniques include geochemical modelling, such as the calculation of
saturation indices i.e. the degree of saturation with respect to specified minerals in a
groundwater sample (Section 5.3.2). This is useful for comparing interactions between aquifers
and tracing groundwater evolution along flowpaths. Mass balance calculations also allow
mixing proportions between waters to be estimated, where the end-member compositions are
well constrained and mixing is conservative.
Having outlined the key hydrochemical methods that will be utilised in this chapter, it is
important to note that in general, different graphical/numerical techniques will be relevant to
different hydrogeological settings. Therefore, hydrochemical analyses have greatest value when
analysed and interpreted within a sound and reliable hydrogeological context (Herczeg and
Edmunds, 2000). Moreover, interpretation of hydrochemical data for tracing water flowpaths is
best done by collecting data along hydraulic transects, or where hydraulic connection can be
assured (Woessner, 2000).
5.2 HYDROCHEMICAL PATTERNS
Groundwaters of the lower Herbert River catchment cover a broad compositional range in
relation to major ion chemistry. As illustrated in Figure 5-2, groundwaters are dominated by
sodium (Na+), calcium (Ca2+), bicarbonate (HCO3-) and chloride (Cl-) ions and are relatively
depleted in magnesium (Mg2+) and sulfate (SO42-). Potassium (K+) is not included on the Piper
diagram due to its low concentration compared to other major ions. Note that for convenience,
ions are herein referred to without their charge, with the exception of species of nitrogen. A
summary of major and minor inorganic chemistry for all groundwater samples is provided in
Appendix A.
Hydrogeochemical Framework
97
80 60 40 20 20 40 60 80
20
40
60
80 80
60
40
20
20
40
60
80
20
40
60
80
Ca Na HCO3 Cl
Mg SO4
RainfallHSsHSdSeawater
Figure 5-2 Piper diagram for deep (HSd) and shallow (HSs) aquifer samples collected during three sampling periods: May 2004, October 2004, June 2005. Note that only samples along the Herbert River valley and lower Stone River are shown.
Analysis of stable isotopes of water indicates that deep and shallow groundwaters primarily lie
on the LMWL (Figure 5-3). Differences in isotopic values along the LMWL result from a
combination of the elevation, amount and continental effects (Section 5.1.1.3). The amount
effect and differences in the sources of recharge can account for the observed isotope
differences between the two aquifers. For example, in general, deep aquifer waters are more
depleted (more negative) in heavy isotopes, consistent with large recharge events. In contrast,
shallow aquifer waters are more enriched in heavy isotopes, representative of smaller rainfall
events and diffuse recharge. The most depleted groundwaters depicted in Figure 5-3 (enclosed)
are from the upper Stone River valley. Note that the uncertainty in δ2H and δ18O is 1 and 0.15 o/oo, respectively; thus, the departure of some samples from the LMWL, outside of this
uncertainty, is indicative of other processes influencing the groundwater isotopic composition.
Further discussion of isotope data is deferred to Section 5.2.2. Note that although the rainfall
sample for major ion analysis was collected inland, stable isotopes were only determined for a
coastal rainfall sample.
Chapter 5
98
-7 -5 -3 -1 1-40
-30
-20
-10
0
10
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
SMOW
coastal rainfall MoretonBay
GMWL
Local meteoricwater line
Brisba
ne W
L
RainfallHSsHSdSeawater
-7 -5 -3 -1 1-40
-30
-20
-10
0
10
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
SMOW
coastal rainfall MoretonBay
GMWL
Local meteoricwater line
Brisba
ne W
L
RainfallHSsHSdSeawater
-7 -5 -3 -1 1-40
-30
-20
-10
0
10
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
SMOW
coastal rainfall MoretonBay
GMWL
Local meteoricwater line
Brisba
ne W
L
RainfallHSsHSdSeawater
Figure 5-3 Oxygen-18 (δ18O) and deuterium (δ2H) stable isotope data for deep and shallow aquifer samples and a coastal rainfall event in May 2004. The isotopic composition of seawater at Moreton Bay (QLD) (Cresswell 2006, pers. comm.) and SMOW are also shown. Trend lines represent the LMWL (blue); GMWL (black solid); and Brisbane water line6 (black dashed). Groundwaters from the upper Stone River valley are enclosed (red).
5.2.1 Compositional groups
Hydrochemical groups within each aquifer can be distinguished based on similarities in the
relative concentrations of major ions and salinities. As outlined in Chapter 3 (Section 3.5),
groundwater samples were collected during three sampling periods: May 2004, October 2004
and June 2005. For the majority of bores sampled in at least two of these periods, there is little
variation in groundwater composition, especially in the deep aquifer (Figure 5-4). Therefore, as
samples from June 2005 have the best spatial coverage, they are used as a basis for identifying
broad hydrochemical groups within each aquifer.
6 Data for the Brisbane water line was obtained from the Isotope Hydrology Section database maintained by the International Atomic Energy Agency (IAEA, 2006). The Brisbane WL was calculated from monthly amount-weighted mean δ18O and δ2H values.
Hydrogeochemical Framework
99
80 60 40 20 20 40 60 80
20
40
60
80 80
60
40
20
20
40
60
80
20
40
60
80
Ca Na HCO3 Cl
Mg SO4
HSs samples
May 2004October 2004June 2005
a
80 60 40 20 20 40 60 80
20
40
60
80 80
60
40
20
20
40
60
80
20
40
60
80
Ca Na HCO3 Cl
Mg SO4
HSd samples
May 2004October 2004June 2005
b
Figure 5-4 Piper diagram for (a) HSs and (b) HSd groundwater samples displayed by month of collection. Note that only bores sampled in at least two of the three sampling periods are shown.
Schoeller plots (Figure 5-5) illustrate that HSs waters can be grouped according to the relative
concentrations of HCO3 and Cl. Within the HCO3-dominated type there are two distinct patterns
based primarily on the relative proportions of the three major cations: one group is enriched in
Na (Figure 5-5a), while the other group has a greater proportion of Ca and Mg (Figure 5-5b).
While Cl-dominated waters overlap in their relative proportions of cations, they are
distinguished by their HCO3/Cl ratio and relative concentrations of Na to Cl. For instance, one
group has a greater proportion of HCO3 to Cl and higher concentration of Na compared to Cl
(Figure 5-5c). Conversely, the other group has a hydrochemical pattern that resembles seawater
Chapter 5
100
and thus has a lower HCO3/Cl ratio and virtually equal concentrations of Na and Cl (Figure
5-5d). The hydrochemical relationships between the identified groups are illustrated in Figure
5-6.
a
b
c
d
Figure 5-5 Schoeller plots illustrating the two dominant hydrochemical groups observed in shallow aquifer samples in June 2005: (a) and (b) represent HCO3-enriched samples while (c) and (d) represent Cl-enriched samples, including seawater (red) (refer to Figure 5-6).
Hydrogeochemical Framework
101
Figure 5-6 Piper diagram for shallow aquifer samples in June 2005: shaded areas highlight waters with similar hydrochemical fingerprints. The blue and green shaded areas represent HCO3-dominated waters whilst the purple and red areas represent Cl-dominated waters (refer to Figure 5-5).
Similar to HSs, two dominant compositional groups are distinguished in HSd based on the
relative concentration of anions (Figure 5-7). HCO3-dominated waters are the most dilute of the
HSd waters: within the group there are samples that are relatively enriched in Na compared to
the other major cations (Figure 5-7a) and other samples enriched in Ca and Mg (Figure 5-7b).
The second compositional group is more enriched in Cl relative to HCO3, with both high and
low salinity trends. There is considerable variation in relative cation and anion concentrations
between samples in the low salinity subgroup; however, these waters define a distinct
hydrochemical cluster (Figure 5-8). Similarly, whilst the concentration of cations is variable
within the high salinity subgroup, particularly with respect to Ca (Figure 5-7d), the waters are
nonetheless grouped as one type due to the overlap in relative anion concentrations (Figure 5-8).
Note that while the Schoeller plots for some bores in both Cl enriched subgroups are similar,
high salinity waters generally have a greater concentration of Mg, Na and Cl; are enriched in Cl
relative to HCO3; and have Na < Cl (compare Figure 5-7 c and d).
Comparison of Schoeller plots indicates that there are similar hydrochemical patterns between
the aquifers. The Na-enriched and Ca-Mg enriched trends of HSs are similar to those in HSd,
while the Na = Cl trend of HSs resembles the high salinity trend of HSd. There is also overlap
between the low salinity and Na > Cl trends in the deep and shallow aquifers, respectively.
Although there are chemical variations within the hydrochemical groups identified, particularly
in regards to relative cation concentrations, these discrepancies are small compared to the major
Chapter 5
102
shifts in anion concentrations between the groups. The spatial distribution of ions in HSs and
HSd is examined in Section 5.2.3.
a
b
c
d
Figure 5-7 Schoeller plots illustrating the two dominant hydrochemical groups observed in deep aquifer samples in June 2005: (a) and (b) represent HCO3-dominated samples enriched in Na and Ca-Mg, respectively, while (b) and (c) represent the low and high salinity Cl-enriched samples, respectively, including seawater (red).
Hydrogeochemical Framework
103
Figure 5-8 Piper diagram for deep aquifer samples in June 2005: shaded areas highlight waters with similar hydrochemical fingerprints. The blue and green shaded areas represent HCO3-dominated waters whilst the purple and red shaded areas represent Cl-enriched waters (refer to Figure 5-7).
5.2.2 Linear trends
The previous analyses identified the main compositional groups in each aquifer based on major
ion concentrations in each sample. Relationships between solutes are examined in this section,
with data presented in terms of the main hydrochemical groups identified above. Plots of TDS
against Cl for each aquifer display a good linear relationship at high salinity (Figure 5-9);
however, at low salinity Cl is not a good indicator of total solutes in solution. Nevertheless, Cl
is used as a conservative reference ion in the following analyses in order to examine changes in
solute concentrations with increasing salinity and to identify hydrochemical clusters. An
increase in the concentration of solutes with Cl is indicative of groundwater evolution due to
processes such as water-rock/mineral interactions, evaporation and ion exchange. However, in
order to infer groundwater evolution with confidence, spatial trends along groundwater
flowpaths must also be examined (Section 5.4).
Chapter 5
104
1 10 100 1000 10000 10000010
100
1000
10000
100000TDS
Cl (mg/l)
HSd samples
SeawaterHigh salinityLow salinityCa-Mg enrichedNa enriched
1 10 100 1000 10000 1000001
10
100
1000
10000
100000TDS
Cl (mg/l)
HSs samples
Na = ClNa > ClCa-Mg enrichedNa enrichedSeawater
Figure 5-9 TDS vs Cl (mg/L) for all deep and shallow aquifer samples (refer to Figures 5-5, 5-6, 5-7 and 5-8 for classification into hydrochemical groups).
5.2.2.1 Deep aquifer
Bivariate plots of major ions against Cl show that groundwater compositions in HSd lie between
rainfall and seawater (Figure 5-10). Na enrichment relative to rainfall is indicative of cation
exchange reactions and interactions with Na-containing minerals. Whilst there is generally a
positive correlation of major ions with Cl, there are distinct deviations from linearity which
correspond to particular hydrochemical groups. Notably, at low salinity there are two clusters of
HCO3-dominated samples that are distinguished in the Ca vs Cl and Mg vs Cl plots: these
clusters correspond to the Ca-Mg enriched and Na enriched hydrochemical groups identified
(Section 5.2.1). Spatial comparison of these water types with the distribution of major soil types
in the area (up to 1 m below ground surface) highlights that Na-enriched waters are associated
with silty clays/loams that have a medium to high cation exchange capacity (CEC) (between 3.9
- 7.9 meq/100g) and have acid volcanic parent material (Wood et al., 2003). In contrast, Ca-Mg
enriched waters are associated with well drained soils (red loam and river bank soils) with a low
CEC of around 3.1 meq/100g (Wood et al., 2003). Calcite (CaCO3) and dolomite (CaMg(CO3)2)
dissolution in the presence of CO2-charged water infiltrating through the soil zone can account
for the Ca-Mg-HCO3 groundwater; coupled with the presence of clay minerals with
exchangeable Na+, Na-HCO3-dominated waters can also be produced through ion exchange.
Hydrogeochemical Framework
105
0 0 1 10 100 10000.01
0.1
1.
10.
100.
1000.Na (meq/l)
Cl (meq/l)0.01 0.1 1. 10. 100. 1000.
0.001
0.01
0.1
1.
10.
100.K (meq/l)
Cl (meq/l)
0.01 0.1 1. 10. 100. 1000.0.001
0.01
0.1
1.
10.
100.Ca (meq/l)
Cl (meq/l)0.01 0.1 1. 10. 100. 1000.
0.01
0.1
1.
10.
100.
1000.Mg (meq/l)
Cl (meq/l)
0.01 0.1 1. 10. 100. 1000.0.1
1.
10.HCO3 (meq/l)
Cl (meq/l)0.01 0.1 1. 10. 100. 1000.
0.001
0.01
0.1
1.
10.
100.SO4 (meq/l)
Cl (meq/l)
HSd samples
RainfallSeawaterHigh salinityLow salinityCa-Mg enrichedNa enriched
Figure 5-10 Bivariate plots of major ions against Cl for deep (HSd) groundwater samples collected during the three sampling periods in 2004-2005. Note that the rainfall sample was collected approximately 20 km inland from the coast.
Chapter 5
106
Geochemical processes that can produce the two HCO3 waters are represented by the reactions:
CaCO3 + H2CO3 → 2HCO3- + Ca2+ (5-3)
Ca2+ + 2Na(ad) ↔ 2Na+ + Ca(ad) (5-4)
where (ad) denotes cations adsorbed on clays (Freeze and Cherry, 1979).
In addition to the two clusters of waters at low salinity, there is a cluster of high salinity samples
in HSd with a different HCO3:Cl relationship to the remaining samples (e.g. HCO3 vs Cl, Figure
5-10). Some of these high salinity groundwaters have anomalous concentrations of Ca relative
to the linear trend, above that of seawater. Stable isotope compositions (for a subset of bores)
illustrate that the high salinity waters plot to the right of the LMWL (Figure 5-11). Although
evaporation prior to recharge can account for these compositions, the linear trend depicted in
Figure 5-11 is consistent with mixing between freshwater and seawater. Further evidence for
this is provided in Section 5.2.3.1 with reference to the spatial distribution of the main water
types. Note that bore 101A lies on the LMWL; however, allowing for the uncertainty in stable
isotope measurements (Section 5.2) it is plausible that the stable isotopic composition of this
bore also corresponds to the seawater mixing line. Stable isotopes for all bores except those in
the high salinity group indicate a purely meteoric origin, with no evidence of evaporation prior
to recharge. This is consistent with lateral recharge to the deep aquifer as opposed to slow
infiltration due to diffuse recharge (e.g. Figure 4-15b, Chapter 4).
The group of low salinity waters show considerable scattering with respect to the major ion-Cl
trends. However, based on visual inspection of the series of bivariate plots presented in this
section, a degree of hydrochemical evolution from HCO3-dominated waters (Ca-Mg and/or Na
enriched types) to low salinity Cl-dominated waters is possible. Lateral hydrochemical trends
along groundwater flowlines in the deep aquifer are specifically examined in Section 5.4 in
order to confirm this observation.
Hydrogeochemical Framework
107
-7 -5 -3 -1 1-40
-30
-20
-10
0
10 HSd samples
RainfallSeawaterHigh salinityLow salinityCa-Mg enrichedNa enriched
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
Seawater mixing line
SMOW
coastal rainfall
Local meteoricwater line
MoretonBay
49A
101B
48A36A
101A
-7 -5 -3 -1 1-40
-30
-20
-10
0
10 HSd samples
RainfallSeawaterHigh salinityLow salinityCa-Mg enrichedNa enriched
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
Seawater mixing line
SMOW
coastal rainfall
Local meteoricwater line
MoretonBay
49A
101B
48A36A
101A
Figure 5-11 Oxygen-18 (δ18O) and deuterium (δ2H) stable isotope data for HSd samples and a coastal rainfall event in May 2004. The isotopic composition of seawater at Moreton Bay (QLD) (Cresswell 2006, pers. comm.) and SMOW are also shown. Trend lines represent the LMWL (blue) and mixing with seawater (red). Bores common to Figure 5-16 are labelled.
5.2.2.2 Shallow aquifer
Bivariate plots of major ions against Cl in HSs are displayed in Figure 5-12. While there is a
linear relationship observed between Na and Cl, there are no dominant trends with respect to the
other ions and Cl. However, the ion-Cl plots in combination with other bivariate plots highlight
distinct groups of waters in HSs that can be attributed to various natural processes. Similar to
groundwaters in HSd, a population of dilute waters in HSs has elevated Ca and Mg
concentrations (e.g. Ca vs Cl and Mg vs Cl, Figure 5-12); this cluster corresponds to the Ca-Mg
enriched hydrochemical group. As established for the deep aquifer, Ca-Mg enriched waters are
associated with sandy soils that have a low CEC of ~3 meq/100g. In contrast, the Na enriched
waters are associated with silty clays/loam and thus have a higher clay content, with CEC ~4
meq/100g.
Chapter 5
108
0.01 0.1 1. 10. 100. 1000.0.01
0.1
1.
10.
100.
1000.Na (meq/l)
Cl (meq/l)0.01 0.1 1. 10. 100. 1000.
0.001
0.01
0.1
1.
10.
100.K (meq/l)
Cl (meq/l)
0.01 0.1 1. 10. 100. 1000.0.001
0.01
0.1
1.
10.
100.Ca (meq/l)
Cl (meq/l)0.01 0.1 1. 10. 100. 1000.
0.01
0.1
1.
10.
100.
1000.Mg (meq/l)
Cl (meq/l)
0.01 0.1 1. 10. 100. 1000.0.1
1.
10.HCO3 (meq/l)
Cl (meq/l)0.01 0.1 1. 10. 100. 1000.
0.01
0.1
1.
10.
100.SO4 (meq/l)
Cl (meq/l)
HSs samples
Na = ClNa > ClCa-Mg enrichedNa enrichedRainfallSeawater
Figure 5-12 Bivariate plots of major ions against Cl for shallow (HSs) groundwater samples collected during the three sampling periods in 2004-2005.
Hydrogeochemical Framework
109
As noted for the deep aquifer, Na = Cl waters in HSs display a different HCO3-Cl relationship to
other samples, with lower HCO3 for the concentration of Cl (e.g. HCO3 vs Cl, Figure 5-12).
Schoeller plots in Figure 5-5d illustrated that these waters resemble dilute seawater. Br-Cl
relationships (Figure 5-13) are also consistent with a dominant seawater origin. Note that while
these waters have a seawater signature, this does not necessarily imply mixing with seawater
per se. Coastal rainfall would also be expected to resemble a seawater source; for example, as a
starting point, the Br/Cl ratio of rainfall can be assumed to be similar to that of seawater
(Herczeg and Edmunds, 2000). Stable isotopes indicate that unlike the deep aquifer, there is no
evidence of mixing with seawater in the shallow aquifer (Figure 5-14). Therefore, samples lying
below the LWML may indicate secondary fractionation processes such as evaporation prior to
infiltration (e.g. bore 61B). The Na > Cl hydrochemical group is only represented by a few
samples. Available evidence from major ions suggests that these waters display similar ion-Cl
relationships to the Na = Cl group, although Na > Cl waters are more enriched in HCO3 relative
to Cl.
0.1 1. 10. 100. 1000.0.0001
0.001
0.01Br/Cl
Cl (meq/l)
HSs samples
Na = ClNa > ClCa-Mg enrichedNa enrichedSeawater
Figure 5-13 Bivariate plot of Br/Cl against Cl for shallow aquifer samples collected during the three sampling periods in 2004-2005.
Based on the above analyses, a summary of the main hydrochemical groups in the deep and
shallow aquifers and possible explanations for these waters types is provided in Table 5-1.
Chapter 5
110
-7 -5 -3 -1 1-40
-30
-20
-10
0
10 HSs samples
Na = ClNa > ClCa-Mg enrichedNa enrichedRainfallSeawater
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
Seawater mixing line
SMOW
coastal rainfall
Local meteoricwater line
MoretonBay
61B
-7 -5 -3 -1 1-40
-30
-20
-10
0
10 HSs samples
Na = ClNa > ClCa-Mg enrichedNa enrichedRainfallSeawater
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
Seawater mixing line
SMOW
coastal rainfall
Local meteoricwater line
MoretonBay
-7 -5 -3 -1 1-40
-30
-20
-10
0
10 HSs samples
Na = ClNa > ClCa-Mg enrichedNa enrichedRainfallSeawater
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
Seawater mixing line
SMOW
coastal rainfall
Local meteoricwater line
MoretonBay
61B
Figure 5-14 Oxygen-18 (δ18O) and deuterium (δ2H) stable isotope data for HSs samples and a coastal rainfall event in May 2004. The isotopic composition of seawater at Moreton Bay (QLD) (Cresswell 2006, pers. comm.) and SMOW are also shown. Trend lines represent the LMWL (blue) and mixing with seawater (red) (derived from deep aquifer samples). Bore 61B is labelled as representative of evaporation prior to recharge.
Table 5–1 Hydrochemical groups of the alluvial aquifers Observation Possible explanation
HCO3-dominated compositions (HSd and HSs)
Ca-Mg enriched
Na enriched
calcite/dolomite dissolution
associated with sandy soils (low CEC)
associated with clay soils (high CEC)
HSd: Cl-dominated compositions
low salinity
high salinity
hydrochemical evolution
mixing with seawater
HSs: Cl-dominated compositions
Na = Cl
Na > Cl
seawater origin
insufficient evidence, but similar to Na = Cl type
Hydrogeochemical Framework
111
5.2.3 Spatial trends
Hydrochemical relationships between the deep and shallow aquifers are examined in Section
5.3, while the lateral chemical evolution of groundwater through HSd and HSs is examined in
Section 5.4. The aim of this section is to place the hydrochemical patterns observed from the
Piper, Schoeller and bivariate plots into a spatial context. Nitrate (NO3-) is also included in the
anion plots to observe relationships with the main water types.
5.2.3.1 Deep aquifer
Spatial representation of the identified hydrochemical groups indicates that HSd waters that
display a high salinity, Cl-enriched trend (Figure 5-7d) are located within 10 km of the coast.
Anion and cation pie charts in Figure 5-15 clearly illustrate that these waters (red shaded area)
characteristically have a high TDS and relatively minor concentrations of anions other than Cl.
Consistent with stable isotope trends (Figure 5-11), a plot of monovalent ions against divalent
ions normalised to chloride provides further evidence of interactions with seawater (Figure
5-16). For example, when seawater intrudes into a coastal freshwater aquifer (trend 1, Figure
5-16), Ca-Na exchange on clays results in an increase in divalent ions in the groundwater
(increase in (Ca + Mg)/Cl). This is balanced by a relative depletion in monovalent ions relative
to seawater (decrease in Na/Cl) such that Δ(Ca + Mg) = -2ΔNa molar units, as depicted by the
theoretical mixing lines in Figure 5-16 (Vengosh, 2004). In contrast, seawater displaced by
freshwater results in an opposite reaction such that there is an increase in Na/Cl and decrease in
(Ca + Mg)/Cl (Appelo, 1994). Bores along trend 2 in Figure 5-16, characterised as low salinity
waters, may be influenced by this process.
Groundwaters in the middle of the catchment, down-flow of the main recharge area for HSd,
belong to the low salinity group (Figure 4-15). Therefore, as suggested from the linear trends
(Figure 5-10), it is possible that these waters represent evolved groundwater compositions.
However, it is noted that low salinity groundwaters (bores 36 and 38) are also observed at the
coast, which define trend 2 in Figure 5-16. Therefore, there may be different processes that give
rise to this hydrochemically diverse group of low salinity waters. It was previously noted that
HCO3-dominated Ca-Mg and Na enriched waters correlate with the distribution of soil type in
the lower catchment (Table 5-1). Ca-Mg enriched waters in the upper part of the catchment, to
the northwest, are characteristically elevated in NO3-. This has significance for the mobility of N
in groundwater, as discussed in Section 5.6.
Chapter 5
112
a
b
Figure 5-15 Pie charts illustrating the spatial distribution of (a) major anions and (b) major cations in HSd in June 2005 as a percentage of total meq/L. The size of the inner circle is proportional to TDS (mg/L). Screened depths are shown in metres below ground level at each bore (italics). Refer to Figure 5-7, Figure 5-8 and Figure 5-10 for corresponding Schoeller, Piper and bivariate plots. Note that the pie chart at bore 101 is displayed for the deeper screened interval (101A or 51A); bore 101B (not shown but in the same water group) is screened at 40-46 m below ground.
Hydrogeochemical Framework
113
0
1
2
3
4
5
6
7
8
0 1 2 3
(Ca+Mg)/Cl (molar)
Na/
Cl (m
olar
)
HSd boresSeawater
49A
101A
36A
48A
38A
53A
1
2
45A46A
47A44A
101B
0
1
2
3
4
5
6
7
8
0 1 2 3
(Ca+Mg)/Cl (molar)
Na/
Cl (m
olar
)
HSd boresSeawater
49A
101A
36A
48A
38A
53A
1
2
45A46A
47A44A
101B
Figure 5-16 Monovalent ions plotted against divalent ions, normalised to Cl for all HSd bores (purple) and seawater (blue). Inset: samples are scaled by TDS (mg/L) and depict two trends away from seawater composition (compare with bores labelled in Figure 5-11). Dashed red lines depict idealised mixing lines for seawater intruding into sediments.
5.2.3.2 Shallow aquifer
Spatial relationships between major anions and cations in HSs are depicted in Figure 5-17, with
the main compositional groups highlighted. Based on the samples collected, the spatial
distribution of hydrochemical groups shows that the Na = Cl waters are restricted to the
northeast. Interestingly, two of the three bores that define this water type have low TDS (150-
300 mg/L), while the third (bore 36B) displays elevated TDS (around 1000 mg/L). High TDS
waters comprising the Na > Cl hydrochemical group do not show an obvious spatial trend.
These observations are further examined in Section 5.4.2.
As illustrated in Figure 5-17, there is one bore that is screened at two depths within HSs: whilst
the shallowest bore (54B) is a Ca-Mg enriched type (pie chart displayed), the deeper interval
(54A) is a Na-enriched type. It was noted above that the spatial distribution of the Ca-Mg and
Na enriched waters can be attributed to soil type (Table 5-1). Furthermore, with reference to the
lithologic cross-sections in Chapter 4 (e.g. Figure 4-4), the two types of HCO3-dominated
waters relate to different sandy units within the shallow aquifer. Hydrochemical relationships
between these units within HSs and justification for their inclusion in one aquifer are discussed
in Section 5.3.3. Importantly, the relative concentration of NO3- is highest in HCO3-dominated
waters that are enriched in Ca and Mg. Factors that influence the spatial distribution of NO3- are
examined in Section 5.6.
Chapter 5
114
a
b
Figure 5-17 Pie charts illustrating the spatial distribution of (a) major anions and (b) major cations in HSs in June 2005 as a percentage of total meq/L. The size of the inner circle is proportional to TDS (mg/L). Screened intervals are shown in metres below ground level at each bore (italics). Refer to Figure 5-5, Figure 5-6 and Figure 5-12 for corresponding Schoeller, Piper and bivariate plots. Note that the pie chart at bore 54 is displayed for the shallowest screened interval (54B), which represents a Ca-Mg enriched composition.
Hydrogeochemical Framework
115
5.3 VERTICAL RELATIONSHIPS BETWEEN AQUIFERS
This section examines hydrochemical relationships between HSs and HSd in order to establish
the degree of vertical interaction between the aquifers and hence verify the conceptual model
developed in Chapter 4. The previous section showed that there is considerable overlap in the
chemistry of shallow and deep groundwaters with respect to major and minor ions, thus it is
difficult to define a unique chemical signature for each aquifer. However, similar to the bore
hydrograph analysis presented in Chapter 4 (Section 4.4), hydrochemical comparisons between
nested bores enable site-specific vertical relationships between the aquifers to be examined.
5.3.1 Unconfined and confined systems
The chemistry of subsurface waters is a complex function of many variables, including: the
composition of groundwater recharge; the mineralogical composition of subsurface rocks; and
the hydrogeologic properties of rocks, that influence the extent of water/rock interaction
(Langmuir, 1997). These factors result in chemical differences between shallow unconfined and
deeper semiconfined/confined aquifers. As discussed in Section 5.1.1, there are numerous key
processes that influence the chemistry of groundwater, namely: evapotranspiration, dissolution
and precipitation, ion exchange, sorption, redox reactions, and gas generation and consumption
(also refer to Figure 2-3). Whilst many of these processes are common to the unsaturated and
saturated zones, the relative importance of each process in each zone may vary. Evaporation, for
instance, is likely to influence the chemistry of a watertable aquifer; however, it is not a
dominant process in deep confined aquifer systems. In addition, the degree and extent of
weathering affects the residence time of water, which is generally shorter in shallow aquifer
systems compared to deeper systems. For example, chemical weathering is very active in
shallow unconfined aquifers due to the aggressiveness of recharge waters, relatively high
water/rock ratios and high groundwater velocities. Conversely, the groundwater chemistry of
confined systems tends to be rock-dominated (many minerals at saturation) due to low
water/rock ratios and long residence times. In general, groundwater systems in sedimentary
rocks become more confined with depth and the water becomes more saline (high TDS),
anaerobic, and isolated from fresh recharge (Langmuir, 1997). An important corollary of the
decline in dissolved oxygen with depth is that the concentration of NO3- also declines, or is
absent, due to microbial activity under anaerobic conditions. This is discussed further in Section
5.6.
Chapter 5
116
5.3.2 Saturation Indices
The calculation of saturation indices is a purely thermodynamic analysis of departure of a
solution from equilibrium, predicting whether samples are undersaturated, saturated or
supersaturated with respect to a particular mineral phase. This type of analysis can be useful for
understanding vertical and spatial hydrochemical relationships between groundwater samples
due to mineral dissolution and precipitation reactions. Whilst calculation of saturation indices is
theoretical, it can constrain the direction of reaction or the potential of a water to dissolve or
precipitate a mineral. The method provides a useful way to combine element concentrations as
well as chemical parameters such as pH, temperature and Eh into plausible minerals and
examine relative differences between samples. The saturation index, SI, is defined as the ratio of
the reaction quotient Q (or ion activity product) and the thermodynamic equilibrium condition
represented by the equilibrium constant Keq at the given temperature (Freeze and Cherry, 1979).
Therefore, SI = Q/ Keq for the particular mineral of interest. If SI > 1 the water contains an
excess of ionic constituents and thus mineral precipitation potentially occurs; conversely, SI < 1
is indicative of undersaturation and hence mineral dissolution is possible. If SI = 1 the reaction
is at equilibrium, which means that the solution is saturated with respect to the mineral of
interest. In this chapter SI’s are expressed in logarithmic form (i.e. SI = log[Q/ Keq]), in which
case a value of zero denotes the equilibrium condition.
Based on the ionic constituents present, SI’s were derived using the program PHREEQC
(Parkhurst, 1995), using the AquaChem7 graphical interface for the common weathering
minerals expected in an alluvial system derived from a granitic source. These minerals included:
chalcedony, halite, calcite, dolomite, gypsum, magnesite, siderite, albite, anorthite, K-mica,
phlogopite, annite, gibbsite, illite, kaolinite, pyrophyllite, Ca-montmorillonite, and goethite. A
phase diagram for the common aluminosilicates (i.e. log([K+]/[H+]) versus log(H4SiO4)
indicates that the sampled shallow and deep groundwaters primarily plot in the montmorillonite
stability field (Langmuir, 1997). Considering the assumptions inherent in phase diagrams, and
their applicability to natural systems, this suggests that smectite clays are likely to be stable.
Given the low permeability and poor drainage of smectite clays, this observation has
implications for the redox condition of groundwater and N leaching potential (Section 5.6).
Input values for the SI calculations included solute concentrations for each sample, field pH,
field Eh (pe) and temperature. Figure 5-18 depicts the Eh-pH distribution for HSs and HSd
groundwaters. Although there is a greater spatial distribution of samples collected in June 2005,
SI’s were only computed for samples from May 2004 and October 2004 due to unreliable Eh
measurements in June 2005. The calculations indicate that waters of HSs and HSd are generally
undersaturated with respect to the majority of minerals examined. However, all samples are
7 AquaChem v.3.7 for Windows 95/98/NT
Hydrogeochemical Framework
117
essentially saturated with respect to chalcedony (SiO2) (log SI ≥ -0.4) and generally
supersaturated with respect to goethite (FeOOH). Saturation indices for both saturated and
unsaturated phases are compared between nested bores in the following section as a semi-
quantitative tool to examine hydrochemical interaction between aquifers. In Section 5.4, the
extent of lateral groundwater evolution along the flowpath is also verified by the application of
these indices.
5 6 7 8100
200
300
400
500Eh (mV)
pH (field)
HSs samples
Na = ClNa > ClCa-Mg enrichedNa enriched
5 6 7 8100
200
300
400
500Eh (mV)
pH (field)
HSd samples
High salinityLow salinityCa-Mg enrichedNa enriched
Figure 5-18 Field measurements for shallow and deep aquifer samples collected in May and October 2004.
5.3.3 Intra-aquifer relationships
In Chapter 4 it was established that based on the available stratigraphic and hydraulic
information, the subsurface can be conceptualised as a two-aquifer system. In order to support
this interpretation, hydrochemical signatures are examined for bores screened at different
intervals within the same aquifer. Three bores in the study area allow this type of comparison:
bore 46 (intervals 74 and 46A) and bore 51 (intervals 101B and 51A/101A) are screened at two
depths in HSd, whilst bore 54 (54B and 54A) is screened at two depths in HSs (refer to Figure
4-3). With reference to the hydrochemical groups identified in Section 5.2, Schoeller plots
indicate that at the respective bores, the two intervals screened in HSd fall within the envelope
of major ion concentrations of the high salinity chemical group (compare Figure 5-19a and b
with Figure 5-7d). The main hydrochemical difference between the nested intervals at each bore
is largely in the Ca/Mg ratio, which is greater in the deeper interval of HSd. Saturation indices
for the two intervals in bore 101 (pipes A and B) suggest that the deeper interval is more
Chapter 5
118
saturated with respect to both dolomite and calcite but otherwise has identical SI’s for other
minerals (Figure 5-20a). At bore 54, the two screened intervals in HSs represent the Ca-Mg
enriched (shallower) and Na enriched (deeper) chemical groups (compare Figure 5-19c with
Figure 5-5a and b). Whilst these groups of waters are distinct, the two intervals have very
similar SI’s for the calculated minerals, suggestive of a common source of water (Figure 5-20b).
Nonetheless, as highlighted in Section 5.2.3.2, the distinction between the Ca-Mg and Na
enriched groups is important from the perspective of tracking NO3- in groundwaters. This is
discussed further in Section 5.6. Therefore, while the hydraulics indicate that the different sand
units behave as one aquifer, the hydrochemistry is a more sensitive variable at small scale.
On balance it is considered that the hydrochemistry of the three bores examined in this section
provide evidence of good hydraulic continuity between the screened intervals within the same
aquifer, consistent with the bore hydrograph trends (Chapter 4).
a
b
c
Figure 5-19 Schoeller plots for two screened intervals within the same aquifer based on available random measurements (1975 - 2005). (a) and (b) represent screened intervals in HSd while (c) represents screened intervals in HSs (refer to Figure 4-3 in Chapter 4). The red trend lines highlight samples from the shallower screened interval in each aquifer. Source: QDNRW.
shallower
deeper shallower
deeper
deeper
shallower
Hydrogeochemical Framework
119
-6
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6
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Side
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Goe
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Cha
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Gyp
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SI 101B
101A
a
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Cal
cite
Dol
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Gyp
sum
log
SI
54B
54A
b
Figure 5-20 Plots of saturation indices (logarithmic form) for intervals screened within (a) the deep aquifer and (b) the shallow aquifer. The A-pipe is the deeper interval (compare with Figure 5-19). The dotted line denotes the equilibrium condition where SI = 1.
5.3.4 Inter-aquifer mixing trends
Hydrochemical comparisons between nested bores enable site-specific vertical relationships
between the aquifers to be examined. Three broad patterns emerge from Schoeller plots of
nested bores screened in the shallow and deep aquifers (Figure 5-21):
Pattern 1: HSs and HSd have ion concentrations of a different magnitude but the separation
is common for all ions i.e. parallel Schoeller plots;
Pattern 2: HSs and HSd have ion concentrations of the same magnitude for all ions i.e.
coincident Schoeller plots; or,
Pattern 3: HSs and HSd have different concentrations of major ions and different proportions
of each ion i.e. non-parallel Schoeller plots.
Chapter 5
120
As discussed in Section 5.3.1, these trends between nested bores arise because of a combination
of factors, including differences/similarities in aquifer material and reactivity of the rocks;
groundwater residence time; volume of water transmitted; water source; and chemical processes
in shallow versus deep aquifers. Therefore, while parallel Schoeller plots indicate that the
groundwater in each aquifer is of a similar composition and non-parallel trends indicate the
converse, these trends do not necessarily correlate with the degree of inter-aquifer connectivity.
However, in combination with SI’s (Figure 5-22) it is possible to infer whether the observed
hydrochemical trends can be attributed to similarities or differences in source waters or
chemical conditions. Therefore, although calculated SI’s are a function of the numerous factors
listed above, similar trends in SI’s at nested bores are indicative of similar processes and/or
similar water-rock interactions. Whilst the potential degree of vertical connectivity between
aquifers was assessed in Chapter 4 (Section 4.4.2) based on hydraulic information, the
interpretation of hydrochemistry can provide further insight into inter-aquifer processes and
interactions than aquifer hydraulics alone.
Schoeller plots for nested bores 68, 61, 60, 36 and 38 display more-or-less parallel trends for the
majority of ions (Figure 5-21). Whilst the deeper aquifer generally represents the more saline
water type (greater upwards vertical displacement of Schoeller plot), in some bores the trends
overlap and give rise to pattern 2, a special case of pattern 1. The remaining nested bores
depicted in Figure 5-21 display non-parallel trends and represent pattern 3.
Hydrographs for bores in the western half of the catchment indicated a downwards vertical head
gradient and strong to good vertical connectivity between HSs and HSd (Figure 4-7). Schoeller
plots for bores 68, 61 and 60 are consistent with this interpretation: the separation between
trends also indicates that the degree of vertical leakage is greatest in the vicinity of bores 60 and
61. The separation between SI trends (between nested bores for the same minerals) provides
further evidence of the same relative degree of connectivity at each location (e.g. Figure 5-22a
and b). Given the similarity in SI’s, the observed mismatch between Schoeller plots at bore 61
in regards to SO4 and HCO3 is indicative of Eh-pH controls between aquifers, such as the
reduction of SO4 as groundwater moves to the more reduced environment of HSd compared to
HSs. Although the disparate ionic concentrations at bore 62 result in non-parallel Schoeller
trends, SI’s illustrate that there is a similar trend of enrichment and depletion of the calculated
minerals in each aquifer (Figure 5-22c). Thus, while there is a degree of confinement between
aquifers in the vicinity of bore 62, a common recharge source is probable. The greater
enrichment in bicarbonate and other solute concentrations in the deep aquifer is consistent with
longer residence times and/or mixing with older water.
Hydrogeochemical Framework
121
For bores in the eastern half of the catchment, the interpretation of non-parallel Schoeller trends
in relation to the degree of vertical connectivity is complicated by secondary processes and/or
seawater influences, that particularly enrich the deep aquifer in NaCl relative to the shallow
aquifer. Where there is an upwards head gradient, there is the additional influence of direct
rainfall to the shallow aquifer that may dilute the HSd signal from vertical leakage. Based on
hydrograph analysis, poor downwards vertical connection between aquifers was established at
bores 46, 51 (101) and 53. Hydrochemical trends between nested bores are consistent with the
observed hydraulic relationships. The large difference in ionic concentrations between aquifers,
particularly at bores 51 (101) and 46, are indicative of different source waters to each aquifer. It
was shown in Section 5.2.3.1 (Figure 5-16) that the hydrochemistry of these bores is consistent
with seawater intruding into the aquifer sediments.
Based on their historical water level elevations, upwards vertical connectivity between HSs and
HSd was evident at bores 48, 49, 36 and 38. Schoeller and SI plots (Figure 5-21 h-k and Figure
5-22 d-f) illustrate that while there is the potential for exchange of water between aquifers, there
is clearly a degree of confinement in the vicinity of bores 48 and 49: poor connectivity is
established. In contrast, the near-coincident Schoeller plots at bores 36 and 38 and identical SI’s
imply that the same waters are present in each aquifer at these locations. Given that the head
gradient is upwards in these bores, the hydrochemistry is consistent with a large component of
vertical discharge from HSd to HSs.
The hydrochemical analyses highlight that differences between Schoeller plots for most nested
bores provide a good indication of the relative degree of vertical hydraulic continuity between
aquifers. This is useful given that the calculation of SI’s requires reliable field measurements.
However, as exemplified by bore 62, disparate Schoeller plots in isolation of SI calculations
may lead to misinterpretation. Hydraulic relationships are also important for interpreting
hydrochemical trends.
Chapter 5
122
a b
c d
e f
Figure 5-21 Schoeller plots for groundwater samples from shallow (blue) and deep (black) nested bores in June 2005. The corresponding bore hydrographs indicate that vertical connectivity is: (a-d) strong/good (downwards head gradient); (e-g) poor (downwards head gradient), (h-i) poor (upwards head gradient); and (j-k) good (upwards head gradient) (refer to Figure 4-7). Note the change in vertical scale in the plots.
Hydrogeochemical Framework
123
cont.
g h
i j
k
#S
#S
#S
#S
#S
#S
#S
#S#S
#S
#S6861
60
Herbert R
4662
3853
51
48
49 36
Figure 5-21 Schoeller plots for groundwater samples from shallow (blue) and deep (black) nested bores in June 2005. Refer to inset map for bore locations.
Chapter 5
124
-10
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2
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68A
a
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SI 61B
61A
b
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SI
62B
62A
c
Figure 5-22 Plots of saturation indices (logarithmic form) for selected nested bores based on samples collected in October 2004 (compare with Figure 5-21). The dotted line denotes the equilibrium condition (SI = 1). Graphs a-c represent bores with a downwards vertical head gradient.
Hydrogeochemical Framework
125
cont.
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48B
48A
d
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e
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36B
36A
f
Figure 5-22 Plots of saturation indices (logarithmic form) for selected nested bores based on samples collected in October 2004 (compare with Figure 5-21). The dotted line denotes the equilibrium condition (SI = 1). Graphs d-f represent bores with an upwards vertical head gradient.
Chapter 5
126
In addition to the graphical methods used in this section to examine vertical connectivity
between the aquifers, semi-quantitative estimates of the proportion of vertical mixing can be
determined by simple mass balance calculations based on solute concentrations. As this
approach assumes conservative mixing between two end-members, the reliability of the
estimates is a function of how well the end-member chemical compositions can be constrained
(Herczeg and Edmunds, 2000). The hydrochemistry of the most dilute HSd waters (in the
northwest of the catchment, Figure 5-15) and that of HSs at the nested bore of interest are
plausible end-members. However, given the semi-confined nature of the deep aquifer, mixing
proportions at any bore represent an overestimate. Therefore, due to difficulties in defining the
end members, the percentage of mixing could not be estimated by this method. More
sophisticated geochemical modelling techniques (Parkhurst, 1995) are considered beyond the
scope of the thesis.
5.3.5 Relationship with the bedrock
A comprehensive search of the QDNRW groundwater database was carried out to select bores
screened in the bedrock which contain stratigraphic logs, screened intervals of known depth and
water chemistry data. The majority of deep bores in the study area and surrounding region are
screened in the alluvial aquifer, with only a few bores screened in weathered granitic bedrock:
no bores are screened exclusively in unweathered bedrock. Schoeller plots indicate that
although groundwater in the weathered bedrock is spatially variable, there is a similar
hydrochemical pattern to the deep alluvial aquifer at each location. The lack of a distinct
hydrochemical signature suggests there is vertical connectivity between the alluvial aquifer and
weathered zone of granitic bedrock; however, the relationship with underlying unweathered
bedrock is unclear.
5.4 SPATIAL RELATIONSHIPS WITHIN AQUIFERS
Lateral relationships in each aquifer are examined in this section to verify the inferred flow
paths established in Chapter 4 (Figure 4-18). The spatial distribution of the main hydrochemical
groups in each aquifer was illustrated in pie charts in Figure 5-15 and Figure 5-17. In Chapter 4
it was noted that with the exception of a marine mud at the coastal fringe, sedimentation in the
case study area occurred from terrestrial deposition (Section 4.2). Given the geological setting
and basement rock, terrestrial sediments are likely to have originated from an igneous source.
As such, the chemical evolution of groundwater is likely to be dominated by the dissolution of
feldspars, micas and other silicate minerals, with the release of Na+, K+, Mg2+ and Ca2+ ions to
the water (Freeze and Cherry, 1979). The key processes that influence the chemistry of
groundwater in both unconfined and confined aquifers were outlined in Sections 5.1.1 and 5.3.1.
Hydrogeochemical Framework
127
5.4.1 Lateral hydrochemical evolution
5.4.1.1 Deep aquifer
A shift from Ca-Mg enriched groundwaters dominated by HCO3 in the northwest of the
catchment, to high salinity waters with seawater-type signatures towards the coast is observed in
the deep aquifer (Figure 5-15). Low salinity group waters with Cl > HCO3 dominate the middle
section of the catchment. By considering the major ions which each comprise greater than 20 %
of the total cation and anion concentrations respectively (expressed in meq/L), the spatial anion
and cation pie charts can be combined into a single hydrochemical facies map (Figure 5-23a).
Analysis of water level contours in Chapter 4 indicated that a major rainfall-recharge zone for
the deep aquifer is present in the northwest of the catchment, while discharge occurs towards the
coast (Figure 4-15b). In accordance with the principles of hydrochemical evolution discussed in
Section 5.1.1.2, there is evidence of a general trend from HCO3-dominated waters with low
TDS values in the main recharge area (e.g. bores 71A, 69A) to Cl-dominated waters with high
TDS values towards the discharge area (e.g. bores 48A, 38A, 36A) during both the beginning
and end of the dry season (e.g. Figure 5-15). However, examination of hydrochemical trends
along groundwater flowpaths in HSd, such as along the main flowpath depicted in Figure 5-23,
indicates that there are clearly other processes occurring that interrupt lateral hydrochemical
development. For example, it is apparent that while there is an increase in TDS away from the
main recharge area in the northwest, waters with TDS less than 300 mg/L are observed further
down the flowpath from bore 68A (Figure 5-23b). In light of the inter-aquifer relationships
established in Section 5.3.4, these dilute waters are consistent with enhanced vertical leakage
from the shallow aquifer and/or proximity to a local recharge area (e.g. north of the Herbert
River, refer to Figure 4-15b). Furthermore, an increase in TDS between bores 61A and 59A is
consistent with groundwater contributions from the southwest.
Evidence of both vertical and lateral recharge sources is provided by stable isotope
measurements. As illustrated in Figure 5-24, there is a progressive enrichment in stable isotopes
from the beginning of the main flowpath (bore 69A) to bore 61A. Given the generally enriched
isotopic signature of the shallow aquifer over the deep (Figure 5-3), this trend along the
flowpath is consistent with isotope enrichment due to mixing between the shallow and deep
aquifers (as discussed in Section 5.3.4). However, a decline in heavy isotopes between bores
61A and 59A is considered to be due to groundwater from the upper Stone River valley, which
is more depleted compared to waters from the northwestern recharge area (Figure 5-3). Note
that the concentration of stable isotopes in groundwater depends mainly on the origin of water:
effects of water-rock interactions on water isotopic composition generally become important
only at high temperature (Gonfiantini et al., 1998).
Chapter 5
128
a
#S
#S
#S#S
#S
#S#S#S
#S
#S
#S
#S
#S
#S
#S
#S#S #S#S
#S
#S
#S
#S
#S
#S
#S
#S
216652
166138
392
260
578
220 794
128
134
136
704
634
384
1076
1050
2416
2662
B: 3644
24941670
1048
3434
14120
A: 4924
> 3000#S
< 300#S
300 - 750#S
2000 - 3000#S750 - 2000#S
TDS (mg/L)
5 0 5 Km
N
HSd TDS
71A69A
68A
67A
61A
62A
63A 58A
59A
60A
52A53A
45A
44A38A
37A46A
47A
36A48A
49A
101A
130A
127B
101B
Stone R
Herbert R
b
Figure 5-23 (a) Hydrochemical facies and (b) distribution of total dissolved solids determined at bores in the deep aquifer in June 2005. Selected flowpaths from the northwest to the coast as well as from the Stone River are also depicted. Note that A and B refer to different screened intervals within the same aquifer (interval A is the deepest). The shaded areas in (a) correspond to the major water types identified: Ca-Mg enriched (green); Na enriched (blue); low salinity (purple); high salinity (red) (refer to Figure 5-15). QDNRW monitoring bores are labelled in (b) (domestic bores are not labelled).
Hydrogeochemical Framework
129
Given that HSd groundwaters are of relatively low temperature (maximum 30oC) it can be
assumed that stable isotopes of water are unaffected by secondary processes. Therefore, the
observed trends are consistent with contributions from isotopically different waters.
-34
-32
-30
-28
-26
-24
-22
69 68A 67 61A 59
HSd bores down the flowpath
Deut
eriu
m (
o / oo,
SM
OW
)
-5.6
-5.4
-5.2
-5
-4.8
-4.6
Oxy
gen-
18 (
o / oo,
SM
OW
)
Figure 5-24 Stable isotopic values along a flowpath for the deep aquifer based on samples collected in May 2004.
Hydrochemical evolution is further obscured in the eastern part of the catchment, where there is
a dramatic increase in TDS values. The corresponding Na-Cl type waters are indicative of a
different aquifer material or mixing with higher salinity waters. From the analysis of ion ratios
(e.g. Figure 5-10 and Figure 5-16) and stable isotope evidence (Figure 5-11), the observed TDS
increase is attributed to mixing with seawater. However, as illustrated in Figure 5-23b, there is a
decrease in TDS at the coast; as such, the spatial distribution of TDS does not correspond with
the present-day coastline. Possible mechanisms to account for these observations are proposed
in Section 5.4.2.
In agreement with the observed TDS distribution, SI’s for the calculated minerals (Section
5.3.2) increase immediately down-flow from the main recharge area (e.g. between bore 69A and
68A). In addition, in the high TDS zone (e.g. bore 47A) there is a marked increase in SI’s,
particularly with respect to gypsum and magnesite: this is consistent with a seawater influence.
Although groundwaters are generally undersaturated with respect to the calculated mineral
phases, fluctuations in SI’s generally mimic major ion and TDS trends in response to influences
from dilute and enriched groundwaters contributing along the flowpath. Importantly, the main
flowpath depicted in Figure 5-23 is a sink for groundwater from different areas in the
catchment. Therefore, whilst there is overall an increase in solute concentrations and TDS from
the main recharge area to the coast, the degree of hydrochemical evolution is suppressed
because of the vertical connectivity between aquifers and the convergence of flowpaths
contributing groundwaters of different compositions.
Chapter 5
130
5.4.1.2 Shallow aquifer
A hydrochemical facies map for the shallow aquifer highlights the distribution of Na-HCO3-Cl
facies groundwater in the west, interspersed with Na-Cl-HCO3 waters in the east; Na-Cl facies
groundwaters are restricted to the northeast (Figure 5-25a). Note that groundwater at bore 101C
(e.g. Figure 5-25b) is classified as belonging to the Na-HCO3-Cl facies (with low TDS) during
other sampling periods, rather than the high salinity Na-Cl type depicted in Figure 5-25a.
Similarly, low TDS measurements have also previously been obtained at bore 57A (QDNRW
database). Given the instability of EC readings during pumping, contamination from the deep
aquifer is possible at both of these bores, although further sampling would be required to
confirm this.
Analysis of water level contours in Chapter 4 indicated that recharge to HSs occurs from both
diffuse and lateral sources (Figure 4-17b). The interpretation of both diffuse and lateral recharge
is supported by available stable isotope data for the shallow aquifer (Figure 5-14). Waters lying
below the LMWL reflect evaporation and therefore slow infiltration during diffuse recharge. In
contrast, waters lying on the LMWL represent rapid infiltration to the shallow aquifer (lateral
and/or diffuse recharge). Given these recharge mechanisms the lack of obvious hydrochemical
evolution is not unexpected. For example, there is no systematic increase in TDS along the main
flowpath depicted in Figure 5-25b. Although groundwater from the northwest becomes more
saline down the flowpath, there is a decrease in TDS at bore 53B. However, the slight increase
in TDS between bores 53B and 46B, as well as between 53B to 48B (and further to the
northeast), are indicative of minor hydrochemical development. It was noted in Section 5.3.3
that the Ca-Mg and Na enriched waters represent different sandy units within the shallow
aquifer. Hydrochemical evolution is more pronounced within the Na enriched unit, as
highlighted by a hypothetical flowpath between bores 68B-60B-50A-49B (Figure 5-25b). This
unit representing Na enriched waters is at a greater depth and has a higher clay content than the
unit representing Ca-Mg waters. These attributes may result in a slightly longer residence time
and hence enhanced TDS evolution along the flowpath. However, along the main flowpath
depicted in Figure 5-25b there is little evidence of hydrochemical evolution. Given that HSs is
characterised as a watertable aquifer, evaporation may affect the hydrochemistry and thus there
may be localised areas of high TDS. The absence of strong evidence for hydrochemical
evolution in the upper sandy unit is consistent with relatively short residence times and
therefore less opportunity for water-rock interactions and other processes to influence the
chemistry of the groundwater.
Hydrogeochemical Framework
131
a
#S
#S
#S#S
#S#S
#S
#S
#S
#S
#S
#S#S
#S
#S
#S
#S
#S
#S#S#S
#S
#S
#S
#S
82
258140
120710
144
154176
114
142
B: 152A: 292
122206
320
172
112
210218
282
576
256
1152
1000
< 150#S
150 - 300#S
800 - 1200#S300 - 800#S
TDS (mg/L)
5 0 5 Km
N
HSs TDS
68B
61B
62B
60B
53B
38B
46B
36B
48B
49B
130B
101C
57A54B54A
50A
b
Figure 5-25 (a) Hydrochemical facies and (b) distribution of total dissolved solids determined at bores in the shallow aquifer in June 2005. Potential flowpaths are depicted. Note that A and B refer to different screened intervals within the same aquifer (interval A is the deepest). The shaded areas in (a) correspond to the major water types identified: Ca-Mg enriched (green); Na enriched (blue); Na > Cl (purple); Na = Cl (red) (refer to Figure 5-15). QDNRW monitoring bores are labelled in (b) (domestic bores are not labelled).
Chapter 5
132
5.4.2 Seawater intrusion
The preceding analyses have identified water types in each aquifer that have high salinity and/or
seawater-type signatures. Based on major element relationships and stable isotope evidence
(Section 5.2), seawater intrusion into the deep aquifer is proposed, while meteoric influences are
suggested for the shallow aquifer. In coastal areas, fresh groundwater derived from precipitation
on the land comes in contact with and discharges into the sea or into estuaries containing
brackish water (Heath, 1987). Therefore, depending on the height of the watertable and
thickness of the freshwater lens, saltwater intrusion can influence the salinity of an aquifer. An
estimate of depth of the saltwater interface as a function of distance from the coast can be made
by application of the Ghyben-Herzberg relation; this represents a theoretical relationship that
generally leads to an underestimate of the depth to the saltwater interface due to the assumption
that the freshwater and seawater zones are static (Freeze and Cherry, 1979).
Groundwater elevations of bores screened in the unconfined aquifer (HSs) were used to depict
the theoretical saltwater wedge below freshwater along three transects (Figure 5-26). Two
periods were chosen, May and October 2004, to approximately represent maximum and
minimum watertable elevations: this translates to minimum and maximum elevations of the
interface, respectively. Comparison of the saltwater interfaces with slotted depths of bores along
transects 1, 2 and 3 indicates that HSd bores within approximately 10 km from the coast are
screened within the seawater zone, and thus have the potential to be intruded by seawater
(Figure 5-27). Whilst this includes bores 37A, 48A and 49A, the estimates suggest that the other
high salinity bores such as 46A, 47A and 101A are screened within the freshwater zone.
However, given that in reality the interface separating freshwater and saltwater is not a sharp
boundary, there may be mixing of saltwater and freshwater in a zone of diffusion around the
interface (Freeze and Cherry, 1979). This is consistent with trend 1 in Figure 5-16.
Figure 5-26 Transects for estimating the theoretical saltwater wedge below freshwater according to the Ghyben-Herzberg relation (Figure 5-27). The red line encloses high salinity type waters. Bores are labelled without their 116000 prefix.
#S
#S
#S
#S
#S
#S#S
#S#S
#S
#S
#S
#S
4737Herbert
R
46
4438
101
48
49 36
53
61
45
62
Transect 1
Transect 2
Transect 3
Hydrogeochemical Framework
133
The spatial distribution of TDS in HSd at the coast (Figure 5-23b), as decreasing from north to
south, is consistent with seawater intrusion from a source from the north, rather than from the
east. Water level contours presented in Chapter 4 indicated a northeasterly trending groundwater
flowpath near the coast (e.g. Figure 4-14a) that may also be a preferential flowpath for the
inflow of seawater to the deep aquifer. Alternative explanations for the source of salinity in the
deep aquifer could be relic seawater in the aquifer sediments or a seawater mixing zone around
the river. Characteristics of residual evaporated seawater include: high TDS (> 35 g/L); Ca-
chloride composition (Ca/(SO4 + HCO3) > 1); molar Na/Cl ratio below that of modern seawater
(0.86); Br/Cl ≥ seawater ratio (1.5 x 10-3) and relative depletion of sulfate (SO4/Cl < 0.05)
(Vengosh, 2004). Whilst not all factors are satisfied by any of the bores in the study area, the
high TDS, low Na/Cl ratio (average 0.6) and Ca-Cl character of bores 101A (51A) and 49A
(Figure 5-23) are indicative of relic seawater in at least these locations. Note that evaporites
have not been described in the aquifer sediments, nor are there characteristically low Br/Cl
ratios to suggest groundwater flow through evaporites (Herczeg and Edmunds, 2000).
Membrane effects such as salt filtering generally occur at depths of greater than 500-1000 m
below the ground surface (Freeze and Cherry, 1979) and are therefore considered unlikely in
HSd, which has a maximum depth of approximately 100 m below the surface.
The estimates of the saltwater interface suggest that coastal bores 36A and 38A are also
intruded by seawater. While available stable isotopic evidence at bore 36A does not exclude this
possibility (e.g. Figure 5-11), these bores have a characteristically low TDS compared to bores
further west. Note that there is an upwards vertical head gradient towards the coast, which
would preclude leakage of freshwater from the shallow aquifer to the deep. As illustrated in
Figure 5-23a, there are two compositional groups distinguished within the Na-Cl facies.
Specifically, bores 53A, 36A and 38A are low salinity-type waters (purple shading) with a
Na/Cl molar ratio ranging from 1.1-1.5. In contrast, the other high TDS bores in the Na-Cl
facies (red shading) have a lower Na/Cl ratio, between 0.8-0.96. As discussed in Section 5.2.3.1,
two trends can be observed as a result of interactions with seawater, arising from the intrusion
of seawater into aquifer sediments or displacement of seawater by freshwater. Based on the ratio
between Na/Cl and (Ca+Mg)/Cl, it is therefore plausible that the low salinity trend can be
explained by the latter process, consistent with trend 2 in Figure 5-16. Furthermore, given the
upwards vertical head gradient and strong vertical connectivity between the shallow and deep
aquifers in the coastal bores (Section 5.3.4), the possibility exists for sites of local groundwater
discharge in the vicinity of bores 36 and 38, where seawater is forced back by freshwater.
Chapter 5
134
Theoreticaldepth May 2004
TheoreticaldepthOctober 2004
Slotted depthHSd
Slotted depthHSs Transect 1
-500
-400
-300
-200
-100
0
051015202530
Dept
h to
inte
rface
(m A
HD)
-500
-400
-300
-200
-100
0
Slot
ted
dept
h of
bor
es (m
AHD
)
seawater zone
freshwater zone
48B101C
36A
36B
49A48A
49B
101B
101A
61B
high salinity bores
-650
-550
-450
-350
-250
-150
-50
50
051015202530
Dep
th to
inte
rface
(m A
HD)
-650
-550
-450
-350
-250
-150
-50
50
Slo
tted
dept
h of
bor
es (m
AH
D)
seawater zone
freshwater zone
37A
38B
38A
46B47B46A
47A
53B
53A
Transect 2
high salinity bores
-250
-200
-150
-100
-50
0
50
051015202530
Distance from coast (km)
Dept
h to
inte
rface
(m A
HD)
-250
-200
-150
-100
-50
0
50
Slot
ted
dept
h of
bor
es (m
AH
D)
seawater zone
freshwater zone
49A
48B
48A
46B
47B
46A
47A
44A 45A
Transect 3
high salinity bores
49B
Figure 5-27 Theoretical seawater interface during two periods in 2004 (Ghyben-Herzberg relation) and slotted depths of bores in HSs and HSd as a function of distance from the coast along three transects (refer to map in Figure 5-26).
W E
W E
S N
Hydrogeochemical Framework
135
Comparison of the theoretical saltwater interface with the slotted depth of HSs bores suggests
that the shallow aquifer is not affected by seawater intrusion. Therefore, the high TDS and
seawater-type signature of bores 36B and 38B (Figure 5-25) must be accounted for by other
mechanisms. It was discussed above (Section 5.3.4) that there is evidence of vertical discharge
from the deep aquifer; the influence of saline coastal rainfall/sea spray are also plausible sources
of salinity. The hydrograph patterns for nested intervals at bores 36 and 38 (Figure 4-9, Chapter
4) indicated that in addition to recharge from the deep aquifer there is an enhanced signal in the
shallow aquifer due to direct rainfall recharge. Therefore, the hydrochemistry of HSs in these
coastal bores is a reflection of both coastal rainfall and groundwater from HSd. Two bores
located adjacent to bore 36B belong to the same hydrochemical facies but have a much lower
TDS (Figure 5-25). As these bores are located within a few hundred metres of the Herbert River
estuary, they are considered to be influenced by the chemistry of the river and the tidal pattern.
Evidence of this is provided by one of the bores which varied from a fresh Ca-Mg type at low
tide to a more saline Na = Cl type water at high tide.
5.5 NITROGEN IN GROUNDWATER
As discussed in Chapter 1 (1.4.1), nitrogen (N) is commonly found in groundwaters beneath
sugarcane growing areas, due largely to fertiliser inputs. Given that sugarcane farming is a
major land use in the study area and that all bores sampled were located in these cropping areas,
it is plausible that N is an important anion in the groundwater. This section explores the
concentration and speciation of N in the two aquifers in order to characterise the N signal of
groundwater. Subsequent chapters examine the N signal in surface waters and consider potential
N contributions from groundwater based on river-aquifer interactions. General principles
regarding nitrogen chemistry and transformation reactions were described in Section 5.1.1.4.
Nitrogen leaching to the subsurface has previously been investigated in the lower Herbert River
catchment, based on recharge estimates and measurements of nitrate in the upper part of the
shallow aquifer along a transect covering the three main geomorphological units (Bohl et al.,
2000b). The spatial distribution of N leaching losses was further investigated based on
pedological and hydrological features in a subcatchment of the Herbert River valley (Bohl et al.,
2001). These studies found that the highest leaching losses are expected to occur on the more
freely draining soils of the alluvial fans and the sand riverbank soils, while the lowest losses are
estimated for the heavy clay soils on the floodplain. While these previous studies yielded
estimates of N leaching potential, the estimates were based on limited spatial coverage of
measured N just below the watertable. Furthermore, potential leaching to deeper groundwater
was not considered as part of the mass balance. Since the time of these previous studies, more
detailed soil mapping of the Herbert valley has been undertaken (Wood et al., 2003). Seven
broad categories based on landscape position and formation are identified, as depicted in Figure
Chapter 5
136
5-28. It can be observed that clay soils dominate the central part of the study area, while alluvial
soils dominate the riverbank of the Herbert River. Based on this 2003 soil data and
understanding of the hydrogeological and hydrogeochemical characteristics, the spatial
distribution of N within each aquifer and partitioning of N between the aquifers are examined
below.
Streams
Terrace loamy soils
Hillslope soils
Sandy soils
Seymour soils
Clay soilsAlluvial soils
5 0 5 Km
N
Figure 5-28 Distribution of mapped soil types in the study area (Wood et al., 2003). Note that the Seymour soils have a high proportion of fine sand derived from acid volcanic parent material. Source: HRIC
5.5.1 Spatial distribution of N
The spatial distribution of total dissolved inorganic forms of nitrogen (DIN) is examined in this
section within the shallow and deep aquifers. Ammonium (NH4+), nitrite (NO2
-) and total
dissolved oxides of nitrogen (NOX) were analysed in all groundwater samples collected: nitrate
(NO3-) is calculated by difference from NOX and NO2
-. DIN is calculated as the sum of the
concentrations of dissolved NO3-, NH4
+ and NO2- (each as mg N/L).
5.5.1.1 Shallow aquifer
As shown in Figure 5-29a, the concentration of DIN is variable throughout the shallow aquifer.
Although the spatial distribution appears random, in general, bores recording the highest
concentrations of DIN are associated with Ca-Mg enriched groundwaters (Figure 5-17).
Comparison with the distribution of soil types shows that bores with DIN > 3 mg N/L are
located on either the alluvial or terrace loamy soils. In contrast, shallow bores with the lowest
Hydrogeochemical Framework
137
measured concentrations of DIN are mostly located on clay soils and are associated with Na
enriched groundwaters (Table 5-1). Note that these shallow groundwaters also plot within the
stability field for smectite clays (Section 5.3.2). The alluvial soils are generally sandy and
permeable, while the clay soils range from silty to heavy clays; terrace loamy soils are similar to
the clay soils although they are situated higher in the landscape and have better structure and
drainage (Wood et al., 2003). The alluvial soils and sandy terrace loams are considered to have
a high risk of leaching due to their well draining nature. Conversely, nitrogen losses by
denitrification due to intermittent waterlogging (anaerobic conditions) are characteristic of the
clay dominated soils (Wood et al., 2003). Therefore, given the soil hydraulic properties and
spatial distribution of DIN in HSs, it is considered that soil type has a major influence on the
amount of N leached to the shallow aquifer. This is consistent with previous studies (Bohl et al.,
2001).
5.5.1.2 Deep aquifer
Figure 5-29b shows that there are three distinct zones in the deep aquifer in regards to the
concentration of DIN. The upper part of the Herbert River valley has Ca-Mg enriched
groundwaters with the highest concentration of DIN, as also shown in Figure 5-15. Further
down-valley, in the middle section of the catchment, groundwater has very little DIN. Whilst
good vertical connection with HSs is evident in some bores, including shallow bores with high
DIN, the near-absence of DIN at depth indicates N loss due to transformations to other species
of nitrogen. For example, denitrification results in the formation of nitrous oxides and nitrogen
gas (Section 5.1.1.4), which were not measured in the samples collected.
Towards the coast, the concentration of DIN is observed to increase in the deep aquifer,
corresponding with Cl-dominated groundwaters. Similar to the TDS distribution (Figure 5-23b),
the highest concentration of DIN in this area is observed in the northeast at bore 49 (0.81 mg
N/L) and progressively declines in groundwater to south. This suggests there may be a
hydrochemical control on the concentration and spatial distribution of DIN in the deep aquifer.
Chapter 5
138
a
b
Figure 5-29 Spatial distribution of dissolved inorganic nitrogen (DIN) (mg N/L) in bores screened in the (a) shallow and (b) deep aquifers in June 2005. A and B refer to different screened intervals considered to be part of the same aquifer at that bore location. The shaded areas in (a) correspond to the major water types identified in Section 5.2.1 (refer to Figure 5-15 and Figure 5-17).
Hydrogeochemical Framework
139
5.5.2 Speciation of N
Whilst the concentrations of the three inorganic nitrogen forms were determined, results
indicate that NO2- is below detection (< 0.002 mgN/L) in groundwaters from HSd and just above
the detection limit in some samples from HSs. Therefore, hydrochemical trends regarding the
other two forms of DIN (NO3- and NH4
+) are the focus of this section.
5.5.2.1 Shallow aquifer
As illustrated in Figure 5-30a, the concentration of NO3- in the shallow aquifer ranges from
below detection up to almost 90 mg/L. In contrast, NH4+ is a relatively minor component of
HSs; the majority of bores have NH4+ concentrations of less than 0.1 mg/L (Figure 5-30b).
Therefore the distribution of DIN observed in Figure 5-29a is largely due to NO3-, which is the
dominant species of N in the shallow aquifer. In contrast to nitrate, the ammonium form of
nitrogen is less mobile in the subsurface environment, and is therefore less likely to be
transported through the unsaturated zone into groundwater. Key processes that inhibit this
transport are adsorption, cation exchange, incorporation into microbial biomass or release to the
environment as a gas. Adsorption onto soil particles is considered to be the major mechanism of
removal of NH4+ in the subsurface environment (Canter, 1997).
In general, an inverse relationship between the two dominant forms of inorganic N is observed
in HSs, such that high NO3- waters have a low concentration of NH4
+ and high NH4+ waters have
low NO3- (compare Figure 5-30a and b). The redox control on the concentration of NO3
- and
NH4+ in the shallow aquifer is clearly illustrated in Figure 5-31: high NO3
- is associated with
more oxidised waters (high Eh), whilst more reduced waters (low Eh) are dominated by NH4+.
These trends are consistent with the respective oxidation states of the nitrate ion (+5) versus that
of the ammonium ion (-3). Where there is sufficient oxygen, an inorganic carbon source, and
specific chemoautotrophic bacteria, biological oxidation (nitrification) of NH4+ produces NO3
-
that is leached to shallow groundwater (equation 5-1, Section 5.1.1.4). In contrast, low NO3-
groundwater is consistent with only minor leaching of NO3- and/or denitrification due to
reduced conditions. Therefore, while soil type influences the amount of DIN leached to the
shallow aquifer (as discussed in Section 5.6.1.1), the redox condition of the groundwater is a
major control on the speciation of N in HSs. Speciation of N is also dependent on the pH
(Appelo and Postma, 1994); however, over the limited pH range of HSs waters (5 - 7) there is no
obvious relationship.
Chapter 5
140
#S
#S
#S#S
#S#S
#S
#S#S
#S
#S
#S#S
#S
#S
#S
#S
#S
#S#S#S
#S
#S
#S
#S
#S#S
#S#S
#S
#S#S
#S
#S
#S#S
#S
#S
#S
#S
#S
#S#S#S
#S
#S
#S
0.03
8.23
0.330.01
A: 2.21
4.020.01
0.31
8.452.57
0.01
27.44
40.06
46.46
38.51
15.71
B: 20.14
87.8715.43
19.26
22.35< 0.01
< 0.01
< 0.01
< 0.3#S0.3 - 4#S
15 - 20#S4 - 15#S
NO3 (mg/L)
5 0 5 Km
N
HSs NO3
#S >20
a
%U
%U
%U%U
%U%U
%U
%U%U
%U
%U
%U%U
%U
%U
%U
%U
%U
%U%U%U
%U
%U
%U
%U
%U
%U
%U%U
%U
%U%U
%U
%U
%U
%U
%U
%U%U
%U
%U
%U
0.34
1.42
0.23
0.020.09
0.010.08
0.01
A: 0.01
0.04
0.02
0.01
0.010.20
0.01
0.47
0.03
< 0.01
< 0.01
< 0.01< 0.01
< 0.01
< 0.01
B: < 0.01
0 - 0.01%U0.01 - 0.1%U
0.5 - 1.5%U0.1 - 0.5%U
NH4 (mg/L)
5 0 5 Km
N
HSs NH4
b
Figure 5-30 Spatial distribution of (a) NO3- (mg/L) and (b) NH4
+ (mg/L) in bores screened in the shallow aquifer in June 2005, with selected flowlines depicted. A and B refer to the two screened intervals considered to be part of the same aquifer at bore 54.
Hydrogeochemical Framework
141
0.0001 0.001 0.01 0.1 1. 10.100
200
300
400
500Eh (mV)
NO3 (meq/l)0.0001 0.001 0.01 0.1
100
200
300
400
500Eh (mV)
NH4 (meq/l)
HSs samples
Na = ClNa > ClCa-Mg enrichedNa enriched
Figure 5-31 Bivariate plots for shallow groundwater samples (2004 samples only), displayed according to the compositional groups determined in Section 5.2.1.
It was noted in Chapter 4 (Section 4.2.2) that there are two sandy units within the shallow
aquifer of varying lateral extent. Whilst the two units display similar hydraulic behaviour, they
are distinct hydrochemically (Section 5.3.3), including with respect to N. Thus, bores screened
in the upper unit, associated with oxidised Ca-Mg enriched waters, are observed to have a high
concentration of NO3- (Figure 5-31). In contrast, groundwater in the deeper unit of HSs,
associated with more reduced Na enriched waters, generally has less than 1 mg/L NO3- and a
detectable concentration of NH4+ (Figure 5-30). Given the high mobility of NO3
- and high
vertical connectivity between the two sandy units in the shallow aquifer (Section 5.3.3),
leaching of NO3- to the deeper unit of HSs would be expected. The near-absence of NO3
- and
NO2- in this deeper unit is consistent with reduction and ultimate denitrification to nitrogen gas,
N2. Vertical movement of N is discussed further in Section 5.6.3.
Chapter 5
142
5.5.2.2 Deep aquifer
As shown in Figure 5-32, N speciation in the deep aquifer is characterised by NO3- in the upland
area to the northwest and NH4+ towards the east. The same trend is noted during other sampling
periods (May and October 2004). In the intervening zone, the concentrations of both NO3- and
NH4+ are low, with no particular dominant species. Therefore, the observed spatial pattern of
DIN (Figure 5-29b) is due to different N species dominating in different parts of the catchment.
The spatial extent of NO3- and NH4
+ enriched waters are associated with Ca-Mg and Na-Cl type
waters respectively (compare Figure 5-32 with Figure 5-29b). Therefore, as suggested in
Section 5.6.1.2, a change in hydrochemical conditions is a plausible explanation for the
speciation difference up-valley compared to on the floodplain.
As for the shallow aquifer, high NO3- concentrations in the northwest are consistent with
oxidising conditions (nitrification) and hence leaching of NO3- to the deep aquifer. Given the
general direction of groundwater flow in HSd (Figure 4-14, Chapter 4), the speciation trends
down gradient of the northwestern recharge area are consistent with reduction of N. Increases in
HCO3 and pH in the intervening zone (low NO3- and NH4
+) indicate reduction of nitrate by
organic matter (denitrification), while elevated concentrations of ferrous iron and sulfate in HSd
in the eastern area (low NO3- and higher NH4
+) are consistent with reduction of N coupled with
pyrite oxidation or reduction by other Fe2+ bearing minerals (Appelo and Postma, 1994).
Nitrogen gas is a product of these reduction processes, which was not specifically measured in
the groundwater samples. Therefore, at least in the middle of the catchment, it is plausible that
DIN is present as N2, which remains in solution until the groundwater discharges to a surface
water body and equilibrates with the atmosphere (Thayalakumaran et al., 2004). Given that N2
is the end product of denitrification, an additional source of N to HSd is required to account for
the observed increase in NH4+ in the lower part of the catchment. It was established above that
seawater intrusion influences the hydrochemistry of HSd in this area. Based on the
hydrochemistry of seawater at Moreton Bay, where the concentration of NO3- was measured at
50 mg/L (Cresswell 2006, pers. comm.), it is therefore plausible that seawater provides the
additional source of N. Dissimilatory nitrate reduction to ammonium (DNRA) (Silver et al.,
2001; Korom, 1992) may thus have a role in the transformation of N. Alternatively, an increase
in ammonium could be attributed to desorption from clays under reduced conditions.
Verification of the factors or processes responsible for the observed speciation/concentration
trends of DIN is beyond the scope of this research. However, it is noted that measurement of
gaseous forms of N; redox indicators; microbiology; and availability of electron acceptors and
donors, could assist in future investigations. Nitrogen isotopes can also provide insight into N
processes (Kendall, 1998).
Hydrogeochemical Framework
143
#S
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0.1 - 10#S
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10 - 20#S
NO3 (mg/L)
5 0 5 Km
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HSd NO3
a
%U
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0.01 - 0.1%U
0.5 - 1.5%U
0.1 - 0.5%U
NH4 (mg/L)
5 0 5 Km
N
HSd NH4
b
Figure 5-32 Spatial distribution of (a) NO3- (mg/L) and (b) NH4
+ (mg/L) in bores screened in the deep aquifer in June 2005. A and B refer to different screened intervals considered to be part of the same aquifer at that bore location. A selected flowline is depicted.
Chapter 5
144
5.5.3 Nitrogen transport
As discussed in Chapter 3 (Section 3.2.1), fertiliser applied to sugarcane is considered to be a
major source of DIN in groundwaters of the study area. Local hot spots of DIN in bores
adjacent to garden beds also provide evidence of point source contributions. Urea ([NH2]2CO)
and/or ammonium sulfate (NH4SO4) are commonly applied crop fertilisers in the study area,
which are mineralised by microorganisms in the soil to NH4+ (Wood et al., 2003). As depicted
in the nitrogen cycle (Figure 5-1), NH4+ has a number of fates: volatilisation, adsorption onto
clay minerals or nitrification to NO3-. In the absence of N measurements in soils of the study
area, it is not possible to determine the proportion of NH4+ bound to soil particles versus
atmospheric losses or transformations to other forms such as NO3-. However, it is sufficient to
note that consistently high concentrations of NO3- are observed in parts of the shallow and deep
aquifer during different sampling periods. Therefore, there is evidence of N leaching to the
groundwater system. It is beyond the scope of this study to carry out a detailed mass balance for
the inputs and outputs of N within and between aquifers and hence to resolve a component
(aquifer storage) of the nitrogen budget framework described in Chapter 2 (Figure 2-5).
However, in light of the recharge-discharge characteristics established in Chapter 4 (Section
4.5) and the vertical and spatial relationships established in Sections 5.3 and 5.4, the following
analysis considers the implications of the hydrogeological characterisation for the transport of N
to groundwaters and potentially to surface waters. Figure 5-33 summarises these concepts.
Visual comparisons between Figure 5-29, Figure 5-30 and Figure 5-32 form the basis of the
following interpretation.
5.5.3.1 Shallow aquifer
It was established in Chapter 4 that recharge to the shallow aquifer is from both lateral and
diffuse sources (Figure 4-17b). Therefore, given the widespread fertiliser source of DIN in the
landscape, the main hydrogeologic control on the distribution of N in HSs is the aquifer
composition and hence redox condition. As discussed above (Section 5.6.2), oxygenated
groundwaters associated with sandy aquifer material result in the observed high NO3-
concentrations, while reducing groundwaters (associated with a higher clay content) result in
low DIN, with NH4+ as the dominant species. Conversion of NO3
- to N2 by denitrification may
also occur where reducing conditions are encountered. Due to the upwards head gradient and
strong vertical connection between HSs and HSd at the coastal bores, the distribution of N in the
shallow aquifer is also influenced by the composition of the deep aquifer. For example, given
that HSd discharges vertically to HSs towards the coast (Section 5.3.4), a component of DIN in
the shallow aquifer may also be from the deep aquifer. Of direct relevance to this study is that
high NO3- groundwaters are found adjacent to the Herbert River: the location of these high DIN
waters has potential implications for the transport of nitrogen to and from surface waters, as
discussed further in Chapters 6 and 7. It was also established in Chapter 4 that in addition to
Hydrogeochemical Framework
145
discharge to the Herbert River, discharge from the shallow aquifer occurs out to sea. Therefore,
based on the available measurements at the coastal bores, there is the potential for relatively low
concentrations of DIN (present as NH4+) to contribute directly from the shallow aquifer to the
offshore marine environment.
5.5.3.2 Deep aquifer
The highest concentrations of DIN are found in the northwest of the catchment, which coincides
with the main recharge area for the deep aquifer. Given that lateral recharge is the dominant
supply of water to HSd, N sourced from this area is a major contributor to the deep aquifer.
Although the shallow aquifer has a high concentration of NO3-, and strong/good vertical
connectivity between HSs and HSd is maintained away from the recharge zone, there is a
dramatic decrease in DIN in the deep aquifer along the flowpath. Therefore, despite there being
multiple sources of DIN (lateral and vertical) and appropriate hydraulics to enable leaching to
occur, the observed concentrations in the deep aquifer are indicative of denitrification arising
from more reducing conditions. Further measurements would be required to verify whether the
decrease in DIN is matched by an increase in the concentration of N2 (Bohlke and Denver,
1995). Alternatively, it is plausible that DIN concentrations are lower in the deep aquifer
compared to the shallow due to HSd groundwaters being older and recharged prior to more
recent anthropogenic inputs.
The potential contribution of DIN from groundwater to surface waters is of particular interest to
this study. Whilst river-aquifer relationships in relation to N are explicitly examined in Chapters
6 and 7, it is noted that high NO3- groundwaters are observed at the beginning of a preferential
pathway that runs parallel to the lower Herbert River. In addition, it was proposed in Chapter 4
that the potential for groundwater discharge from the deep aquifer to the river exists in the upper
reaches of the valley. Therefore, a contribution of DIN from deep groundwater to the river is
plausible. Furthermore, given that a component of lateral discharge from the deep aquifer is out
to sea (Section 4.5.1), there exists an additional pathway for the transport of DIN to surface
waters. In addition to diffuse groundwater discharge to the ocean, point discharge of
groundwater from confined aquifer systems, several kilometres offshore, has been described
within the Great Barrier Reef region (Stieglitz and Ridd, 2000). It has been suggested that a
small net flux of submarine groundwater discharge can deliver a comparatively large flux of
nutrients to sea. Therefore, diffuse and/or point discharge mechanisms may have ecological
significance for the delivery of nutrients to the intertidal zone or inner/mid shelf of the GBR
(Stieglitz, 2005).
Figure 5-33 Conceptual diagram summarising the movement of water and N in the alluvial aquifer system and potentially to the Herbert River (HR)
Hydrogeochemical framework
147
5.6 CHAPTER SUMMARY
A range of hydrogeochemical data have been analysed in this chapter in order to extend the
conceptual model for the hydrogeology in the alluvial aquifer system of the lower Herbert River
catchment. Specifically, hydrochemical analyses have been used to verify key attributes and
processes relating to the distinction of the main aquifers; the degree of interaction between
aquifers; and the spatial relationships within the aquifers. Based on this enhanced understanding
of the hydrogeological framework, the distribution of N in groundwaters of the two aquifers and
the potential implications for contributions to surface waters have also been assessed.
HSs groundwaters in the upper and middle sections of the study area are characterised by Na-
HCO3-Cl facies, with Ca-Mg enriched and Na enriched groups related to clay content and hence
aquifer lithology. Whilst sandy units within the shallow aquifer are chemically distinct, they
nonetheless share a common recharge source and are considered to represent the same aquifer at
the scale of interest. The dominance of HCO3 is consistent with recharge from proximity to a
rainfall-recharge zone: recharge is rapid and, in general, not associated with evaporation prior to
recharge. Na-Cl ± HCO3 facies groundwaters are found in the lower (eastern) section of the
catchment, with high and low salinity groups observed. The highest salinities are associated
with contributions from the deep aquifer, either from upward vertical recharge at the coast or
contamination during pumping. Lower salinity trends are indicative of minor hydrochemical
evolution in some areas and mixing with water derived from the Herbert River in the tidal zone.
The absence of strong evidence for hydrochemical evolution is consistent with relatively short
groundwater residence times and therefore less opportunity for water-rock interactions.
The spatial separation of water types in the deep aquifer is more distinct than in the shallow
aquifer. The main recharge area in the northwest is characterised by Na-HCO3 facies
groundwaters, whilst the central part of the study area is dominated by Na-HCO3-Cl facies. The
lower section of the catchment is represented by Na-Cl facies waters with high salinities. Whilst
there is evidence of hydrochemical evolution in HSd, lateral development is interrupted by
enhanced vertical leakage from the shallow aquifer and contributions of higher salinity waters
from the Stone River valley. Therefore, whilst there is overall a general increase in solute
concentrations and TDS down-flow of the main recharge area, the degree of hydrochemical
evolution is suppressed because of the vertical connectivity between aquifers and the
convergence of flowpaths contributing groundwater of different compositions. In the eastern
part of the catchment, the hydrochemistry of the deep aquifer is influenced by seawater from
past and/or present-day intrusion. A preferential pathway for intrusion in the northeast, rather
than the eastern coastline is suggested. Hydrochemical evidence supports local vertical
groundwater discharge at the coast, where seawater is effectively forced back by freshwater.
Chapter 5
148
The analyses have highlighted the role of hydrochemistry to infer relative degrees of inter-
aquifer mixing. While bore hydrographs indicate the direction of vertical head gradient between
aquifers and the potential degree of connectivity, the hydrochemistry provides evidence of
actual exchange of water and the relative extent. On balance, the hydrochemistry supports the
conceptualisation of the subsurface into a two-aquifer system, with there being spatial
variability in hydraulic connectivity between aquifers. Geochemical modelling, which is beyond
the scope of this research, would allow the degree of inter-aquifer mixing to be quantified.
Connectivity relationships between the aquifers are important for understanding the partitioning
of N in groundwater. In general, concentrations of DIN are greatest in the shallow aquifer, with
the distribution and speciation influenced by aquifer composition, including soil type and redox
state. Nitrification, associated with sandy soils and oxidising conditions, produces NO3- that is
readily leached to HSs. In contrast, clay soils, prone to water-logging, result in denitrification
and hence low DIN groundwaters. High DIN in HSd is restricted to the main recharge area in the
northwest, which is the main source area for N leached to the deep aquifer. Despite good
vertical connectivity between the aquifers in some areas, there is strong partitioning of DIN due
to the redox control on the fate of N in groundwater. Thus, whilst oxidising conditions in the
main recharge area favour NO3-, a dramatic decline in DIN is evident away from the recharge
zone due to N reduction processes. Mass balance calculations, including measurements of other
forms of N, would be required to verify the proposed mechanisms for the observed speciation of
DIN in both aquifers. The application of nitrogen isotopes is also an area for future research.
Whilst it is beyond the scope of the thesis to undertake a detailed mass balance for the inputs
and outputs of DIN within and between aquifers, the hydrochemical analyses provide evidence
of N leaching to the groundwater system. Furthermore, the spatial distribution of DIN in each
aquifer allows the potential risk of N transport from groundwater to surface waters to be
considered. In particular, the observed high concentration of DIN in both aquifers in locations
adjacent to the Herbert River indicates the potential for N in groundwater to contribute to the
river system or vice versa. Given the ultimate discharge of groundwater from both aquifers to
the sea, there is an additional pathway for the movement of N offshore. The following two
chapters explicitly examine the connectivity between groundwater and the Herbert River, from
both a physical (Chapter 6) and chemical (Chapter 7) perspective. Therefore, while this chapter
has highlighted areas of high DIN in the alluvial aquifers, subsequent chapters examine (i)
whether exchange of water from groundwater to surface water is a plausible mechanism and (ii)
whether there is hydrochemical evidence of N transport to the Herbert River via this
mechanism.
149
Chapter 6 Physical River-Groundwater Interactions
6.1 INTRODUCTION
A framework for subsurface water movement in the lower Herbert River catchment was
developed in Chapters 4 and 5 through interpretation of physical and chemical hydrogeological
datasets. The following two chapters incorporate surface water into the conceptual model by
considering river-groundwater interactions. Resource and Environmental Management (REM,
2002) devised a classification system for defining the nature of stream-aquifer interactions in
order to assist with targeting management in areas where these interactions are important.
Although the classification system was originally proposed for use in the Murray-Darling Basin,
where the issues are largely concerned with managing conjunctive water use, the principles are
also relevant to the current study, in which the nature of these interactions can have important
water quality implications (refer to introductory chapters 1-3).
Table 6–1 Classification system for stream-aquifer interactions relevant to conjunctive use management (REM, 2002).
(1) Hydraulic connection (2) Stream-aquifer interaction process
(3) Potential impacts of poor quality groundwater on
surface water quality
connected gaining stream high
connected losing stream no impact
connected variable gaining/losing stream low
disconnected losing stream no impact
As shown in Table 6-1, the three levels of classification are aimed at distinguishing whether (1)
there is hydraulic connection; (2) the direction of interaction; and (3) the likely impact of
groundwater on stream water quality. This chapter specifically addresses (1) through
comparisons of groundwater elevation to the topography (bed and floodplain) of the Herbert
River, while (2) is assessed through groundwater elevation-river stage relationships. These
techniques represent qualitative hydrometric approaches (Section 2.4.1.2). In addition, various
hydrological methods (e.g. Section 2.4.1.3) are applied to temporal streamflow data to provide
further insight into the direction of flux between the aquifers and the river. Whilst the aim of
Chapter 6
150
this chapter is to characterise the physical hydraulic relationships between groundwater and
surface water along the lower Herbert River, Chapter 7 considers chemical interactions and
hence addresses classification (3).
6.1.1 Data availability and preparation
Numerous gauging stations have been installed in the Herbert River catchment to collect flow
and/or water quality data. Two flow gauges are currently operational along the lower Herbert
River (Figure 6-1). Although the lower catchment is the focus of this section, measurements
from gauge 116004, immediately upstream of the Herbert River gorge, are referred to where
appropriate. The length of historical data available at each of the stream gauges is variable: a
long time series of flow and stage data (from 1915) is available at gauge 116001. Water quality
data is also available at some gauges during part of the gauging period (analysed in Chapter 7).
A summary of gauge characteristics is provided in Table 6-2.
%U
%U
%U
116004
116006
116001
stream gauges%U
lower Herbert Riverstreams
20 0 20 kmN
#
Herbert RGorge
#S
Figure 6-1 Selected QDNRW stream gauges along the lower (116006, 116001) and upper (116004) Herbert River which are referred to in the following text.
Physical River-Groundwater Interactions
151
In order to compare groundwater and river heights, the data were converted to a common
datum, in this case the Australian Height Datum (AHD). River stage data provided by QDNRW
were corrected for the gauge zero value and then a general correction factor (advised by
hydrographers at QDNRW) was applied to convert the stage height data from the State Datum
to AHD. Groundwater elevations (in m AHD) were derived by subtracting groundwater levels
(depth from reference point) from the elevation of the bore reference point (in m AHD) (as
discussed in Section 4.4.1).
Table 6–2 Features of the QDNRW stream gauging stations in the lower catchment and selected upper catchment gauges.
Stationa Streamb Start date End date Typec
116001 (A-E) L Herbert River 1/8/1915 open F, WQ
116006 (A-B) L Herbert River 2/2/1968 open F, WQ
116004 (A-C) U Herbert River 31/5/1922 open F, WQ
a Letters attached to the station number after ‘A’ correspond to minor shifts in the gauge position over time: the start date is given for the first gauge to be installed and end date for the last b L and U refer to Lower and Upper extents of the streams c F and WQ refer to Flow and Water Quality (WQ is available for at least part of the gauging period).
Surveyed river cross-sections were also provided by the Hinchinbrook Shire Council at a
number of sites along the lower Herbert River. The lowest point of the riverbed and the
maximum heights of the left and right banks (all in m AHD) were recorded from the most recent
cross-section available at each site. The distance from the river outlet of each survey location
was also determined from a GIS. As can be observed in Figure 6-2a, there is a close distribution
of surveyed cross-sections for approximately 60 km upstream from the river mouth. Within this
stretch of river there is a stream gauge (116001) at around 22 km upstream; at around 70 km
from the river mouth there is an additional stream gauge (116006). Both of these gauges have
stage height records in metres relative to AHD. In the absence of extensive gauge data along the
entire length of the lower Herbert River, stage height is calculated at locations away from the
gauges by adding the measured water column at the sampling sites to the riverbed depth. Note
that measurement of the water column allows for changes in river stage due to variation in the
width of the river along its length but not to variations in height with discharge (Figure 6-2b).
Where sampling sites and cross-section locations are not coincident, the riverbed depth is
derived from the average between adjacent (upstream and downstream) surveyed cross-sections.
Table 6-3 shows that there is reasonably good agreement (within 1 m) between derived and
gauged stage heights at the two QDNRW stream gauges on the same day of measurement. This
provides confidence in using derived river stage heights (where necessary) to compare with
groundwater elevations, particularly during the dry season. This approach is examined in
Section 6.3.
Chapter 6
152
-10
0
10
20
30
0 10 20 30 40 50 60 70 80 90 100Distance from river mouth (km)
Sur
veye
d riv
erbe
d (m
AH
D)
a
0
50
100
150
200
250
0 10 20 30 40 50 60 70 80 90 100Distance from river mouth (km)
Rive
r w
idth
(m)
116006116001
b
Figure 6-2 (a) Surveyed riverbed and (b) estimated river width (June 2005) as a function of distance from the mouth of the Herbert River. The location of the two stream gauges along the river is also indicated. Source: Hinchinbrook Shire Council (surveyed riverbed).
Table 6–3 Comparison between gauged (QDNRW) and derived stage heights on the same day of measurement during the dry season.
Date Gauge Gauged (m AHD)
Derived (m AHD)
Difference (m AHD)
26/10/2004 116001 1.4 0.6 0.8
3/06/2005 116001 1.6 1.8 -0.2
2/11/2004 116006 16.6 16.0 0.6
8/06/2005 116006 17.0 16.1 0.9
Physical River-Groundwater Interactions
153
Time series stage heights are derived at locations away from the gauges by adjusting the gauged
data by the difference between the gauged and derived stage (away from the gauge) at the same
period in time. As water column measurements were recorded during two dry season months
(early and late in the season), an average value is used. For example, the surveyed depth of the
riverbed at Trebonne (located approximately 7 km upstream of gauge 116001) is 2.7 m AHD,
while the measured water column in October 2004 (late dry season) was 0.8 m. Therefore, stage
height at Trebonne is estimated as the summation i.e. 3.5 m AHD. The recorded stage height at
the gauge on the same day was 1.36 m AHD; hence, the difference between the gauged and
calculated stage is approximately 2 m. A similar difference is obtained in June 2005 (early dry
season). This translates to a 2 m vertical shift upwards of historical stage heights at gauge
116001 to derive the time series at Trebonne (Figure 6-3). Note that this approach assumes that
a constant factor can be applied to the gauge data for the entire record based on shifts in stage
heights at Trebonne during the dry season. While this assumption is plausible during the dry
season due to the low variability in river stage, there is greater uncertainty in the derived stage
heights during high flow periods.
0
5
10
15
20
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
Stage @gauge116001
Stage @Trebonne
HR @Trebonne
Riverbed @Trebonne
2 m vertical shift
Figure 6-3 Derived historical stage height (m AHD) in the Herbert River at Trebonne. Circles represent the calculated river stage in October 2004 and June 2005 based on surveyed riverbed and measured water column depth, which were used in conjunction with gauge 116001 values to derive the river stage time series at Trebonne.
Chapter 6
154
6.2 POTENTIAL FOR HYDRAULIC CONNECTION
As outlined in Chapter 2 (Section 2.2.1), a connected river-aquifer system arises where there is
direct contact between a stream reach and an underlying aquifer via a zone of saturated material
or a narrow unsaturated zone (Bouwer and Maddock, 1997). The potential for hydraulic
connection was discussed briefly in Chapter 4 (Section 4.2.2.1) based on the construction of
lithologic cross-sections and river profiles, which illustrated that in the upper part of the valley
the Herbert River intersects the sediment comprising the deep aquifer, while further downvalley
the river incises the shallow aquifer only (Figure 4-4). Hydraulic connection is further assessed
in this section by comparing groundwater elevations with stream topography: connection is
assumed to potentially exist where the groundwater elevation lies within the elevation of the
channel, defined as between the bed and bank of the river. Given the availability of surveyed
river profiles and the relatively small number of bores along the Herbert River, site-based
comparisons are considered to be feasible. In order to simplify the analysis, the lower Herbert
River is divided into four distinct reaches: (A) from the river mouth to gauge 116001; (B) from
gauge 116001 upstream to the junction with the Stone River; (C) upstream of the Stone River
junction to Long Pocket; and (D) from Long Pocket to Abergowrie, towards the western extent
of the lower catchment (Figure 6-4 and Figure 6-5). These reaches are chosen because of the
locations of the stream gauges along the lower Herbert River and consideration of the tidal zone
(downstream of gauge 116001) and bore distribution. Bores selected for the analysis are located
within 3 km of the Herbert River (at the closest point) and correspond to one of the four river
reaches (Figure 6-5).
-10
0
10
20
30
40
50
0 10 20 30 40 50 60 70 80 90
Distance from river mouth (km)
m A
HD
River bed
Max elevation left bank
Max elevation right bank
Figure 6-4 Surveyed topographic features of the lower Herbert River channel in m AHD. Vertical lines mark the division of the lower Herbert River into the four reaches (labelled). Note that left and right banks are defined from the perspective of looking downstream.
A
B
C
D
Physical River-Groundwater Interactions
155
%U
%U
%U
%U
%U%U
%U
%U
#
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#S
#S#S
#S
#S#S
#S
#S
#S
#S
#S
#S
#S
#S#S
%U
#S#S
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#S
#S#S #S #S
#S
#S
#S
#S
#S
#S
#S
#S
$T$T$T$T
$T$T
$T
$T$T
$T
$T
$T$T$T
$T
$T$T$T$T
$T
$T$T
$T
Halifax
Trebonne
Lannercost
#
Abergowrie bridge
Timrith
Ingham
#
Gairloch bridge
99025 99047
36
48
51
52
59
61
62
67
68
69
7071
%U
McNade
58
53
49
#
Gedges Xing
River mouth
Abergowrie
LongpocketD
C
B
A
Stone Riverjunction 116001
116006
Stream gauges%UWater column#S
River x-sections$TBores#S
StreamsPlaces%U
5 0 5 KmN
Ripple Ck
Ston
e R
Herbert R
Figure 6-5 Map showing geographical features which relate to the analyses of groundwater elevations and river topography/stage along the lower Herbert River. Bores are labelled without their 11600 prefix. Also indicated are locations of water column measurements and where river cross-sections are available. The four river reaches are defined as: (A) river mouth to gauge 116001; (B) gauge 116001 to Stone River junction; (C) Stone River junction to Long Pocket; and (D) Long Pocket to Abergowrie. Note that reach A is tidal.
An example of the approach used to compare historical groundwater elevations to the bed and
bank height of the Herbert River is shown in Figure 6-6 for bore 11600048. A similar approach
is applied to each of the bores within the four reaches; where the bore and cross-section
locations do not coincide, an average of bank heights (on the relevant side of the river) and
riverbed measurements from the nearest upstream and downstream cross-sections are used. Note
that the relevant bank is that which is on the same side of the river as the bore, where the bank is
defined from the perspective of looking downstream. Thus, for bore 11600048, the height of the
right bank is applicable, while for bore 11600061, the left bank is appropriate for comparison
with groundwater elevation (Figure 6-5). The analyses are performed on both shallow and deep
groundwaters because of the vertical connectivity between HSs and HSd in some parts of the
catchment and incision of the river channel into both aquifers (Figure 4-4, Chapter 4).
The analyses indicate that throughout the historical record, groundwater elevations for the
shallow and deep aquifer are within the corresponding elevation of the river channel within each
of the four reaches i.e. above the elevation of the base of the stream and below the level of the
stream bank at each comparison site. Based on the available information this indicates that there
is potential for stream-aquifer connectivity along the entire length of the lower Herbert River in
both wet and dry season months. According to the framework for classifying stream-aquifer
interactions (Table 6-1) the next step is to establish the potential direction of flux.
Chapter 6
156
-5
0
5
10
1976
1978
1980
1982
1984
1986
1988
1990
1992
1994
1996
1998
2000
2002
2004
m A
HD
11600048B
Riverbed
Riverbank
Stre
am c
hann
el
Figure 6-6 Comparison of time series groundwater elevation at bore 1160048 (reach A) with the corresponding surveyed streambed and bank height in the Herbert River.
6.3 DIRECTION OF INTERACTION
Having established the potential for river-aquifer connectivity, the direction of flux between the
aquifer and the stream remains to be ascertained in order to characterise the interaction process
(Table 6-1). In Chapter 4, the shape and upstream orientation of groundwater elevation contours
where they cross the river suggested a potential gaining stream system over a large section of
the lower Herbert River during selected wet and dry periods (Figure 4-14 and Figure 4-16).
Detailed comparisons of groundwater elevation and river stage throughout the historical record
provide further insight into stream-aquifer relationships. Whilst height comparisons provide an
indication of the potential direction of interaction at a small scale, examination of flow
characteristics provides evidence of actual volumetric flux over a broad area. Therefore,
hydrological approaches such as streamflow analysis, flow duration curves and hydrograph
separation are also examined.
6.3.1 Groundwater elevation - river stage relationships
If the groundwater elevation is above the elevation of the stream, groundwater potentially
moves into the stream, which is termed gaining; the converse gives rise to a losing stream
situation (Section 2.2.1, Chapter 2). This is the underlying principle behind the following
analyses. Although the catchment has only two stream gauges, river stage is estimated at
intervening sites with reference to the measured water column and surveyed depths to the
riverbed (as discussed in Section 6.1.1). To aid with the analysis, the river is divided into the
four reaches defined in Figure 6-4 and Figure 6-5. A list of bores and the relevant comparison
Physical River-Groundwater Interactions
157
stream gauge at the closest sites along each reach is provided in Table 6-4 (also refer to Figure
6-5). An explanation of features specific to the analyses along each reach is provided below,
with examples illustrating the approach used to compare groundwater elevations and river stage.
Table 6–4 Bores and comparison stream gauges at sites along the four river reaches defined as: (A) river mouth to gauge 116001; (B) gauge 116001 to Stone River junction; (C) Stone River junction to Long Pocket; and (D) Long Pocket to Abergowrie (refer to Figure 6-5).
Reach Bore number1 Stream distance2 (km)
Comparison stream gauge
A
101A, 101C3
48A, 48B, 49A, 49B
36A, 36B
52A, 53A, 53B
0.82
0.15; 2.6
1.8
0.73; 2.7
adjusted4 gauge 116001 at Ripple Ck
adjusted gauge 116001 at McNade
adjusted gauge 116001 at Halifax
adjusted gauge 116001 at Gairloch
B
61A, 61B, 62A, 62B
99025, 99047
59A, 58A
0.97; 1.7
0.65; 0.45
0.42; 2.2
adjusted gauge 116001 at Trebonne
adjusted gauge 116001 at Trebonne
adjusted gauge 116001 at Gedges Crossing
C
67A
68A, 68B
0.18
0.65
adjusted gauge 116001 & 116006 at Timrith
adjusted gauge 116001 & 116006 at Lannercost
D
69A
70A
71A
2.1
0.81
1.7
adjusted gauge 116006 at Abergowrie Bridge
gauge 116006
adjusted gauge 116006 at Abergowrie 1 Bores are denoted without their 116000 prefix 2 Closest distance to the Herbert River
3 Bore 101 is the replacement for 51: 51A = 101A and 51B = 101C 4 River stage has been adjusted from the gauge value based on the measured water column and surveyed depths to the riverbed (refer to Section 6.1.1).
6.3.1.1 Reach A: river mouth to gauge 116001
The river reach downstream of gauge 116001 is tidal and therefore stage heights measured at
the gauge do not necessarily represent stage heights further downstream. However, in the
absence of detailed tidal information, river stage has been derived (from the measured water
column, riverbed depth and gauge data) at selected locations along the reach and compared with
groundwater elevations (Figure 6-7). The analyses indicate that all bores screened in the shallow
and deep aquifers have groundwater elevations above the corresponding river stage
measurements during the dry season. In addition, the analyses highlight that the river-
groundwater relationship generally reverses during streamflow events (peaks) in the wet season;
however, as groundwater heads (for both aquifers) increase further upstream along the reach,
such as at bores 11600052 and 11600053, this reversal is less frequent.
Chapter 6
158
-5
0
5
10
15
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
HR @ CSR
HR @ Gairloch
Stage @ CSR
Stage @Gairloch
11600048B
11600053B
a
-5
0
5
10
15
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
HR @ CSR
HR @ Gairloch
Stage @ CSR
Stage @Gairloch
11600048A
11600052A
b
Figure 6-7 Comparison of historical groundwater elevations in selected (a) shallow bores and (b) deep bores with corresponding river stage (adjusted) along reach A of the Herbert River. Circles represent calculated river stage based on the riverbed elevation and measured water column. Bore 11600048 is compared with river stage at CSR (blue), while bores 11600052 and 11600053 are compared with river stage at Gairloch (orange).
Physical River-Groundwater Interactions
159
6.3.1.2 Reach B: gauge 116001 to Stone River junction
Comparison between historical groundwater elevations and derived stage heights for HSs and
HSd bores along reach B indicates that river stage remains below the groundwater elevation in
both aquifers during the dry season (Figure 6-8). However, in the wet season river stage
sometimes exceeds the groundwater elevation, especially at bores on the left bank of the river
(e.g. bores 11600061, 99047 and 11600059) which tend to have a lower groundwater elevation
than on the right bank (e.g. bores 99025 and 11600062).
0
5
10
15
20
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
HR @Trebonne
Stage @Trebonne
11600061B
11600062B
99025
99047
a
0
5
10
15
20
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
HR @ Gedges Xing
HR @Trebonne
Stage @Gedges Xing
Stage @Trebonne
11600061A
11600062A
11600059
b
Figure 6-8 Comparison of historical groundwater elevations in selected (a) shallow bores and (b) deep bores with corresponding river stage heights (adjusted) along reach B of the Herbert River. Circles represent calculated river stage based on the riverbed elevation and measured water column. Bores 11600061, 11600062, 99025 and 99047 are compared with river stage at Trebonne (orange), while bore 11600059 is compared with river stage at Gedges Crossing (blue). Note that bores 99025 and 99047 have only partial records.
Chapter 6
160
6.3.1.3 Reach C: upstream of Stone River junction to Long Pocket
There are only two monitoring bores along reach C. Given that there are no stream gauges along
the reach, historical stage height data at gauges 116001 and 116006 have been adjusted based on
field measurements (Section 6.1.1). Table 6-5 shows that in the Herbert River at Timrith, the
derived stage height is 1.4 m above gauge 116001 and 13.9 m below gauge 116006 during both
the end (October 2004) and beginning (June 2005) of the subsequent dry season. In the Herbert
River at Lannercost, the measured stage height is 4.1 m above gauge 116001 and 11.2 m below
gauge 116006 during both dry season measurements. Note that both stream gauges give similar
stage height time series when adjusted for measurements at Timrith or Lannercost.
Table 6–5 Comparison of derived river stage at Timrith and Lannercost and actual stage heights at gauging stations 116001 and 116006 in the lower Herbert River.
Location Date
Derived river stage (m AHD)
Stage height gauge 116001 (m AHD)
Stage height gauge 116006 (m AHD)
Timrith 28/10/04 2.8 1.4 16.7
5/06/05 3.0 1.6 16.9
Lannercost 27/10/04 5.5 1.4 16.7
1/06/05 5.7 1.6 16.9
0
5
10
15
20
25
30
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
HR @Lannercost
Stage @Lannercost(116001)
Stage @Lannercost(116006)
11600068A
11600068B
Figure 6-9 Comparison of groundwater elevation in shallow (11600068B) and deep (11600068A) bores with river stage (adjusted) at Lannercost. Similar stage heights are attained when adjusted using the records at gauge 116001 or 116006. Circles represent calculated river stage based on the riverbed elevation and measured water column.
Physical River-Groundwater Interactions
161
Comparison of groundwater and adjusted stream elevations indicate that for the deep aquifer
(e.g. bore 11600068A) the groundwater elevation is above river stage during the dry season and
generally less during the wet season. However, for the shallow aquifer (e.g. bore 11600068B)
the groundwater elevation generally exceeds river stage throughout both wet and dry season
months (Figure 6-9).
6.3.1.4 Reach D: Long Pocket to Abergowrie
Stream gauge 116006 is located on the most upstream reach of the lower Herbert River. As for
the other river reaches, stage height has been adjusted at locations distant from the gauge (close
to the monitoring bores) based on derived stage heights (Table 6-4). As there was no river depth
measurement available for the Herbert River at Abergowrie, a conservative estimate of stage
height was made from a surveyed cross section of the riverbed (assuming a zero water column
as a minimum in the dry season). Stage height and groundwater elevation comparisons indicate
that the elevation of groundwater in the deep aquifer (e.g. bore 11600070) is generally above
river stage throughout the year, except during some wet season events when there is a large rise
in river stage (Figure 6-10).
15
20
25
30
35
1975
1977
1979
1981
1983
1985
1987
1989
1991
1993
1995
1997
1999
2001
2003
2005
m A
HD
Stage @116006
11600070
Figure 6-10 Comparison of groundwater elevation at bore 11600070 and stage height at adjacent gauge 116006.
Chapter 6
162
6.3.1.5 Interpretation of flux direction
The above analyses indicate that the direction of flux between the Herbert River and the two
alluvial aquifers varies both seasonally and spatially. Based on the available information, the
relationships of the four reaches of the Herbert River with groundwater are discussed below.
As summarised in Table 6-6, river reaches in the floodplain area (reaches A and B) display a
seasonal pattern: the dominant direction of potential flux is towards the river during the dry
season and from the stream to the aquifers during high flow periods in the wet season. In the dry
season, stream-aquifer relationships along reach C display a similar trend to the downstream
reaches. However, while the direction of flux reverses with respect to HSd in periods of the wet
season, the elevation of groundwater in HSs is maintained above stage height. The potential
flow direction is thus from the shallow aquifer to the stream throughout the year. Similarly, the
high elevation of groundwater in reach D means that groundwater levels are generally above
river stage throughout the year: the dominant direction of potential flux is from the deep aquifer
to the river.
Table 6–6 Summary of the dominant river-groundwater elevation relationships along the lower Herbert River during wet and dry seasons in the historical record.
Reach Season Aquifer(s) River-groundwater elevation relationship
Theoretical direction of flow
A dry shallow & deep gw elevation > stage gw → river
wet shallow & deep stage > gw elevation+ gw ← river
B dry shallow & deep gw elevation > stage gw → river
wet shallow & deep stage > gw elevation+ gw ← river
C dry shallow & deep gw elevation > stage gw → river
wet shallow gw elevation > stage+ gw → river
wet deep stage > gw elevation+ gw ← river
D dry deep gw elevation > stage gw → river
wet deep gw elevation > stage+ gw → river
+ This relationship is observed during the majority of wet season periods in the historical record.
Physical River-Groundwater Interactions
163
On balance, the analyses indicate that the Herbert River can be characterised as a dominantly
gaining stream with respect to the shallow and deep aquifers, consistent with the water level
contours constructed in Chapter 4 (Figure 4-14 and Figure 4-16). However, during streamflow
events in the wet season the lower reaches (A and B) of the river potentially represent a losing
system, noting that the direction of flux switches from the aquifers back to the stream as river
stage relaxes. While a reversal in flow direction is consistently observed at some locations in the
deep and shallow aquifers, at other locations this relationship depends on the extent of river
stage rise during a particular streamflow event. As there is greatest uncertainty in the river stage
data during high flow events, there is less confidence in the river-aquifer relationships
established during the wet season. However, assuming that the relative error between
groundwater elevations and river stage is small, in general there is a transition from a
dominantly gaining system to variably gaining-losing system from the upper to lower reaches of
the lower Herbert River.
An important observation from the analysis of paired stream-bore hydrographs is that even
though the direction of stream-aquifer flux is towards the river during the dry season,
groundwater elevation is always maintained above minimum river stage. This suggests that
while there is evidence of potential leakage from the aquifer to the stream, there is a constriction
in flow such that the groundwater head remains elevated. This is particularly evident in the
upper reaches of the catchment where there is up to a 10 m difference between the elevation of
the stream and shallow groundwater at a distance of approximately 800 m from the river (e.g.
bore 11600068B in Figure 6-9). It is beyond the scope of the thesis to examine in detail the
reasons for this observation; however, it was noted in Chapter 4 that mud units present
throughout the lithologic sequence may act as semi-confining/confining layers and hence retard
flow either vertically or laterally. Therefore, the hydraulic properties of the aquifers or
surrounding material may provide an explanation. Whilst a zero point error in either the
groundwater elevations or river stage is a possibility, a difference of up to 10 m seems unlikely.
The approach taken in this section has enabled the potential direction of flux between the lower
Herbert River and the adjacent aquifers to be determined. However, the potential direction of
flux does not necessarily equate to actual movement of water between the aquifers and the
stream, as this will depend on factors such as aquifer hydraulic properties and the infiltration
characteristics of the soils and riverbed sediment. In addition, the approach has assumed that
stage height and the elevation of the riverbed, bank and groundwater have been accurately
determined. Therefore, the following section considers the characteristics of streamflow to
establish whether there are volumetric gains and/or losses to the Herbert River that can be
attributed to stream-aquifer fluxes.
Chapter 6
164
6.3.2 Flow characteristics
There are a number of standard hydrological techniques used to study the flow characteristics of
rivers (Section 2.4.1.3). Several of these methods are applied in the following sections based on
the availability of historical streamflow data. Given that this thesis is primarily concerned with
interactions during the dry season, the low-flow regime of the river is the focus of the analyses.
6.3.2.1 Stream hydrographs
A stream hydrograph shows the fluctuations in stream discharge through time and can be used
to gain insight into the relationships between rainfall and runoff in a catchment as well as the
role of baseflow. As depicted in Figure 6-11a, streamflow in the river displays a distinct
seasonal pattern. In the wettest months, generally from December - March, flow in the river is at
least 200 GL/day and greater than 900 GL during large flood events in the historical record. In
comparison, dry season flows are generally less than 50 GL/day in April/May and decline to
less than 200 ML/day by October/November, before the next wet season commences. Although
streamflow declines by the end of the dry season, flow remains above zero despite little or no
rainfall (Figure 6-11b). In the absence of lakes (or glaciers) in the area this indicates that
releases from groundwater storage (± bank storage) must sustain streamflow in the lower
Herbert River during low-flow periods. Given the topographic relationships between the
streambed and aquifers and the relative river-groundwater elevations (Sections 6.2 and 6.3.1) it
is considered that a ‘true’ groundwater body is drained rather than just near-channel/bank
storages (Smakhtin, 2001).
In contrast to the lower reaches of the Herbert River, dry season flows at gauge 116004 in the
upper catchment (Figure 6-1) are commonly less than 30 ML/day and zero-flow has been
recorded several times in the historical record. This suggests that unlike in the lower catchment,
groundwater does not always sustain flow in the upper Herbert River during the dry season.
Physical River-Groundwater Interactions
165
0
100
200
300
400
500
600
700
800
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
Flow
(GL/
day)
Flow @ 116001Flow @ 116006
a
1
10
100
1000
10000
100000
1000000
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
Flow
(ML/
day)
0
100
200
300
400
Rai
nfal
l (m
m)
Flow @ 116001Flow @ 116006Rainfall
b
Figure 6-11 Daily flow at the two lower Herbert River stream gauges during 1995-2005. Source: QDNRW
Chapter 6
166
6.3.2.2 Flow duration
In addition to hydrograph analysis of streamflow, the characteristics of flow can be further
assessed by constructing flow duration curves. These plots show the percentage of time during
which the flow of a stream is equal to or greater than any specified discharge, regardless of
chronological order. If a stream is dominated by baseflow, the flow duration curve (FDC) is
characterised by a low slope, which means that for most of the year the stream has a fairly
uniform discharge rate. Conversely, if the FDC is steep, discharge is more variable throughout
the year and the stream is likely to be dominated by direct runoff: these streams are also likely
to show long periods of no flow. Therefore, unregulated streams that exhibit a low slope on the
FDC, and/or do not show a cease-to-flow point, indicate potential for strong linkages to
groundwater systems (REM, 2002).
Figure 6-12 illustrates the FDCs for gauges in the lower catchment as well as for one in the
upper catchment based on daily historical streamflow data. Comparison of the three FDCs
indicates that the probability of attaining or exceeding a particular discharge increases from
upstream to downstream (Figure 6-12a). A log-normal representation of the FDCs delineates the
low-flow and high-flow ends of the curves: Figure 6-12b clearly illustrates the different
characteristics of flow in the upper and lower catchments, particularly at low flow. Whilst the
Herbert River in the lower catchment flows throughout the historical record and can be
considered to be perennial, flow in the upper catchment is not always sustained. This is
consistent with a source of baseflow to the stream below the gorge, which divides the upper and
lower catchments.
In order for low-flows in the river to be maintained by groundwater storages, (i) the draining
aquifer must be recharged seasonally with sufficient volumes of water; (ii) the watertable must
be shallow enough to be intersected by the stream; and (iii) the aquifer’s size and hydraulic
properties must be sufficient to maintain flows throughout the dry season (Smakhtin, 2001).
While the lower catchment comprises alluvial aquifers, the majority of aquifers in the upper
catchment are in fractured rock (Johnson and Murray, 1997). Therefore, the difference in flow
characteristics between the upper and lower catchments could be due to differences in aquifer
properties as well as in the other factors listed above.
Physical River-Groundwater Interactions
167
0.01
0.1
1
10
100
1000
10000
100000
1000000
0 10 20 30 40 50 60 70 80 90 100
Probability (%)
Stre
amflo
w (M
L/da
y)gauge 116001gauge 116006gauge 116004
a
0.01
0.1
1
10
100
1000
10000
100000
1000000
-4 -3 -2 -1 0 1 2 3 4
Normal variate
Stre
amflo
w (M
L/da
y)
gauge 116001gauge 116006gauge 116004
(0.1) (2.3) (15.9) (50.0) (84.1) (97.7) (99.9)
b
Figure 6-12 Flow duration curves for stream gauges in the lower (116001, 116006) and upper (116004) Herbert River catchment (refer to Figure 6-1 for gauge locations). Note that the FDCs are represented in (b) as a log-normal plot, with probabilities (%) shown in brackets on the x-axis.
6.3.2.3 Hydrograph separation
Baseflow is defined as that part of streamflow that is not attributable to direct runoff from
precipitation or snowmelt and is usually contributed by groundwater storage or other delayed
sources (e.g. shallow subsurface storage). In many hydrological applications it is useful to be
able to separate the baseflow, or slow flow component, from quick flow (runoff, interflow and
direct rainfall), and hence attempt to isolate the low-frequency signal of a stream. In a wet
season discharge comprises baseflow and quickflow, which represent the catchment response to
rainfall; conversely, stream discharge in a dry season is dominated by baseflow (Smakhtin,
2001). The steady decline of stream discharge during periods without rainfall is referred to as
recession, representing the gradual drainage from subsurface storages.
Chapter 6
168
In the context of this study, it would be useful to quantify the groundwater input to the lower
Herbert River, especially during the recession period when the proportion of subsurface runoff
to surface runoff is considered to be greatest. At different times along a stream hydrograph the
baseflow contribution is comprised of varying proportions of groundwater discharge and
streambank storage. Contributions to baseflow from bank storage generally decline as stream
levels fall; therefore, by the end of the dry season the baseflow contribution should be largely
composed of groundwater discharge (Ward, 1967).
Baseflow separation techniques include digital filtering, graphical methods and chemical
separation approaches (Sponberg, 2000). Although digital filtering can be a simple and robust
method for evaluating baseflow for a large range of unregulated catchments, the results are very
sensitive to the filter parameter value, which needs calibration before the results can be
considered to be numerically valid (REM, 2002; SKM, 2001). The Lyne-Hollick digital filter
(Lyne and Hollick, 1979) is a one-parameter mathematical algorithm that separates the runoff
signal, or quick response; baseflow is then calculated by subtracting the filtered quick response
from total streamflow (Sponberg, 2000). Although there are numerous documented baseflow
filters (Furey and Gupta, 2001; Chapman, 1999; Nathan and McMahon, 1990; Lyne and
Hollick, 1979), the Lyne-Hollick filter was considered to be appropriate for the purposes of this
study as it has been widely applied to daily data and has a recommended filter parameter for
daily data (Grayson et al., 1996). Hence, the filter was applied to historical streamflow data at
the two lower Herbert River gauges in order to generate baseflow time series.
Figure 6-13 shows the relationships between total (observed) streamflow at the stream gauges
and the corresponding calculated baseflow over a 1-year period. Visual inspection indicates that
the contribution of baseflow as defined by this algorithm is greater in absolute terms during the
wet season (November-April); however, the proportion of baseflow to streamflow is greater
during the dry season (May-October). Consistent with the flow duration and hydrograph
analyses above, modelled baseflow at the end of the dry season correlates with total flow i.e.
stream discharge is maintained by subsurface storages. The marked seasonality of the baseflow
signal as calculated by this methodology is clearly evident in Figure 6-14, consistent with the
seasonal trends observed from the bore hydrographs in Chapter 4 (Figure 4-6).
Physical River-Groundwater Interactions
169
0
100
200
300
400
500
600
700
800
Jan
00
Feb
00
Mar
00
Apr
00
May
00
Jun
00
Jul 0
0
Aug
00
Sep
00
Oct
00
Nov
00
Dec
00
Jan
01
Flow
(GL/
day)
116001 total flow116006 total flow
116001 baseflow116006 baseflow
a
100
1000
10000
100000
1000000
Jan
00
Feb
00
Mar
00
Apr
00
May
00
Jun
00
Jul 0
0
Aug
00
Sep
00
Oct
00
Nov
00
Dec
00
Jan
01
Flow
(ML/
day)
116001 total flow116006 total flow
116001 baseflow116006 baseflow
b
Figure 6-13 Baseflow separation using the Lyne-Hollick algorithm at the lower Herbert River stream gauges. Note that the filter was applied to the entire historical record of streamflow data but for clarity is only shown for 2000-01. Source: daily streamflow data from QDNRW.
Chapter 6
170
The baseflow index (BFI) is an important index for low-flow studies and is defined as the
volume of baseflow divided by the volume of total streamflow (Smakhtin, 2001). As shown in
Figure 6-15, there is a range in the proportion of baseflow contributed to streamflow from
approximately 10-100% on a daily basis. However, on a bimonthly basis, the contribution is
generally within 50-90% throughout the course of a year at gauge 116001. Calculation of the
seasonal BFI (volumetric ratio of historical mean baseflow to historical mean streamflow for
months within each season) at gauge 116001 expressed as a percentage, gives 34% and 67% for
the wet and dry seasons, respectively. Similar percentages of 31% and 66% for average wet and
dry season baseflow contributions are calculated at gauge 116006. Although the BFI calculated
at the two gauges along the lower Herbert River are very similar, the volume of baseflow is
between 20-100 % greater at the downstream gauge at any time (Figure 6-14). Given that
calculated baseflows represent the aggregated contributions from the entire catchment area
upstream of each gauge, these observations indicate that groundwater contributions to the river
from the additional catchment area between the two gauges are significantly large. Based on
analysis of the baseflow contribution, there is evidence that groundwater discharges to both the
upper and lower reaches of the lower Herbert River. Therefore, the large subsurface
contribution between the gauges may reflect enhanced river-aquifer connectivity due to
aquifer/soil hydraulics, recharge sources and volumes, as well as physical river-aquifer
relationships such as relative elevations of the stream-aquifer topography and water levels.
6.4 IMPLICATIONS FOR N TRANSPORT
Based on the available information, the analyses in this chapter have shown that there is
evidence of groundwater contributions to the lower Herbert River throughout the year. Whilst
there is potential flux from the river to the aquifers along some reaches during periods of high
river stage, the dominant direction of flux is from the aquifers to the river. These interaction
characteristics have important implications for the transport of N from groundwater to surface
waters. It was established in Chapter 5 that high concentrations of DIN are found in both
shallow and deep aquifers, including adjacent to the Herbert River. In HSd a hotspot for NO3- is
found in bores along reach D, while for HSs high concentrations of NO3- are observed in bores
along reaches A, B and C. Hence, given the hydraulic relationships, the potential exists for NO3-
to be contributed to the lower Herbert River from groundwater sources along the entire length of
the river. However, near-stream processes such as dentrification, resulting from contact with
reducing sediments or riparian vegetation prior to groundwater discharge, may reduce or
prevent the emergence of DIN in surface waters (Bohlke and Denver, 1995). River chemistry
data are thus evaluated in the following chapter to observe whether there is evidence of N in the
Herbert River derived from the alluvial aquifers.
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Figure 6-14 Time series of calculated baseflow at stream gauges 116001 and 116006 and observed total flow at gauge 116001 (for reference). Note the baseflow recession period during the dry season (May to October). Note the larger volume of baseflow at the downstream gauge (116001) compared to the upstream gauge (116006).
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Figure 6-15 Calculated daily baseflow as a percentage of observed streamflow at gauge 116001. The red trend represents a bimonthly moving average through the baseflow proportions. Flow at the gauge highlights the wet (November-April) and dry (May-October) seasons. The baseflow filter was applied to the entire historical record of streamflow data but for clarity is only shown for a 10-year period. Note the trend of increasing baseflow proportion from the beginning to end of the dry season, which corresponds with progressively decreasing streamflow.
Physical River-Groundwater Interactions
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6.5 CHAPTER SUMMARY
Physical relationships between the alluvial aquifers of the lower catchment and the Herbert
River have been examined in this chapter. In accordance with the classification scheme for
stream-aquifer interactions outlined in the introduction to this chapter, there is evidence of
potential for hydraulic connection along the entire lower Herbert River. The dominant direction
of potential flux is from the aquifers to the river, although during short periods of high river
stage in the wet season the direction reverses along some reaches in the floodplain area. Flow
characteristics provide an indication of actual flux between the aquifers and the stream.
Baseflow is an important source of recharge to the river, particularly during the dry season when
contributions from groundwater can be up to 100 %. While groundwater contributions to both
the upper and lower reaches of the lower Herbert River are apparent, there is an enhanced
contribution of baseflow between the two gauges in the study area (i.e. below gauge 116006) as
indicated by digital filtering. This may reflect enhanced river-aquifer connectivity due to
aquifer/soil hydraulics, recharge sources and volumes, as well as physical river-aquifer
relationships. On balance, the upper reaches of the lower Herbert River can be considered to be
a dominantly gaining system, while towards the coast the river is characterised as a variably
gaining/losing stream. Given that similar physical river-groundwater relationships have been
identified in this chapter for each aquifer, it is not possible based on the available information to
determine the relative proportion of groundwater flux from each aquifer to the river.
Hydrochemical data is thus examined in Chapter 7 to build on the conceptual model for river-
groundwater interactions. The physical relationships established in this chapter provide a
platform for the subsequent hydrochemical analysis and assessment of the potential impacts of
stream-aquifer interactions on surface water quality, particularly in relation to species of N.
Based on the spatial distribution of DIN in both shallow and deep aquifers, the stream-aquifer
hydraulics indicate that there is potential for N in groundwater to contribute to the lower Herbert
River. Evidence for this is provided in the following chapter.
175
Chapter 7 Chemical River-Groundwater Interactions
7.1 INTRODUCTION
In the previous chapter it was established that based on physical hydraulic relationships there is
evidence of stream-aquifer interaction along the lower Herbert River. In addition, it was
demonstrated that the dominant direction of interaction is from groundwater to the river,
especially during the dry season. This chapter explores the chemistry of the surface water
system in order to verify the physical river-aquifer linkages and enrich the conceptual model
that the previous three analytical chapters (4-6) have developed. Temporal river chemistry data
are analysed in order to identify hydrochemical signatures and seasonal patterns. Longitudinal
trends are also examined, from data collected for this study, in order to identify variations in
water sources to the stream during months at the beginning and end of the dry season. A variety
of dissolved constituents are examined, both conservative and non-conservative tracers,
including field parameters, major ions, and stable isotopes. Given the focus of the thesis on
river-groundwater interactions, radioactive isotopes are also analysed in order to trace
groundwater discharge and estimate the flux along the river. The synthesis of the extensive
hydrochemical database provides a powerful tool to characterise the chemistry of the river in
space and time, and for the key processes that influence its chemistry to be identified. Based on
the observed concentrations of DIN in the shallow and deep aquifers adjacent to the river, it was
proposed in Chapter 6 that there is the potential for N transport to the river from groundwater
sources. Therefore, in accordance with the objectives of the thesis, this chapter evaluates the
role of groundwater as a transport vector for dissolved inorganic forms of N to the river.
Furthermore, the environmental significance of the results for in-stream and marine ecosystems
is considered. A conceptual diagram summarising water and N movement between the aquifers
and the lower Herebert River is presented at the end of this chapter (Figure 7-38).
7.1.1 Factors that influence river chemistry
Surface water and groundwater contain dissolved solutes and gases derived through a variety of
processes within the hydrologic cycle. As discussed in Chapter 2 (Section 2.2.2), climatic,
geological and biochemical factors influence the chemistry of natural waters; mixing between
waters from different sources also impacts on stream and groundwater chemistry. Compared to
groundwater, the chemistry of a stream is generally more dynamic due to faster water
movement; rapid response to rainfall events; and potential for multiple sources which contribute
Chapter 7
176
water to the stream, including groundwater. In addition, the direct exposure of surface waters to
the atmosphere means that climatic effects such as evaporation in general have a greater
influence on surface water chemistry than subsurface waters, especially in comparison to deep
groundwaters.
Based on the compositions of the world’s major rivers, Gibbs (1970) presented a diagram which
relates total dissolved solids (TDS) and the Na/(Na+Ca) ratio to predict whether a stream
sample is dominated by: (1) the chemistry of precipitation (rain dominated); (2) rock-weathering
reactions (rock dominated); (3) evaporative concentration-crystallisation; or by a combination of
these influences (Figure 7-1). Seawater plots in the top right hand corner of the diagram, while
rainfall generally plots in the lower right hand corner at low TDS.
1
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Precipitation dominance
Evaporation/ fractional crystallisation dominance
seawater
Rock dominance
Figure 7-1 Gibbs diagram depicting the key processes that control the chemistry of surface waters. The world’s major rivers plot within the ‘boomerang’ envelope (modified after Gibbs 1970).
In addition to environmental factors are the effects of human activities. For instance,
agricultural development can impact on stream water quality through the use of fertilisers and
pesticides as well as pumping of surface and/or groundwaters for irrigation. As the case study
area is largely comprised of dryland cropping, the potential effects of irrigation will not be
discussed further. However, as the catchment is dominated by sugarcane farming, nitrogen is an
important diffuse contaminant to the surface water network (Section 3.2.1, Chapter 3).
Superimposed on the processes which affect the transport of chemical constituents to a stream
are in-stream biogeochemical reactions such as acid-base reactions; mineral precipitation and
Chemical River-Groundwater Interactions
177
dissolution; sorption and ion exchange; oxidation-reduction (redox) reactions; dissolution and
exsolution of gases; and biodegradation (Winter et al., 1998). These reactions affect the fate and
concentration of chemical species from their point of entry to further downstream.
A comprehensive discussion of the key thermodynamic principles and chemical processes that
ultimately control the composition of natural waters is provided in Hem (1985). The effects of
each of the natural processes and anthropogenic activities outlined above on the observed
hydrochemistry of a stream are summarised below.
7.1.1.1 Climatic factors
The amount and rate of rainfall, runoff and evaporation, as well as temperature, are the key
components of climate that influence surface water composition. Solutes introduced through
rainfall reflect inputs from seawater, dust, and atmospheric gases, which in turn impart a
characteristic chemistry to the rainfall that hits the land surface. The ultimate chemistry of a
stream largely depends on the pathways that this rainfall moves along and hence the time it
takes for the water to reach the stream. While overland flow can have an important effect on
stream quality following storm events, groundwater inflow can dominate the hydrochemistry
during dry periods. In general, the solute concentration of river water tends to be inversely
related to flow rate, such that at very high flow rates runoff may be nearly as dilute as rainwater
(Hem, 1985). This observation stems from the fact that direct runoff has a relatively short
contact time with soil or vegetation. Thus, an increase in streamflow is generally accompanied
by a decrease in solute concentration due to dilution by waters containing a lower concentration
of particular ions than originally present in the stream. An exception is where the chemical
constituent is insoluble and present within the stream sediment load, such that a positive
correlation between stream discharge and concentration may be found (Langmuir, 1997). In
contrast, water infiltrating through the soil zone and interacting with rocks undergoes more
extensive reactions due to the longer residence time, resulting in baseflow that has a higher
concentration of dissolved solids. Therefore, groundwater discharge to a stream would be
expected to increase the concentration of solutes derived from the soils and/or geology.
Temperature is another aspect of climate that can influence stream chemistry through different
rates of weathering, biogeochemical reactions and evaporation. While weathering and
biogeochemical reactions differentially affect the concentration of the solutes in a stream,
evaporation results in an increase in the concentrations of all dissolved components in the
stream, constant ion ratios, and enrichment in stable isotopes of water. These are important
diagnostic features for distinguishing between processes. Climates characterised by distinct wet
and dry periods may favour weathering reactions and therefore produce larger amounts of
soluble inorganic matter during particular seasons of the year (Hem, 1985). Also related to the
seasons are flow and river stage, which influence the variability and overall concentration of
Chapter 7
178
solutes in-stream (McNeil et al., 2005). Therefore, in catchments such as the Herbert that are
distinctly seasonal in regards to temperature, rainfall and flow, river chemistry would be
predicted to change between the seasons due to one or more of these factors.
7.1.1.2 Geologic factors
Dissolution and leaching of minerals are major sources of dissolved solutes reaching surface
waters. Therefore, the hydrochemical signature of a stream is characteristic of the dominant
source(s) of water, such as a particular aquifer system. Reactions between streambed sediments
and the stream can have an additional impact on the observed hydrochemistry.
The noble gas radon (222Rn) is derived from radioactive decay of uranium-series isotopes;
therefore, aquifers containing uranium-bearing minerals will also contain 222Rn. Due to the short
half-life of 3.8 days and loss via gas exchange with the atmosphere, high radon concentrations
are only present in surface waters in the vicinity of points of groundwater discharge (inflow)
and at relatively short distances downstream of such locations (Cook et al., 2003). For these
reasons 222Rn is a particularly useful environmental tracer for identifying zones of preferred
groundwater discharge into surface water bodies. In contrast to dissolved inorganic tracers, 222Rn is also conservative and is not modified by biogeochemical reactions. Whilst the
concentration of 222Rn in groundwater is significantly different to that in surface waters, radon
data must be interpreted with caution, as variations in radon concentration along a river may be
due to numerous factors. For example, aquifers comprised of different geology may have an
inherently different composition of uranium and hence radon. In addition, tributaries supplied
by groundwater may also contribute radon to the stream that is not related to direct inflow from
an aquifer. A multi-tracer approach is thus desirable, coupled with a good understanding of the
hydrogeology. In addition to examining qualitative trends, 222Rn data can be used to derive
estimates of groundwater discharge to surface waters, as described in Stieglitz (2005) and Cook
et al. (2004). The approach is applied to data collected as part of this study (Section 7.5).
7.1.1.3 Biogeochemical factors
Biochemical processes that affect the chemistry of surface waters include processes that require
a net energy input, such as photosynthesis; redistribute chemically stored energy; convert
chemically stored energy; or do not involve energy transfer (Hem, 1985). Given the focus in this
thesis on nitrogen in surface water and groundwater, bacteria-assisted transformation reactions
such as nitrification and denitrification (Section 5.1.1.4, Chapter 5) represent important
biogeochemical mechanisms that influence the concentration and speciation of nitrogen in
water. The concentrations of other solutes such as silica and trace elements such as iron may
also be controlled by biological processes (Hem, 1985).
Chemical River-Groundwater Interactions
179
7.1.1.4 Mixing of waters
Groundwater is generally more enriched in dissolved solids than runoff due to the longer time
for mineral-water reactions to occur. This means that discharge of baseflow to a stream typically
results in mixing between hydrochemically distinct waters, leading to a shift in hydrochemical
signature i.e. a change in solute concentrations, ion ratios, water quality indicators (e.g. pH, Eh,
temperature), and isotope composition. In addition to groundwater, a stream receives water from
tributaries that may also have a distinct chemical signature due to contributing sources and in-
stream reactions. This results in mixing relationships between different waters. A further
hydrochemical mixing trend occurs where fresh and saline waters meet in the estuary of a river.
This mixing zone fluctuates in position depending on river flow, winds and ocean tides (Hem,
1985). Mixing relationships in the tidal zone are complex because of the greater density of
seawater compared to freshwater and hence vertical and horizontal variability. Nonetheless, as
shown in this chapter, the tidal zone represents an important part of the river system that
warrants hydrochemical investigation because of its junction between terrestrial and marine
waters. Hydrochemical consequences of mixing between fresh groundwater and saline water,
both in the Herbert River estuary and the sea, were also discussed in Chapter 5 (Sections 5.4.2
and 5.6.3).
The potential mixing relationships within a stream can generally be distinguished if the end
member compositions are known. For instance, if the hydrochemical signatures of groundwater
and runoff have been identified, then the proportion of mixing can be calculated by a simple
mass balance relationship for solutes that differ in concentration between the two waters. The
inherent assumption is that the solute behaves conservatively upon mixing. Therefore, if Cl is
the constituent of interest, the mass balance relationship is described by:
[Cl]M = x[Cl]A + (1-x)[Cl]B (7-1)
where [Cl]M is the Cl concentration of the mixture; [Cl]A and [Cl]B are the Cl concentrations of
water A and B, respectively; x is the fraction of water A; and (1-x) is the fraction that water B
contributes to the mixture.
7.1.2 Assessment principles and methods
Given the variety of factors that can potentially influence the chemistry of the river, different
tracers are required to distinguish between processes (refer to Section 5.1.1.1 for a summary of
tracer types). An environmental tracer is particularly useful for estimating groundwater inflows
to rivers where the concentration of the tracer in groundwater is relatively uniform and
significantly different to that in the river (Cook et al., 2003). Importantly, when surface water
chemistry is used to identify stream-aquifer interactions, the groundwater component in the
river is the cumulative result of the hydrogeological and hydrological processes along the entire
Chapter 7
180
upstream watercourse length. Therefore, the interpretation of environmental tracer data requires
knowledge of upstream processes that affect both surface and groundwater quality. Based on a
review of tracer techniques, REM (2002) conclude that methods for investigating groundwater-
surface water interactions fall into two main groups:
• Stream routing approaches: where the concentration of the tracer of interest is measured
along the entire stream reach, allowing identification of zones of groundwater
discharge.
• Groundwater approaches: where the concentration of a tracer in (nested) piezometer
transects is measured and compared to stream water concentrations, allowing
determination of stream water contributions to groundwater.
Using a variety of tracers, a combination of these approaches is used in this chapter in order to
characterise the dynamics of the system (Section 2.4.1.4).
Table 3-2 (Chapter 3) summarised the field and laboratory parameters that were measured in
surface water samples and the primary reason(s) for their measurement. In order to assess
processes in detail this extensive dataset of hydrochemical information must be analysed in a
targeted manner. Various methods for organising groundwater data in order to assist with
hydrochemical interpretation were presented in Section 5.1.2 (Chapter 5). Similar techniques are
also applicable to surface water samples, such as Piper, Schoeller and bivariate plots. In
addition, spatial relationships are examined in this chapter through longitudinal plots, while
temporal hydrochemical trends are explored through time series analyses.
The physical and chemical factors outlined in Section 7.1.1 result in: (i) a change in the
concentration of one or more constituents in the river; and/or (ii) a change in chemical signature,
which may be at a single location in the stream or follow a trend along the river. Therefore, the
challenge is to select appropriate chemical species and relationships with other parameters that
will reveal conclusive information about specific processes that impact on the chemistry of
surface waters. The key natural physicochemical processes identified above were: overland
flow, evaporation, groundwater discharge, and mixing with tributaries and seawater.
Denitrification and nitrification were also noted as important biogeochemical processes that
specifically influence stream nitrogen concentrations. Summarised in Table 7-1 are the
distinctive chemical characteristics observed in a stream arising from the range of processes that
can influence the stream’s chemistry. Salinity is an important chemical measure in the stream
for numerous processes and can be represented by the total dissolved solids (TDS) content as
well as total dissolved ions (TDI) and electrical conductivity (EC). While each of these
measures represents the majority of dissolved constituents, there are important differences. As
summarised by McNeil and Cox (2000), TDS is the concentration of dissolved substances in
Chemical River-Groundwater Interactions
181
water, including mineral and organic matter, whether or not in ionic form: this includes SiO2. In
contrast, TDI is the total number of ions in solution, whether they are dissociated or not: the
TDI is calculated as the sum of the major ions expressed in mg/L. EC is the ability of a solution
to conduct an electrical current: it is not affected by dissolved silica or undissociated ions such
as H2CO3 which do not carry an electric charge. McNeil and Cox (2000) note that TDI is the
most useful salinity measure in hydrological studies, although EC is the most convenient to
measure. Both of these measures are analysed in this chapter.
Table 7–1 Distinctive chemical characteristics of processes that influence the chemistry of surface waters. Key attributes used in the analyses are highlighted. Process Influence on stream chemistry Overland flow Decrease in salinity (EC, TDI) and dilution of ion concentrations; increase in
Cl/HCO3
Isotope composition of the stream may reflect mixing between different waters (baseflow-dominant and rainfall)
Inverse relationship between solute concentration and flow
Generally low pH (4-6)+
Evaporation Enrichment in stable isotopes of water, particularly in δ18O relative to δ2H
Increase in salinity (e.g. EC, TDI)
Constant ion ratios
Groundwater discharge
Receiving stream has a hydrochemical signature (ions/ratios/isotopes) similar to that of groundwater (Chapter 5) or is consistent with a mixing trend
Presence of radon
Greater salinity than surface waters
Often more reducing (anaerobic) than surface waters (lower Eh)
Tributary inflow
Receiving stream has a hydrochemical signature (ions/ratios/isotopes) similar to that of the contributing tributary or is consistent with a mixing trend (downstream of the tributary entry point)
Seawater mixing
Increase in salinity (TDS, TDI, EC) and other solute concentrations except silica
Decrease in radon concentration
Increase in pH (8.1)
Mixing trends between fresh and oceanic water (ions/ion ratios/isotopes) +pH varies from 4-6 (continental) to 5-6 (coastal) based on global atmospheric precipitation, with initial rainfall from a given event tending to be the most acid (Langmuir, 1997)
7.1.3 Study site & terminology
As discussed in Chapter 3, water quality data were collected as part of this study soon after
cessation of the wet season (May) and at the end of the dry season (October) in 2004, as well as
in the early part of the dry season (June) in 2005. The Herbert River gorge marks the upstream
extent of the lower catchment area. The most upstream samples were collected just below the
Chapter 7
182
gorge at Nash’s Crossing (within Yamanie National Park), whilst the furthest downstream
samples were collected below gauge 116001 within the tidal zone of the river (Figure 7-2). Note
that Nash’s Crossing is located upstream of sugarcane production areas. To aid with the
analysis, the Herbert River is divided into a series of reaches (1 - 24): the downstream extent of
each reach is marked by the location at which a full water quality sample was collected (for
major ion analysis). As this thesis is concerned with river-groundwater interactions, river
reaches are numbered from upstream to downstream, consistent with the general direction of
groundwater flow (Chapters 4 and 5). Therefore, reach 1 refers to the reach downstream of the
Herbert River gorge, to Nash’s Crossing (2.9 km). Locations along the river are also referred to
in the text as distance in km downstream from the gorge, as this provides a better indication of
distances between observed hydrochemical changes in the stream than reach numbers (Figure
7-2).
U
U
#S#S
#S#S
#S#S
#S
#S #S
#S
#S#S
#S#S#S#S
#S#S#S #S #S#S#S#S
#S
#S
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#S
#S
#S
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#S#S #S
#S
#S
#S
#S #S
#S#S
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#S #S#S
#S#S
#S
#S
#S
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#S
#S
#S
#S
#S
#S#S
#S
#S#S
$T
$T
$T$T
$T
$T$T
$T
$T
$T
$T
$T$T
$T
$T
$T
$T
$T$T
$T$T
r
r
Elphinstone Ck
Gow
rie Ck
Stone R
Lannercost Ck
Ripple C kH aw
k ins
Ck
Dal
rym
ple
Ck
#
Herbert RGorge
Palm C k
Tre bonne Ck
Seymour R
HERBERT RIVER
2.9 km
34.0 km
57.4 km
68.3 km
76.4 km
84.5 km 93.8 km
100.0 km
24.4 km 44.5 km
HinchinbrookIsland116006
116001
May 2004$T
October 2004#S
June 2005#SRainfall 2004r
N
5 0 5 Km
Figure 7-2 Surface water sampling sites during the three collection periods and locations of rainfall samples. Selected reach numbers and distances downstream of the Herbert River gorge are also indicated. Note that the two QDNRW stream gauges are located at approximately 33 km (116006) and 76 km (116001) downstream of the gorge.
In this chapter the months of May-October are referred to as dry season months, while
November-January are months of the wet season, noting that the beginning of the wet season
varies from year to year and that there is a transition period between seasons. The beginning and
end of the dry season are interchangeably referred to as higher flow and lower flow periods,
respectively.
1
2
3
4
5
6 8
910
11
16
18
20 14
24
22
Nash’s Crossing
Chemical River-Groundwater Interactions
183
7.2 GENERAL HYDROCHEMISTRY
Hydrochemical data are analysed in this section to establish general chemical characteristics of
the Herbert River and some of its tributaries. Given the focus in this chapter on river-
groundwater interactions, surface water compositions are also compared with that of
groundwater (Chapter 5).
The Piper diagram depicted in Figure 7-3 illustrates that the Herbert River is relatively enriched
in HCO3, except in the tidal zone, where Cl is the dominant anion. Na, Ca and Mg comprise the
major cations. Note that while the Herbert River samples appear to define a linear trend in the
Piper diagram, this trend does not translate to a systematic evolution in chemistry along the
river. The exception is for samples in the estuary with Cl > 60 % of the major anions, which
correspond to progressively increased proportions of seawater towards the river mouth. Overlap
in relative major ion concentrations is observed between the Herbert River and groundwaters as
well as with seawater, consistent with waters contributing to the river from multiple sources.
However, given the narrow trend for surface waters compared to groundwaters, multiple sources
might be expected to give rise to considerable scatter if the groundwater inputs are as variable as
indicated in Figure 7-3. Surface waters appear to be relatively more enriched in Mg and
depleted in SO4 compared to groundwaters. A summary of major and minor inorganic chemistry
for all surface water samples is provided in Appendix A.
80 60 40 20 20 40 60 80
20
40
60
80 80
60
40
20
20
40
60
80
20
40
60
80
Ca Na HCO3 Cl
Mg SO4
Herbert RTributariesRainfallSeawaterHSsHSd
Figure 7-3 Piper diagram for surface water and groundwater samples collected during three sampling periods: May 2004, October 2004, June 2005.
Chapter 7
184
A modified Gibbs diagram (Figure 7-4), whereby TDI is plotted instead of TDS (compare with
Figure 7-1), shows that samples collected along the lower Herbert River are generally of low
salinity (TDI) and are Na-dominated, similar to the Type D water types of McNeil et al. (2005)
only with a dominance of HCO3 over Cl. Type D waters are characteristic of the steep, high
rainfall, northern coastal streams of Queensland which are typically associated with granitic
terrains that release Na, K and Ca upon weathering of feldspar. Low salinity samples also have a
Na:Cl ratio (meq) of greater than 1, indicative of Na released during silicate weathering
(Meybeck, 1987). Similarly, according to the classification scheme proposed by Meybeck
(2004), based on the sum of cations as an indicator of weathering, the freshwater reaches of the
Herbert River generally correspond to dilute or medium dilute waters, indicative of rainfall and
silicate weathering controls. The higher TDI estuarine samples, which have Na/Cl (molar) equal
to 0.86 and Na/(Na+Ca) approximating that of seawater, reflect an oceanic origin of salts
(Figure 7-4). These waters correspond to the Type A water type of McNeil et al. (2005)
associated with evaporation and/or tidal influences.
1
10
100
1000
10000
100000
0.0
0.5
1.0
Na/(Na + Ca) (meq/L)
TDI (
mg/
L)
Seawater
Rainfall
Groundwater
HR - QDNRW
HR - this study
Evaporation
Dilution
Weathering
Figure 7-4 Modified Gibbs diagram with logarithmic plot of TDI against Na/(Na+Ca) for water samples collected along the lower Herbert River (HR): estuarine samples are circled. Groundwater (both aquifers), seawater and rainfall (collected inland from the coast) are also shown (adapted from McNeil et al. 2005).
7.2.1 Compositional groups
Samples were collected along the entire length of the lower Herbert River, including within the
tidal reaches. Schoeller plots indicate that there are two main compositional groups, with some
samples representing mixtures between these end members (Figure 7-5a).
Chemical River-Groundwater Interactions
185
Consistent with the Piper and Gibbs diagrams (Figure 7-3 and Figure 7-4), surface water in the
tidal zone has a hydrochemical signature that resembles seawater. In contrast, freshwater
compositions are generally more depleted in Cl and SO4 and relatively enriched in HCO3 and
Ca, similar to the Ca-Mg groundwaters of the shallow and deep aquifers (Figure 7-5b). In
addition to Schoeller plots, saturation indices provide evidence of hydrochemical similarities
between the Herbert River and Ca-Mg enriched groundwaters. Note the two distinct groups
within the low salinity trend depicted in Figure 7-5a which are further examined in Section
7.3.2.2.
a
b
Figure 7-5 Schoeller plots of (a) samples in the Herbert River, with seawater highlighted in red and (b) Ca-Mg enriched groundwaters of the shallow aquifer (compare also with the Ca-Mg group of the deep aquifer, Figure 5-7b).
7.2.2 Stable isotopes
The theory behind the use of stable isotopes and factors that cause isotopic fractionation were
outlined in Chapter 5 (Section 5.1.1.3). Stable isotope compositions of samples collected in May
2004 in the Herbert River and two of its tributaries (Stone River and Ripple Creek, Figure 7-2)
are provided in Figure 7-6, with groundwater compositions also indicated. Surface waters
generally cluster around the LMWL, derived from analyses of groundwaters in the lower
catchment (Figure 5-3, Chapter 5). Although isotopic enrichment is observed between some
consecutive reaches (e.g. reaches 9-11), as well generally from upstream (reach 1) to
downstream (e.g. reach 22), there is no evidence of a continuous evaporation trend (inset Figure
7-6). Longitudinal isotopic trends are further analysed in Section 7.4.3.1. Depletion in heavy
isotopes in the upper reaches of the Herbert River (reaches 1-3) is consistent with the elevation
effect, while the isotope composition of Ripple Creek, which lies on the LMWL, is consistent
with enrichment due to rainfall from a short-duration event (the amount effect). Note that the
isotope composition of some shallow and deep groundwaters is similar to that of the Herbert
River, indicative of common source waters.
Herbert River Ca-Mg enriched (HSs)
Chapter 7
186
-7 -5 -3 -1 1-40
-30
-20
-10
0
10Herbert RStone RRipple CkRainfallSeawaterHSsHSd
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
SMOW
coastal rainfall MoretonBay
GMWL
Local meteoricwater line
Brisba
ne
WL
12
3
5
9
18
11
20
16
24
22
14
-7 -5 -3 -1 1-40
-30
-20
-10
0
10Herbert RStone RRipple CkRainfallSeawaterHSsHSd
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
SMOW
coastal rainfall MoretonBay
GMWL
Local meteoricwater line
Brisba
ne
WL
-7 -5 -3 -1 1-40
-30
-20
-10
0
10Herbert RStone RRipple CkRainfallSeawaterHSsHSd
δ2H (o/oo, SMOW)
δ18O (o/oo, SMOW)
SMOW
coastal rainfall MoretonBay
GMWL
Local meteoricwater line
Brisba
ne
WL
12
3
5
9
18
11
20
16
24
22
14
12
3
5
9
18
11
20
16
24
22
14
12
3
5
9
18
11
20
16
24
22
14
Figure 7-6 Oxygen-18 (δ18O) and deuterium (δ2H) stable isotope data for surface water and groundwater samples and a coastal rainfall event in May 2004. The isotopic composition of seawater at Moreton Bay (QLD) (Cresswell 2006, pers. comm.) and SMOW are also shown. Trend lines represent the LMWL (blue); GMWL (black solid); and Brisbane water line (black dashed) (see Figure 5-3). The inset shows isotopic compositions of samples along the Herbert River (labelled by reach number) and the Stone River (near the confluence with the Herbert River), as well as groundwaters with similar isotopic values.
7.3 TEMPORAL DATA
Historical measurements of field parameters and a suite of major and minor elements are
available from QDNRW at several stream gauges in the catchment. These datasets, combined
with data collected for this study, are analysed below to examine temporal hydrochemical trends
in surface waters in the catchment. Although the lower catchment, including its two gauges
(116001 and 116006), is the focus of the analysis, data from an upper catchment gauge
(116004) are included for comparison where applicable. Figure 7-7 illustrates the location of the
stream gauging stations in relation to the entire catchment. Given the emphasis of this thesis on
nitrogen, temporal trends for different forms of N in the lower Herbert River are also examined
based on data from CSIRO (Bramley and Muller, 1999).
Chemical River-Groundwater Interactions
187
%U
%U
%U
116004
116006
116001
stream gauges%U
lower Herbert Riverstreams
#S upstream sampling site
20 0 20 kmN
#
Herbert RGorge
#S
Figure 7-7 Selected QDNRW stream gauges along the lower (116006, 116001) and upper (116004) Herbert River which are referred to in the following text. Also shown is the location of the most upstream sampling site below the gorge, at Nash’s Crossing.
7.3.1 Field parameters
Automatic recorders installed by QDNRW have collected daily water quality data at gauge
116001 since November 1999. Plots of each parameter relative to streamflow highlight distinct
seasonal patterns, especially with respect to electrical conductivity (EC) and temperature (T)
(Figure 7-8). The following sections examine trends within and between seasons for these field
parameters based on available time series data.
Chapter 7
188
1
10
100
1000
Nov
-99
May
-00
Nov
-00
May
-01
Nov
-01
May
-02
Nov
-02
May
-03
Nov
-03
May
-04
Nov
-04
May
-05
Fiel
d pa
ram
eter
s
0
100
200
300
400
500
600
700
800
Flow
(GL/
day)
EC @ 116001 pH @ 116001T @ 116001 Flow @ 116001
Figure 7-8 Time series water quality data relative to flow at gauge 116001. Vertical lines approximate the beginning of the dry season (May). Source: QDNRW
7.3.1.1 Electrical conductivity
Electrical conductivity displays a marked negative correlation with streamflow at gauge 116001
(Figure 7-9). Rapid declines in river salinity correspond with the rising stage of individual
streamflow events, with minimum EC values correlating with streamflow peaks (Figure 7-10a).
In contrast, there is an observed salinity rise after a streamflow event (falling limb of the stream
hydrograph). In addition to salinity fluctuations during streamflow events, the salinity in the
river shows a distinct seasonal trend of low EC during the wet season, rising steadily to a
maximum value by the end of the dry season (Figure 7-10b).
0
40
80
120
160
1 10 100 1000 10000 100000 1000000
Flow (ML/day)
EC
(uS/
cm)
Figure 7-9 Electrical conductivity versus flow at gauge 116001. Source: QDNRW
EC
pH
T
Flow
Chemical River-Groundwater Interactions
189
0
50
100
150
200
250
Jan-
02M
ar-0
2
May
-02
Jul-0
2
Sep
-02
Nov
-02
Jan-
03
Mar
-03
May
-03
Jul-0
3
Sep
-03
Nov
-03
Jan-
04
Mar
-04
May
-04
Jul-0
4
Sep
-04
Nov
-04
Jan-
05M
ar-0
5
EC (u
S/cm
)R
ainf
all (
mm
)
1
10
100
1000
10000
100000
1000000
Flow
(ML/
day)
EC @ 116001Rainfall @ 32045Flow @ 116001
a
0
50
100
150
200
250
Jan-
02M
ar-0
2
May
-02
Jul-0
2
Sep
-02
Nov
-02
Jan-
03M
ar-0
3
May
-03
Jul-0
3
Sep
-03
Nov
-03
Jan-
04M
ar-0
4
May
-04
Jul-0
4
Sep
-04
Nov
-04
Jan-
05M
ar-0
5
EC
(uS/
cm)
Rai
nfal
l (m
m)
1
10
100
1000
10000
100000
1000000
Flow
(ML/
day)
EC @ 116001
Rainfall @ 32045Seasonal flow @ 116001
b
Figure 7-10 Continuous daily electrical conductivity at gauge 116001 relative to streamflow and rainfall: (a) depicts the actual stream hydrograph based on daily data, while (b) shows the approximate seasonal stream hydrograph (60-day moving average). Dotted circles in (b) show the three distinct clusters of EC values at particular stages along the stream hydrograph. Source: BoM (rainfall); QDNRW (flow)
The declining EC trends are consistent with flushing and dilution of dissolved salts in the river
as relatively less saline water (runoff) is introduced into the river system during streamflow
events. In contrast, concentration of dissolved solutes can occur as a result of evaporation and/or
baseflow contributions (including groundwater). Evaporation is especially plausible throughout
the course of the dry season, characterised by a progressive increase in river temperature (Figure
falling limb
peak
falling limb peakrisinglimb
WET DRY
Chapter 7
190
7-13). However, assuming that EC is proportional to TDS, the theoretical trajectory that EC
would follow under different rates of evaporation is defined by the parallel trends depicted in
Figure 7-11. Given that these evaporation lines are much steeper than the observed EC trend at
different flow rates (Figure 7-9), evaporation cannot be the dominant mechanism driving the
increase in EC as flow decreases. However, evaporation is evident in Figure 7-11 where EC
values parallel the evaporation trends. Therefore, while evaporative effects cannot be
discounted, the most likely explanation for the increasing trends in river EC, within individual
events and seasonally, is an increase in the proportion of baseflow discharged to the river
relative to surface runoff. This is consistent with the large baseflow flux determined by
hydrograph separation (Section 6.3.2.3, Chapter 6). Assuming that the EC of baseflow remains
approximately constant throughout the year, seasonal EC values in the river can be interpreted
as tracking particular stages within the seasonal stream hydrograph. As illustrated in Figure
7-10b for 2004, the lowest EC values occur during the streamflow peak, when streamflow is
dominated by surface runoff. After the peak, a higher EC is maintained in the river due to
increased proportions of baseflow contributed by streambank and groundwater discharges. The
highest EC in the river is attained by the end of the dry season, when low flows in the river are
sustained by groundwater. A similar pattern is observed for other years, although the three
stages are particularly evident in 2004 because of the dilution effect caused by high rainfall
during the 2004 wet season. Figure 7-11 further illustrates that EC in the river during
intermediate flows reflects mixing between baseflow-dominated and runoff-dominated waters.
Note that the minimum EC measured in shallow and deep groundwater samples is
approximately 100 μS/cm; above this value there is negligible surface flow contribution (Figure
7-11). EC can thus be used as an indicator of the pure baseflow component.
0
40
80
120
160
1 10 100 1000 10000 100000 1000000
Flow (ML/day)
EC (u
S/cm
)
baseflow
mixing
runoff
evaporation lines
minimum EC (gw)
Figure 7-11 Electrical conductivity versus flow at gauge 116001, depicting theoretical evaporation lines starting from different flow/EC combinations. The minimum EC measured in groundwater (gw) is indicated. Distinct clusters of EC values corresponding to baseflow-dominated, runoff-dominated and mixing between these waters, are also highlighted. Source: flow and EC data from QDNRW
Chemical River-Groundwater Interactions
191
7.3.1.2 pH
River pH can be influenced by factors such as the pH of the source water, turbidity (affecting
the dissolved CO2 concentration) and mineral/weathering reactions in the river-bed/water
column. In addition to natural factors, human influences such as stream pollution can influence
the chemistry of a stream (Hem, 1985). Although pH at gauge 116001 does not display marked
seasonal fluctuations compared to the other water quality parameters (e.g. Figure 7-8), time
series data indicate that large streamflow events produce a rapid decline in pH (more acidic)
followed by a return to neutral values within a few days of the event (Figure 7-12). The pH of
rainfall generally varies from 4-6 (continental) to 5-6 (coastal) (Langmuir, 1997). Therefore,
depressed pH values during large streamflow events are consistent with a high rainfall input. In
addition, the rapid pH response to streamflow events is consistent with the contribution of low
pH waters to the Herbert River from overland flow. As the lower catchment is dominated by
sugarcane farming, which uses nitrogenous fertilisers, soil acidity can result (Wood et al.,
2003); acids can hence be leached out of the soils during runoff events. High concentrations of
ammonia are also likely to be present in the atmosphere during crop fertilisation (Freyney et al.,
1994) that can lower the pH of precipitation (Hem, 1985). The low correlation between pH and
streamflow in the dry season suggests that factors other than streamflow influence river pH
during low flows. Further analysis is beyond the scope of the thesis.
6.0
6.5
7.0
7.5
8.0
Jan-
02
Mar
-02
May
-02
Jul-0
2
Sep
-02
Nov
-02
Jan-
03
Mar
-03
May
-03
Jul-0
3
Sep
-03
Nov
-03
Jan-
04
Mar
-04
May
-04
Jul-0
4
Sep
-04
Nov
-04
Jan-
05
Mar
-05
pH
10
100
1000
10000
100000
Flow
(ML/
day)
pH @ 116001
Flow @ 116001
Figure 7-12 Continuous daily pH at gauge 116001 relative to streamflow. Source: QDNRW
Chapter 7
192
7.3.1.3 Temperature
River temperature can be influenced by numerous interrelated factors such as air temperature;
temperature of the source water (eg. rainfall, groundwater); and volume of water in the water
column. Time series temperature data highlight that the mean temperature of the Herbert River
closely tracks the mean air temperature. Troughs in stream temperature coincide with minimum
air temperatures in July. The temperature of the stream increases in the latter half of the dry
season, consistent with higher air temperatures and lower volumes of water in the river.
10
20
30
40
Nov
-99
May
-00
Nov
-00
May
-01
Nov
-01
May
-02
Nov
-02
May
-03
Nov
-03
May
-04
Nov
-04
May
-05
Tem
pera
ture
(o C)
1
10
100
1000
10000
100000
1000000
Flow
(ML/
day)
Mean air T @ 32045 River T @ 116001 Flow @ 116001
DRY SEASON
WET SEASON
Figure 7-13 Time series of mean air temperature, mean river temperature and streamflow at gauge 116001 in the lower Herbert River. Source: BoM (air T); QDNRW (river T and flow)
Given the strong positive correlation between river and air temperatures (Figure 7-14a), climate
is considered to be a dominant influence on the temperature of the Herbert River. Furthermore,
as groundwater generally has a temperature range between 25-29oC, an increase in baseflow
contribution to the stream (as discussed in Section 7.3.1.1) could be an additional factor that
influences river temperatures in the latter part of the dry season. The split regressions displayed
in Figure 7-14b for river temperatures less than 25oC, between 25-29oC, and greater than 25oC,
are consistent with groundwater buffering the mean river temperature relative to the mean air
temperature towards the end of the dry season.
Chemical River-Groundwater Interactions
193
y = 0.92x + 0.82R2 = 0.82
10
20
30
40
15 20 25 30 35
Mean water T (oC)M
ean
air T
(o C)
a
y = 1.2x - 6.0R2 = 0.61
y = 0.59x + 10R2 = 0.21
y = 1.0x - 2.4R2 = 0.53
10
20
30
40
15 20 25 30 35
Mean water T (oC)
Mea
n ai
r T
(o C)
< 25 degrees
> 25-29 degrees
>29 degrees
b
Figure 7-14 Correlation between mean air and river temperatures at gauge 116001 in the lower Herbert River: (a) regression over all samples, (b) regression separated for river temperatures < 25 oC, 25-29 oC and > 26 oC. Source: BoM (air T); QDNRW (river T).
7.3.2 Major ions
7.3.2.1 Inter-seasonal trends
Trilinear plots based on historical major ion chemistry indicate that there is overlap in the
proportions of major anions in the Herbert River at each of the stream gauges (Figure 7-15).
However, the composition of the stream in the upper catchment (gauge 116004) is distinctly
more enriched in Mg compared to the lower catchment gauges (116006 and 116001). Rivers in
Queensland generally reflect geographical factors such as geology and climate (McNeil et al.,
2005); hence, the observed hydrochemical clusters are consistent with differences in geology
and soil types, which are basaltic in origin in the upper catchment and granitic in the lower
catchment. Different flow characteristics of the Herbert River in the two catchments were also
noted in Chapter 6 (Section 6.3.2.2). Figure 7-15 also highlights that the composition of the
river at gauge 116001 is more variable than at the gauges upstream. Comparison of wet and dry
season data indicate that this variation in hydrochemistry occurs during months of the wet
season. For example, Schoeller plots of samples at the downstream gauge (116001) illustrate
that during the wet season there are both Cl-enriched and Cl-depleted signatures relative to
Chapter 7
194
HCO3 (Figure 7-16a). Comparison of HCO3 and Cl with streamflow indicates that while the
concentrations of both anions decrease with increasing flow, the decrease in HCO3 is
proportionally greater (Figure 7-17). Hence, the composition of the Herbert River during the
wet season reflects the timing of sample collection relative to the flow event, with an increase in
Cl/HCO3 close to the flow peak. An increase in the Cl/HCO3 ratio is therefore a diagnostic for
overland flow. Note that Schoeller plots for dry season samples at gauge 116001 resemble the
Cl-depleted signature of the wet season (Figure 7-16b), consistent with a constant source of
water to the river throughout the year.
80 60 40 20 20 40 60 80
20
40
60
80 80
60
40
20
20
40
60
80
20
40
60
80
Ca Na HCO3 Cl
Mg SO4
Gauge 116001Gauge 116006Gauge 116004
a
Figure 7-15 Piper plot of historical major ion compositions (1973-2004) of the Herbert River at three stream gauges: 116001 (downstream) and 116006 (upstream) are located in the lower catchment while 116004 is located in the upper catchment. Source: QDNRW
a b
Figure 7-16 Schoeller plots of historical (a) wet season and (b) dry season water quality samples collected at gauge 116001 since the 1970’s. Samples highlighted in red correspond to waters that are relatively enriched in Cl relative to HCO3. Source: QDNRW
Chemical River-Groundwater Interactions
195
y = -3.88Ln(x) + 33.05R2 = 0.70
0
10
20
30
40
50
60
0.01
0.10
1.00
10.0
0
100.
00
1000
.00
Flow (GL/day)
HCO
3 (m
g/L)
wet season
dry season
a
y = -0.78Ln(x) + 9.66R2 = 0.57
0
5
10
15
20
0.01
0.10
1.00
10.0
0
100.
00
1000
.00
Flow (GL/day)
Cl (m
g/L) wet season
dry season
b
Figure 7-17 HCO3 and Cl concentrations against streamflow at gauge 116001 during months of the wet and dry seasons. Source: QDNRW
Although the frequency of collection of major ion data at gauge 116001 is in general biannual,
examination of time series data highlight that anions of HCO3, Cl and SO4 display a consistent
increase in concentration during the dry season, followed by a decrease in response to wet
season rainfall. This seasonal trend is particularly apparent for HCO3, which is the dominant
anion in the river. Major cations (Na, Ca, Mg) and SiO2 display a similar seasonal trend to the
anions. The seasonal patterns are consistent with the EC trends identified in Section 7.3.1.1:
dilution occurs during runoff events as relatively fresh surface waters are introduced to the river,
while the subsequent increase in concentrations can be attributed to increased baseflow.
Chapter 7
196
7.3.2.2 Intra-seasonal trends
Similar to the inter-seasonal trends, within the freshwater zone of the river (upstream of tidal
influence), Schoeller plots illustrate that months at the beginning (May/June) and end (October)
of the dry season also have distinct chemical traces. For example, by the end of the dry season
the composition of the Herbert River shifts to higher concentrations of the major cations and
anions, except for Cl (Figure 7-18a). Although the absolute concentration of Cl does not change
significantly, there is a change in the proportion of Cl relative to HCO3 such that the Cl/HCO3
ratio decreases by the end of the dry season.
a
b
Figure 7-18 Major ion chemistry for samples collected in the lower Herbert River during the beginning (May 2004, June 2005) and end (October 2004) of the dry season. Schoeller plots in (a) are for all freshwater compositions except at Nash’s Crossing (reach 1), with end of dry season samples highlighted in red; (b) samples collected at the most upstream sampling site (Nash’s Crossing, Figure 7-2) during months at the extremes of the dry season.
Whilst a change in hydrochemistry can be attributed to numerous factors (as outlined in Section
7.1.1), the EC data examined in Section 7.3.1.1 indicated a transition from surface runoff-
dominated to baseflow-dominated flow during the course of the dry season. Furthermore, based
on physical relationships it was established in Chapter 6 that groundwater is an important source
of recharge to the river, particularly during low flow conditions. Therefore, the shift in
hydrochemical signature is consistent with a change in the composition of water discharging
into the Herbert River. Figure 7-18b illustrates that in contrast to the other reaches, the
hydrochemical signature at reach 1 (Nash’s Crossing) is similar at the two extremes of the dry
season, particularly in regards to relative anion concentrations. This suggests that baseflow
contributions downstream of Nash’s Crossing are the dominant influence on the chemistry of
the lower Herbert River at the end of the dry season, consistent with the baseflow analysis in
Chapter 6 (6.3.2.3).
May 2004October 2004June 2005
end of dry season
beginning of dry season
Chemical River-Groundwater Interactions
197
7.3.2.3 Tributaries of the Herbert River
Schoeller plots for selected tributaries of the lower Herbert River are provided in Figure 7-19.
Similar to the Herbert River (e.g. Figure 7-18a), the plots indicate a shift in hydrochemical
signature at the end of the dry season compared to the beginning. An increase in the
concentration of major ions and a decrease in the Cl/HCO3 ratio are generally observed,
consistent with a change in tributary source waters during the course of the dry season. With the
exception of Hawkins Creek, the hydrochemical trends at the end of the dry season (and
beginning for most tributaries) resemble the Ca-Mg enriched groundwaters of the shallow and
deep aquifers (Figure 5-5 and Figure 5-7, Chapter 5). This observation is supported by
saturation indices, which highlight the hydrochemical similarity between groundwater
compositions and tributaries in the upper part of the catchment (Figure 7-20). Hence, as noted
for the Herbert River, groundwater is a likely source. Furthermore, the intra-seasonal
hydrochemical trends of the tributaries suggest that the change in composition of the lower
Herbert River during the dry season can be attributed to a change in the contribution of baseflow
to the river from both indirect (via tributaries) and direct groundwater discharge. Note that the
hydrochemical trends observed particularly in Hawkins Creek and Ripple Creek (at the
beginning of the dry season) resemble the Na > Cl water type in HSs (Figure 5-5c, Chapter 5).
Given the uncertainty in the origin of this groundwater type (Table 5-1, Chapter 5) it is beyond
the scope of the thesis to further assess these trends.
Chapter 7
198
Figure 7-19 Schoeller plots for tributaries of the lower Herbert River during the beginning (May 2004, June 2005) and end (October 2004) of the dry season (refer to Figure 7-2 for tributary locations). Red trend lines highlight measurements at the end of the dry season.
Chemical River-Groundwater Interactions
199
-7
-6
-5
-4
-3
-2
-1
0
1
Mag
nesi
te
Cal
cite
Gyp
sum
Cha
lced
ony
Dol
omite
log
SI
71A
69A
Gow rie Ck
Elphinstone Ck
Dalrymple Ck
Figure 7-20 Saturation indices (logarithmic form) for bores and tributaries in the upper part of the catchment based on samples collected at the end of the dry season (October 2004).
7.3.3 Nitrogen
Temporal trends for nitrogen in the lower Herbert River are examined in this section based on
data from a water quality project conducted by CSIRO (Bramley and Muller, 1999). Although
dissolved inorganic nitrogen (DIN) is of particular interest to this study, total particulate N and
dissolved organic nitrogen (DON) data are also analysed in order to determine the relative
contribution of DIN to the total N load of the river. Sampling sites along the Herbert River
include Nash’s Crossing, Abergowrie Bridge, John Row Bridge and Gairloch Bridge, which lie
approximately 2 km, 44 km, 78 km and 81 km downstream of the gorge, respectively (Figure
7-2). A summary of N chemistry was presented in Chapter 5 (Section 5.1.1.4).
7.3.3.1 Particulate vs dissolved N
Analysis of time series nitrogen data (total, soluble and particulate) shows that the highest total
N concentrations in the river generally correspond to high flow events in the wet season, which
are dominated by particulate N (Figure 7-21). As observed in Figure 7-21b and also noted by
Mitchell et al. (1997), the concentration of particulate N increases with rising flow and reaches
a maximum close to the flow peak. In contrast, the dry season is generally dominated by
dissolved forms of N. Based on the available data, the proportion of soluble N in the Herbert
River relative to total N is generally between 80-90 % during the dry season months.
Chapter 7
200
0
1500
3000
4500
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
Tota
l N (u
g N/
L)
1
100
10000
1000000
Flow
(ML/
day)
Nash's Crossing Abergowrie bridgeJohn Row bridge Gairloch bridgeFlow @ 116001 Flow @ 116006
a
0
1500
3000
4500
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
Tota
l par
ticul
ate
N (u
g N/
L)
1
100
10000
1000000
Flow
(ML/
day)
Nash's Crossing Abergowrie bridgeJohn Row bridge Gairloch bridgeFlow @ 116001 Flow @ 116006
b
0
1500
3000
4500
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
Tota
l sol
uble
N (u
g N/
L)
1
100
10000
1000000
Flow
(ML/
day)
Nash's Crossing Abergowrie bridgeJohn Row bridge Gairloch bridgeFlow @ 116001 Flow @ 116006
c
Figure 7-21 Time series concentrations of (a) total; (b) particulate and (c) soluble N at various sites along the lower Herbert River. Flow at gauges 116006 and 116001 is also shown. Source: N data from Bramley and Muller (1999); flow data from QDNRW
Chemical River-Groundwater Interactions
201
7.3.3.2 Dissolved organic vs inorganic N
Examination of dissolved N at each sampling location along the lower Herbert River indicates
that the concentration of DON is generally higher than DIN, particularly at the upstream
sampling site (Figure 7-22). DON does not show a strong seasonal pattern but is highly variable.
In contrast, there is a noticeable trend in DIN characterised by a marked increase at the
beginning of the wet season followed by a decline to a base concentration during the dry season.
Accordingly, the proportion of DIN relative to total dissolved N ranges from a maximum of 70
% during the start of the wet season, to less than 2 % by the end of the dry season: this
represents a maximum of 30% and down to 1% of total N. As illustrated in Figure 7-22 and also
observed by Mitchell et al. (1997), high DIN in the wet season corresponds to the first major
flow event (first flush) following the dry season (high availability of DIN in the catchment),
while a decrease in DIN occurs around the flow peak due to dilution by rainwater.
0
100
200
300
400
500
600
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
ug N
/L
1
100
10000
1000000
Flow
(ML/
day)
DIN Abergowrie bridge DON Abergowrie bridge
Flow @ 116001
a
0
200
400
600
800
1000
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
ug N
/L
1
100
10000
1000000
Flow
(ML/
day)
DIN John Row bridge DON John Row bridge
Flow @ 116001
b
Figure 7-22 Concentrations of dissolved N as inorganic (DIN) and organic (DON) components at (a) upstream and (b) downstream sampling sites along the lower Herbert River. Flow at gauge 116001 is also shown. Source: N data from Bramley and Muller (1999); flow data from QDNRW
Chapter 7
202
7.3.3.3 Inorganic species
The inorganic components of dissolved nitrogen are comprised of nitrate (NO3-), ammonium
(NH4+) and nitrite (NO2
-), which differ in their oxidation states (Section 5.1.1.4, Chapter 5). As
depicted in Figure 7-23, NO3- is consistently the dominant inorganic species over NH4
+ (NO2- is
generally present in very low concentrations in surface waters). Whilst NO3- in the river varies
during the wet season, the concentration is relatively constant during the dry season. This
observation is consistent with NO3- inputs from groundwater, as discussed in Section 7.6.
0
100
200
300
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
ug N
/L
1
100
10000
1000000
Flow
(ML/
day)
NO3 Abergowrie bridge NH4 Abergowrie bridge
Flow @ 116001
a
0
300
600
900
Jan-
1992
Jan-
1993
Jan-
1994
Jan-
1995
ug N
/L
1
100
10000
1000000
Flow
(ML/
day)
NO3 John Row bridge NH4 John Row bridge
Flow @ 116001
b
Figure 7-23 Nitrate and ammonium concentrations at (a) upstream and (b) downstream sampling sites along the lower Herbert River. Flow at gauge 116001 is also shown. Source: N data from Bramley and Muller (1999); flow data QDNRW
7.4 LONGITUDINAL DATA
The above analyses have highlighted that there are differences in the hydrochemistry of the
Herbert River and its tributaries both inter-seasonally (between the wet and dry seasons) as well
as intra-seasonally (between the beginning and end of the dry season). Notably, as streamflow
declines the:
Chemical River-Groundwater Interactions
203
• salinity of the river (and tributaries) increases due to an increase in major ion
concentrations, not due to evaporation alone;
• Cl/HCO3 ratio of the river (and tributaries) decreases due to relative enrichment in
HCO3; and
• mean river temperature decreases relative to air temperature.
In addition, there is similarity between the hydrochemical signature of surface waters and Ca-
Mg-HCO3 enriched groundwaters of the shallow and deep aquifers. These collective
observations provide strong evidence of baseflow contributions to the Herbert River and its
tributaries. Moreover, consistent with the flow characteristics examined in Chapter 6 (Section
6.3.2), the available hydrochemical data demonstrate that the contribution of groundwater
increases from the wet season to the end of the dry season. Hence, there is a seasonal shift from
runoff-dominated to baseflow-dominated waters contributing to the Herbert River.
Given the above hydrochemical characteristics of surface waters, the aim of this section is to
examine chemical trends along the length of the lower Herbert River that may indicate variation
in water sources to the stream. Field parameters as well as other solutes are analysed for
longitudinal trends, both within particular sampling periods, and between different periods in
the dry season (refer to Appendix A for the entire dataset). The range of parameters analysed
provide different levels of information. Thus, EC is useful for examining broad hydrochemical
trends, whilst individual solutes yield information on small variations due to particular
processes/influences. Isotopes such as 222Rn are a particularly powerful tool for identifying
recently added groundwater sources contributing to the Herbert River (Section 7.1.1.2). Note
that no sampling was undertaken during the wet season. However, samples collected at the
beginning (May, June) and end (October) of the dry season provide an indication of
hydrochemical changes in the river as flow declines.
7.4.1 Salinity
Electrical conductivity was recorded in the field at regular intervals along the length of the
lower Herbert River and selected tributaries. As observed in Figure 7-24a, there is a marked
increase in EC within the tidal zone, which extends to a maximum of approximately 25 km
upstream from the mouth of the river (75 km downstream from the gorge). Compared to the
tidal reach there is little variation in EC along the freshwater section of the river. However,
detailed examination highlights that EC progressively increases downstream from the top of the
catchment, particularly during lower flow conditions (Figure 7-24b). This is indicative of a
source of dissolved constituents downstream from the gorge and/or evaporation.
Chapter 7
204
Note that there is a strong positive relationship between EC and TDI, hence EC is considered to
be a suitable indicator of river salinity for this population of waters during the sampling periods
(McNeil and Cox, 2000). Although the longitudinal variations are similar, there is a distinct
difference in the magnitude of salinity in the river and its tributaries between the beginning and
end of the dry seasons. These observations are discussed in the sections below in the context of
the individual constituents of salinity.
10
100
1000
10000
100000
0 20 40 60 80 100
Distance downstream from gorge (km)
EC (u
S/c
m)
October 2004June 2005
tidal limit
a
40
80
120
160
200
240
0 20 40 60 80
Distance downstream from gorge (km)
EC (u
S/cm
)
Oct-04
Jun-05Tribs - Oct
Tribs - June
21 3 4 6 8 10 12 15 17
b
Figure 7-24 Measurements of field electrical conductivity along the lower Herbert River and its tributaries (Tribs) during the beginning (June) and end (October) of the dry season. Note that (b) excludes the high salinity section of the river; selected reach numbers corresponding to sample locations are provided on the upper horizontal axis.
Chemical River-Groundwater Interactions
205
7.4.2 Major ions in the tidal zone
In general, there is little longitudinal variation observed in the concentrations of major ions
within the freshwater zone of the river. However, consistent with the EC measurements, there is
a marked increase in the concentrations of most constituents within the tidal zone, except HCO3
and SiO2 (Figure 7-25). Similarities with the relative proportions of major ions in seawater (e.g.
Figures 7-3, 7-4, 7-5) confirm that the high solute concentrations are due to mixing with
seawater. In addition, the dramatic increase in the Mg/Ca ratio is consistent with a seawater
influence (Figure 7-26) (Vengosh, 2004).
June 2005
0.1
1.0
10.0
100.0
1000.0
10000.0
0 20 40 60 80 100
Distance downstream from gorge (km)
Con
cent
ratio
n (m
g/L) HCO3
ClSO4NaCaMgKSiO292 km
HCO3
SiO2
tidal limit
October 2004
0.1
1.0
10.0
100.0
1000.0
10000.0
0 20 40 60 80 100
Distance downstream from gorge (km)
Con
cent
ratio
n (m
g/L) HCO3
ClSO4NaCaMgKSiO286 km
HCO3
SiO2
tidal limit
Figure 7-25 Longitudinal comparison of major ion and oxide concentrations along the lower Herbert River during months representing the beginning (June) and end (October) of the dry season. Seawater influence is observed in samples downstream of the vertical dotted lines. The longitudinal trends for HCO3 (green) and SiO2 (dark blue) are also labelled.
Chapter 7
206
Within the tidal zone, the extent of seawater influence on the chemistry of the Herbert River is
determined by the river stage and sampling time in relation to the tidal patterns (relative sea
level). Hence, the location and chemistry of the mixing zone varies seasonally as well as daily.
For example, in samples from October 2004 mixing is evident downstream of 86 km (reach 20),
while in June 2005 seawater influence is apparent at 92 km (reach 23) downstream from the
gorge (Figure 7-25 and Figure 7-26).
The near-constant concentrations of HCO3 and SiO2 upstream of, and within the tidal zone, are
due to the relatively small difference in concentration of these dissolved constituents in seawater
compared to freshwater. For instance, the amount of SiO2 in seawater is approximately 5 mg/L
compared to around 6-7 mg/L in freshwaters of the river; therefore, in the zone of seawater
mixing there is only a minor decrease (dilution) in SiO2. Conversely, HCO3 is slightly enriched
in seawater (142 mg/L) compared to freshwater (20-40 mg/L); therefore, a minor increase in
concentration is observed in the tidal zone. In comparison, the concentrations of the other major
ions are at least two orders of magnitude greater in seawater than in freshwater of the Herbert
River, resulting in a marked change in their concentrations where fresh and saline waters mix.
Although longitudinal variations in major anions, cations and SiO2 within the freshwater
reaches are relatively small in magnitude, they are important for detecting variations in source
waters contributing to the river. Concentration differences between higher flow and lower flow
periods also provide insight into different processes influencing the lower Herbert River during
the extremes of the dry season, as examined below.
0.0
1.0
2.0
3.0
4.0
5.0
6.0
1 2 3 4 5 6 7 8 9 10 11 12 14 15 16 17 18 20 21 23 24
Reach number
Mg/
Ca (m
eq/L
)
October 2004June 2005
Figure 7-26 Mg/Ca ratio for samples collected along the lower Herbert River during October 2004 and June 2005. Reaches are numbered from the gorge to the river mouth (refer to Figure 7-2).
Chemical River-Groundwater Interactions
207
7.4.3 Processes in the freshwater zone
As discussed in Section 7.1.1, there are a number of processes that can influence the chemistry
of the Herbert River. Along the freshwater reaches, the key processes include evaporation,
overland flow, tributary inflow, and groundwater discharge (Table 7-1). The analyses below aim
to distinguish between these processes based on major ion and isotope trends. Whilst
concentrations were measured for a number of dissolved constituents, longitudinal trends are
illustrated for selected solutes and their ratios that demonstrate the main hydrochemical
variations along the river attributed to the key processes. Note that although the downstream
extent of the freshwater zone is variable between sampling periods (e.g. Figure 7-25 and Figure
7-26), for ease of comparison longitudinal plots representing freshwater compositions are
depicted from the gorge (reach 1) to 86 km downstream (reach 18).
7.4.3.1 Evaporation
An increase in salinity and individual ion concentrations, constant ion ratios, and isotopic
enrichment, are indicative of evaporation (Table 7-1). It was noted in Section 7.2.2 that
enrichment in stable isotopes is observed between some consecutive reaches in May,
particularly between reaches 1-2 and 9-14 (Figure 7-27). Although the Na/Cl ratio is virtually
constant between reaches 1-2, the δ2H/δ18O slope between these reaches is -1.3, compared to a
typical evaporation slope from open waters in the range 4-6 (Gat, 1980). In contrast, between
reaches 9-11, the δ2H/δ18O slope is 4.6, the corresponding TDI increases, and Na/Cl is virtually
constant over the reach (Figure 7-27). However, the δ2H/δ18O slope between reaches 11-14 is
10.8, which is incompatible with evaporation. Therefore, whilst there is evidence for
evaporation along at least one section of the river in May (e.g. between reaches 9-11), based on
the available data evaporation is not considered to be a major influence on the chemistry of the
river at the beginning of the dry season. It was noted in Section 7.3.1 that evaporative effects
would be expected to increase towards the end of the dry season as a result of an increase in
river and air temperatures. However, while TDI generally increases from upstream to
downstream in October, ion ratios such as Na/Cl are not constant between consecutive reaches
displaying large salinity increases (e.g. between reaches 1-2 in October) (Figure 7-27). This is
consistent with the conclusions from the temporal EC analyses in Section 7.3.1.1: whilst
evaporation cannot be discounted, it is not the major mechanism driving the increase in river
salinity through the dry season. The collection of complementary stable isotopic data at the end
of the dry season would assist in quantifying the evaporative influence along different reaches
of the river.
Chapter 7
208
40
50
60
70
80
90
100
0 20 40 60 80
Distance downstream from gorge (km)
TDI (
mg/
L)
May-04
Oct-04
21 3 4 6 8 10 12 15 17
a
0.5
1.0
1.5
2.0
2.5
3.0
0 20 40 60 80
Distance downstream from gorge (km)
Na/
Cl (
meq
/L)
May-04
Oct-04
21 3 4 6 8 10 12 15 17
b
-5.5
-5.0
-4.5
-4.0
-3.5
0 20 40 60 80
Distance downstream from gorge (km)
18O
(o / oo S
MO
W)
May-0421 3 5 9 11 14 16 18
c
Figure 7-27 Salinity, Na/Cl and oxygen-18 (δ18O) along the lower Herbert River. Reach numbers corresponding to each sample location are provided on the upper horizontal axis.
Chemical River-Groundwater Interactions
209
7.4.3.2 Overland flow
Key diagnostic features of overland flow include a decrease in salinity and dilution of ions
present in the stream prior to the runoff event. In addition, it was noted in Section 7.3.2 that the
Cl/HCO3 ratio is related to flow such that the ratio increases close to the flow peak. Whilst
sampling was undertaken during periods in the dry season, with little or no expected overland
flow, rainfall occurred during the sampling period in June 2005: mean discharge increased by
approximately 200 ML/day at the upstream gauge (116006) and 2000 ML at the downstream
gauge (116001). In contrast, in May and October 2004 sampling was undertaken during the
declining phase of the stream hydrograph, with there being no major runoff events throughout
the periods of sample collection. The effect of the flow event in June 2005 on the chemistry of
the Herbert River is clearly evident as a trough (dilution) in longitudinal plots between reaches
7-10, whereby the concentrations of all constituents except for Cl and SO4 decline compared to
adjacent reaches (Figure 7-28). In contrast, all solute concentrations increase within this section
of the river during May and October 2004. As Cl and SO4 do not similarly decrease, this
suggests that there is a source of these solutes in the contributing catchment area; the ions are
thus readily mobilised during a runoff event. The ratio of Cl/HCO3 also increases over this reach
in June which provides further evidence of overland flow.
0.2
0.3
0.4
0.5
0 20 40 60 80
Distance downstream from gorge (km)
meq
/L
Na (June)
HCO3 (June)
Cl (June)
21 3 4 6 8 10 12 15 17
reaches 7-10
Figure 7-28 Longitudinal trends for selected major ions along the lower Herbert River in June 2005. Reach numbers corresponding to each sample location are provided on the upper horizontal axis, with reaches 7-10 highlighted.
Chapter 7
210
7.4.3.3 Tributary inflow
Numerous tributaries enter the Herbert River, particularly in the upper reaches which
collectively drain a significant area of the catchment. Based on saturation indices and Schoeller
plots it was observed that there is similarity between the hydrochemistry of groundwater and
most tributaries, especially at the end of the dry season (Section 7.3.2.3). In addition, high radon
activities in measured tributaries (Figure 7-29) indicate a groundwater supply during low flow
conditions (Section 7.1.1.2). Therefore, where a tributary is groundwater-fed, it is difficult to
distinguish between direct groundwater discharge and tributary inflow as the source of water to
the Herbert River. Nonetheless, visual comparisons between solute concentrations within the
tributaries and upstream and downstream of the entry point into the river indicate that
decreasing Cl between reaches 1-4 is consistent with inflow of low Cl tributaries such as
Gowrie Creek and Elphinstone Creek (Figure 7-30a). However, additional sources (other than
Gowrie Creek) must be invoked to account for the concentrations of solutes such as HCO3 and
Mg between reaches 1-2 of the river (Figure 7-30b, c). These sources could include the other
tributaries south of the river (Figure 7-2) and/or direct groundwater discharge.
0.0
1.0
2.0
3.0
4.0
5.0
6.0
Reach number
222 Rn
(Bq/
L)
21 3 4 6 8 10 11 12 145 7 9 15 16 17 18
October 2004
Dal
rym
ple
Ck
Haw
kins
Ck
Ston
e R
Rip
ple
Ck
Figure 7-29 Radon (222Rn) activities along the freshwater reaches of the lower Herbert River and sampled tributaries (blue) in October 2004. Reach numbers on the horizontal axis relate to the river, while sampled tributaries are individually labelled.
Assuming that the solute changes in the Herbert River are due entirely to tributary inflow, the
required salt load (TDI x flow) contributions from the tributaries can be quantified to test
whether additional sources must be invoked. Based on the TDI in measured tributaries and the
change in TDI of the river (upstream and downstream of each tributary), the required discharge
from the tributaries to account for the river solute loads is estimated (Table 7-2). In the absence
of flow gauges on the tributaries (with the exception of Gowrie Creek), tributary flow rates are
estimated based on field measurements of flow velocity (assuming 0.1 m/s if not visibly
flowing) and river depth and width (assuming a triangular cross-sectional area of the river).
Chemical River-Groundwater Interactions
211
0.1
0.2
0.3
0.4
0.5
0.6
Reach number
Cl (
meq
/L)
21 3 4 6 8 10 11 12 145 7 9 15 16 17 18
October 2004
Gow
rie C
k
Elph
inst
one
Ck
Dal
rym
ple
Ck
Haw
kins
Ck
Ston
e R
Rip
ple
Ck
a
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1.0
1.1
Reach number
HCO
3 (m
eq/L
)
21 3 4 6 8 10 11 12 145 7 9 15 16 17 18
October 2004
Gow
rie C
k
Elph
inst
one
Ck
Dal
rym
ple
Ck
Haw
kins
Ck
Ston
e R
Rip
ple
Ck
b
0.0
0.1
0.2
0.3
0.4
0.5
Reach number
Mg
(meq
/L)
21 3 4 6 8 10 11 12 145 7 9 15 16 17 18
October 2004
Gow
rie C
k
Elph
inst
one
Ck
Dal
rym
ple
Ck
Haw
kins
Ck
Ston
e R
Rip
ple
Ck
c
Figure 7-30 Longitudinal plots for selected ions along the freshwater reaches of the lower Herbert River and sampled tributaries (blue) in October 2004. Reach numbers on the horizontal axis relate to the river, while sampled tributaries are individually labelled.
Chapter 7
212
Table 7–2 Required and estimated flow rates in tributaries of the lower Herbert River and groundwater based on changes in total solute loads upstream and downstream of selected tributaries in October 2004
Tributary TDI of tributary
(mg/L)
Δ solute load in Herbert R (106 mg/day)+
Required tributary discharge (ML/day)
Estimated tributary discharge
(ML/day)*
Required gw input
(ML/day)#
Gowrie Ck 71 5100 70 20 38
Elphinstone Ck 62 8300 130 80 36
Dalrymple Ck 91 -300 -3 10 0
Stone R 140 3500 30 60 0
Ripple Ck 120 300 3 4 0 + Defined as downstream minus upstream load estimated from Herbert River gauges (Figure 7-7) * Estimated from field measurements of flow velocity and river width and depth (gauge value assumed for Gowrie Ck) # TDI of contributing aquifer (HSd) is 96 mg/L (near Gowrie Ck) and 93 mg/L (near Elphinstone Ck)
The calculations indicate that tributaries such as Gowrie Creek and Elphinstone Creek most
likely contribute to the solute load in the Herbert River; however, additional sources are
required to explain the total solute loads in the river. Assuming a groundwater contribution, the
required groundwater input to account for the solute loads in the river is estimated to be around
40 ML/day. While there are additional tributaries above and below the confluence of Gowrie
Creek within reach 2 (Figure 7-2) that could be supplied by groundwater, there are no major
tributaries other than Elphinstone Creek within reach 4; hence, direct groundwater discharge to
the river is plausible. The mass balance indicates that inflow of Dalrymple Creek does not
contribute to the change in total solute load in the river. Note that Hawkins Creek was not
included in the calculations as the TDI is well below that of the upstream and downstream
concentrations in the river; the tributary also has a very small catchment area (Figure 7-2) and
would be unlikely to contribute to noticeable flow (or solutes) in the river at the end of the dry
season. However, the calculations indicate that while the estimated discharge rates are small,
solute loads from the Stone River and Ripple Creek are consistent with loads in the Herbert
River. Furthermore, in the absence of tributaries in between reaches 8-17, the increase in solute
loads must be attributed to other sources, such as direct groundwater discharge or evaporation.
7.4.3.4 Groundwater discharge
Based on hydrochemical similarities between surface water and groundwater (Section 7.2) and
temporal chemical trends in the river (Section 7.3) it was established that groundwater is a
source of water to the river, with an increasing relative contribution towards the end of the dry
season. Furthermore, longitudinal plots along the freshwater reaches of the river highlight a
marked change in the concentration of solutes between reaches 1 and 2 (3-34 km downstream),
particularly in October, which cannot be accounted for by evaporation alone (e.g. Figure 7-27
Chemical River-Groundwater Interactions
213
and Figure 7-31). In the absence of overland flow during this period, other processes must be
invoked. Similarly, it has been shown that variations in ion concentrations and ratios along other
reaches are not consistent with evaporation as the dominant processes influencing the chemistry
of the river at the end of the dry season. Based on the available information, groundwater
discharge is the most likely mechanism, either directly or via groundwater-fed tributaries. In
addition, longitudinal trends in the concentration of radon (refer to Section 7.5) demonstrate that
groundwater contributes over the entire length of the Herbert River during months at the
extremes of the dry season. Evidence for changes in the groundwater inflow rate along the river
is examined in Section 7.5.
Variations in ion concentrations in the river are evident between reaches 6-11, particularly in
October, for Na, Mg and HCO3 (Figure 7-31). Similar oscillations in cation concentrations are
also observed in June; however, due to the flow event the trends are subdued. In addition, the
concentration of Cl gradually increases downstream of reach 5 in May, June and October. These
subtle changes in chemistry are indicative of changes in source waters contributing to the river.
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0 20 40 60 80
Distance downstream from gorge (km)
Na
and
HC
O 3 (
meq
/L)
0.1
0.2
0.3
0.4 Na (Oct)
HCO3 (Oct)
Cl (Oct)
Mg (Oct)
21 3 4 6 8 10 12 15 17
Cl a
nd M
g (m
eq/L
)
reaches 6 -11
Figure 7-31 Longitudinal trends for selected major ions along the lower Herbert River in October 2004. Reach numbers corresponding to each sample location are provided on the upper horizontal axis, with reaches 6-11 highlighted.
Due to the complex stratigraphy and variable degree of interconnectivity between the aquifers
(Chapter 4), it is difficult to isolate which aquifer(s) is the dominant influence on the river along
particular reaches. However, given that HSd is the main aquifer in the northwest upland area,
where it is also incised by the Herbert River (Figure 4-4) and has the potential to discharge to
the river (Table 6-6), it can be inferred that the deep aquifer discharges into the upper reaches of
the river. However, further downstream, in the vicinity of reach 6, the channel depth decreases
such that the river only intersects the shallow aquifer. Moreover, as discussed in Chapter 5
(Section 5.2.3.2), there is evidence of an additional deeper sand unit in the shallow aquifer that
Chapter 7
214
is more enriched in Na, Mg, HCO3 and Cl than the upper unit of HSs. Hence, the enriched and
depleted concentrations of these ions within reach 6-11 are consistent with relative changes in
the contributing sand units of the shallow aquifer. Based on the available evidence, the changes
in river chemistry downstream of reach 5 are consistent with spatial variation in the
hydrogeology and hence hydrogeochemistry.
7.5 TRACING GROUNDWATER
The application of radon as a tracer of groundwater discharge to surface waters was discussed in
Section 7.1.1.2. Due to potential losses of radon through processes such as gas exchange with
the atmosphere and radioactive decay, the radon content of a stream represents the balance of
additions and losses. Using the formulation of Cook et al. (2004) and under the assumption that
radon contributions to the stream are solely from groundwater inflow (either directly or via
groundwater-fed tributaries), changes in radon content in the stream can be expressed as:
dwckwcIcxc
i λ−−=∂∂Q
(7-2)
where c is the radon activity in the stream; ci is the activity in groundwater inflow; Q is the
streamflow rate (m3/day); I is the groundwater inflow rate per unit of stream length (m3/day/m);
k is the gas transfer velocity across the water surface (m/day); λ is the radioactive decay
constant (day-1); w is the width of the river surface (m); d is the mean stream depth; and x is the
distance in the direction of flow. In the absence of surface water inflow or direct rainfall input,
change in flow with distance is given by:
EwIx
−=∂∂Q
(7-3)
where E is the evaporation rate (m/day). Hence, the activity of radon with distance becomes:
cdwkwcwEcccIxcQ i λ−−+−=∂∂ )( (7-4)
The last three terms in equation 7-4 represent changes in activity due to evaporation (increases
radon activity in residual water), gas exchange (decreases radon activity), and radioactive decay
(decreases radon activity). Assuming that for a particular reach of the river the groundwater
inflow rate, evaporation rate and gas transfer velocity do not change with distance, the change in
radon activity over the reach can be expressed as:
λλ dwkwwEIIc
wExIxQQ
dwkwwEIIc
cxc i
p
i
++−+⎟⎟
⎠
⎞⎜⎜⎝
⎛−+
⎟⎠⎞
⎜⎝⎛
++−−=
o
oo)( (7-5)
Chemical River-Groundwater Interactions
215
where wEI
dwkwwEIp−
++−=
λ and Qo is the river flow at the start of the river reach (x = 0).
The above mass balance equations highlight that there are a number of parameters that influence
the net radon flux in the river. The following analyses examine longitudinal trends in radon in
light of these considerations.
7.5.1 Radon distribution in groundwater
Samples were collected at selected bores in the study area for the measurement of 222Rn in
groundwater. Whilst the concentration of 222Rn is relatively constant between sampling periods
(October 2004 and June 2005) there is considerable spatial variation. The ultimate source of 222Rn to groundwater is likely to be uranium-bearing minerals in the granite bedrock that
underlies the alluvial aquifer system. As 222Rn is a gas, the flux to groundwater is related to the
travel time (diffusion rate) of radon from the source. Greater travel distances and retardation
caused by impeding layers (e.g. clays) would be expected to decrease the vertical 222Rn flux due
to radioactive decay. Therefore, aquifer thickness and characteristics such as the degree and
direction of vertical connectivity are plausible factors that could influence the measured radon
content in groundwater. In general it is observed that the average concentration of radon in
groundwater of the shallow aquifer decreases from an average of around 58 Bq/L in the upper
part of the catchment to 9 Bq/L towards the coast. In addition, there is an increase in thickness
of the alluvial profile (Section 4.2.2, Chapter 4) as well as strong upwards vertical flow (Section
5.3, Chapter 5,) towards the coast. Whilst it is beyond the scope of the thesis to quantify in
detail the competing processes that can account for the spatial distribution of radon in
groundwater, the observed decline in concentration is important for interpreting longitudinal
changes in the activity of 222Rn. For example, low concentrations in the river do not necessarily
translate to a proportionately low flux of groundwater (or vice versa), as the input concentration
varies along the river.
7.5.2 Temporal trends in radon along the river
Longitudinal measurements of radon in the lower Herbert River indicate that months at the two
extremes of the dry season display different trends (Figure 7-32). In June there is a continual
decline in the concentration of 222Rn along the entire length of the river, with a marked decrease
downstream of reach 6 (62 km). In contrast, the October trend is characterised by an increase in
the flux of 222Rn for approximately 58 km downstream from the gorge (reaches 1-5), with a
progressive decline below reach 10 (70 km) towards the mouth of the Herbert River.
Furthermore, the concentration of 222Rn just below the gorge is greatly elevated in June
compared to October.
Chapter 7
216
0.0
0.5
1.0
1.5
2.0
0 20 40 60 80 100
Distance downstream from gorge (km)
222 R
n (B
q/L)
OctoberJune
9 10 12 16 18 20 22 2421 3 5 6 8
prior to flow event
dilution after flow event
peak flow at gauge 116006
rising stage at gauge
peak flow at gauge 116001
prior toflow event
gauge 116006
gauge 116001
(tidal limit)
Figure 7-32 Concentration of radon (222Rn) along the lower Herbert River during periods representing the end (October 2004) and beginning (June 2005) of the dry season. The approximate position of the stream gauges in reaches 2 and 12 are also indicated. Reaches in June are labelled with reference to the stages of the streamflow event that occurred during the sampling period.
Disparity between the June and October radon trends can largely be explained by the difference
in river discharge at the extremes of the dry season. As noted in Section 7.4.3.2, samples in June
were also influenced by the streamflow event that occurred during the sampling period. Prior to
the event, mean daily discharge at gauge 116006 (reach 2) in June was 770 ML/day compared
to 116 ML/day in October 2004; at gauge 116001 (reach 12) mean daily discharge was 1380
ML/day in June compared to 450 ML/day in October. Whilst higher discharge and
corresponding river width may result in greater loss of radon due to gas exchange (equation 7-
4), there is a greater reach contribution of radon at higher flow rates (due to the extra distance
that a molecule of radon can travel prior to decay). Therefore, the elevated concentration of 222Rn in the upper and lower reaches (reaches 1-3 and 12-24) at the beginning of the dry season
(June) compared to the end (October) is consistent with differences in stream discharge. A
greater presence of clays in the riverbed, contributing additional radon to the stream, would also
be expected in June (following the wet season) compared to October. Furthermore, samples in
June corresponding to reaches 1 and 2 were collected just after the peak flow recorded at gauge
116006, which may have additionally enhanced the in-stream radon concentration. Note that the
downstream samples (reaches 12-24) in June were collected during the rising stage of the stream
hydrograph and at peak flow recorded at gauge 116001 (Figure 7-32). Therefore, whilst river
discharge in the absence of rainfall is inherently higher at the beginning of the dry season
compared to the end, it is not conclusive whether the observed in-stream concentrations of 222Rn
in June are elevated in the tidal zone due to the flow event. Further sampling downstream of
Chemical River-Groundwater Interactions
217
reach 11 at the beginning of the dry season (in the absence of a flow event) would be required to
confirm this.
In the middle reaches of the Herbert River the longitudinal radon trends cross over in June and
October. Samples in June corresponding to reaches 7-10 were collected two days after the peak
of the flow event, which resulted in an increase in mean discharge of approximately 2000
ML/day at gauge 116001 (reach 12). Note that the increase was of far greater magnitude at the
downstream gauge compared to the upstream gauge (increase of 300 ML/day at gauge 116006).
Therefore, as also observed with major ions (Figure 7-28), the radon concentrations between
reaches 7-10 reflect dilution from overland flow. The reversal in seasonal trends at reach 5 is
consistent with the higher 222Rn concentration of the inflowing tributary (Dalrymple Creek) in
October (4.5 Bq/L) than June (1.9 Bq/L). Additional tributary sampling would be required to
verify if the 222Rn concentration of Elphinstone Creek (between reach 3 and 4) is also elevated
at the end of the dry season.
7.5.3 Relative flux of groundwater along the river
Whilst 222Rn is a useful tracer of groundwater discharge to surface waters, there are numerous
factors that can influence the measured in-stream concentration, that are not necessarily related
to river-aquifer interactions. However, based on the approach of Cook et al. (2004), the
groundwater flux along a reach can be estimated under the assumptions outlined for equation 7-
5. Whilst several of the parameters are not accurately known, the estimates provide an
indication of the relative contribution over the specified reaches and whether a groundwater
source can account for the observed radon concentrations at the beginning and end of a reach.
This semi-quantitative approach is particularly useful given that the measured radon activity in
surface waters is not necessarily proportional to the flux of groundwater. Given that radon
samples in June were influenced by a significant flow event, measurements of radon from
October are considered to be more reliable for examining changes in groundwater discharge
along the river. However, for comparison, estimates in June are also generated for reaches
sampled prior to the flow event.
Inputs into equation 7-5 include a combination of assumed and measured values. The
evaporation rate, gas transfer velocity and decay are assumed to be constant at E = 7mm/day,
k = 1 m/day and λ = 0.18 day-1, respectively (Cook et al., 2004); mean river width, mean depth,
and in-stream radon concentrations (initial and final) for each reach were determined in the
field. The distance between sampling points was determined from a GIS. Given the variability
of 222Rn in groundwater, the input concentration is taken as an average value for bores centered
on a particular reach. Although HSd is the dominant aquifer in the upper reaches of the river, for
simplicity a similar flux is assumed from each aquifer. Therefore, an average radon
concentration of HSd and HSs is used. However, in the lower reaches, only HSs concentrations
Chapter 7
218
are used. Given the available data these assumptions are considered reasonable, which also
allow, in a relative sense, for the dramatic decline in the concentration of radon potentially
discharging to the lower reaches of the river (Section 7.5.1). The streamflow rate was only
recorded at the two gauges along the river (Figure 7-2). Hence, the input value is the average of
the gauge discharge values (on the days of radon sampling for a particular reach), unless the
reach is upstream or downstream of the gauge, in which case the nearest corresponding gauge
value is assumed. Note that within the tidal zone, radon activities have been corrected where
mixing with seawater (dilution of 222Rn) is apparent from major ion analysis. The mass balance
method described in Cook et al. (2004) using electrical conductivity measurements was applied
to determine the proportion of seawater mixing and hence the undiluted concentration of radon
in the stream. Input parameters for the determination of groundwater fluxes are displayed in
Table 7-3.
Table 7–3 Measured input parameters for modelling radon activities in the river (equation 7-5). Refer to text for assumed input values for E, k and λ.
Reach co (Bq/L) ci (Bq/L) Qav (m3/day)+ w (m) d (m)
1-2 0.57 55 120000 53 0.5
2-5 0.81 58 270000 50 0.3
5-6 1.5 35 300000 50 0.4
6-8 1.1 35 290000 50 0.4
8-9 1.4 35 410000 80 0.5
9-10 1.1 35 390000 30 0.4
10-14 1.4 33 390000 50 0.5
14-17 0.74 30 390000 100 1.7
17-19 0.28 16 390000 100 2.4
19-24 0.43 9.2 390000 100 1.6 + Note that Qav is an average of the gauge values on the day(s) that radon samples were collected along a particular reach, which is not necessarily the stream discharge rate at the start of the reach.
An approximate groundwater inflow rate was determined for each reach (using equation 7-5) in
order to minimise the difference between the calculated and measured downstream radon
activities. Discharge estimates compared with measured radon concentrations in October
(Figure 7-33) illustrate that the observed radon trend can be accounted for by changes in the flux
of groundwater. Within the freshwater zone, increases in radon concentration are consistent with
an increase in groundwater flux, while a decrease in radon concentration along a reach
corresponds with a decrease in the groundwater discharge rate. However, downstream of reach
6, the magnitude of the change in groundwater flux between reaches is greater than that for
radon. This is largely due to the fact that a large change in groundwater discharge is required to
Chemical River-Groundwater Interactions
219
compensate for a small change in the concentration of radon. i.e. the estimated groundwater flux
is sensitive to the in-stream radon activity. Note that the groundwater flux estimated in June for
selected reaches sampled prior to the flow event (Figure 7-32) is similar to October estimates
between reaches 2-5; however, between reaches 5-6 the flux is greatly elevated in June
compared to October.
0
1
2
3
4
5
0 20 40 60 80 100
Distance downstream from gorge (km)
222 R
n (B
q/L)
0
1
2
3
4
5
Gro
undw
ater
flux
(m3 /d
ay/m
)
Rn (Oct)
groundwater flux (Oct)
groundwater flux (June)
9 10 1214 17 19 2421 5 6 8tidal limit
seawater mixing
Figure 7-33 Measured radon concentrations and estimated groundwater flux along the lower Herbert River in October 2004 and at selected reaches (2-5 and 5-6) in June 2005 based on the approach of Cook et al. (2004). Reach numbers are displayed on the horizontal axis, with the tidal limit and zone of seawater mixing (determined from major ions in October 2004) indicated.
In the tidal zone the reverse trend is observed compared to upstream, whereby the concentration
of radon declines, while the flux of groundwater increases. This trend arises because of the
increase in river width and depth (at least double) downstream of reach 14 compared to
upstream; a large increase in groundwater flux is thus required to account for the increased
losses from gas exchange and decay (equation 7-4). Sensitivity analysis indicates that the
estimated groundwater inflow is particularly sensitive to river width. In addition, the radon
concentration of groundwater declines within the tidal zone, especially downstream of reach 17
(Table 7-3); sensitivity analysis indicates that the groundwater flux is also responsive to this
parameter. In a study by Cook et al. (2004) it was also found that estimated groundwater
inflows are most sensitive to changes in radon activity of groundwater inflow, followed by the
gas exchange velocity and river width.
The groundwater flux estimates in the freshwater zone can be compared with gauged stream
discharge measurements during the October 2004 sampling period. Gauges 116006 and 116001
are located at approximately the beginning of reaches 2 and 12 (Figure 7-2). During the period
of interest, the difference in recorded discharge between the gauges is approximately
Chapter 7
220
240 ML/day. Given minimal overland flow during October 2004, this discharge difference
represents the baseflow contribution. Correspondingly, the summation of groundwater flux
estimates between reaches 2-12 is around 10 m3/day/m, which equates to around 400 ML/day
(40 km distance between the gauges). Whilst the estimated groundwater contribution from the
radon mass balance is greater than that determined from the difference in streamflow between
the gauges, the two results are of a similar magnitude.
The results of the above radon mass balance approach indicate that groundwater discharges
along the entire length of the lower Herbert River at the end of the dry season, either directly or
via groundwater-fed tributaries. This is consistent with the physical river-groundwater
relationships established in Chapter 6 (Table 6-6), which showed that there is potential for
hydraulic connection along the four main reaches (A-D) and that the dominant direction of flux
is from the aquifers to the river. Whilst the estimated discharge rate is underpinned by numerous
assumptions, the approach highlights the relative changes in groundwater inflow along the river,
with there being a considerably greater flux between reaches 6-10.
7.5.3.1 Comparison with hydrochemistry
Hydrochemical data coupled with the groundwater discharge estimates provide a powerful tool
to verify and describe the river-groundwater relationships along the length of the river. For
example, over the first 58 km of the river downstream from the gorge (reaches 1-5), there is a
marked decrease in Mg/Ca and increase in Na/Cl in the river in October, coinciding with an
increasing trend for 222Rn activity (Figure 7-34). These trends are indicative of increasing
contributions from groundwater sources below the gorge, which have a lower Mg/Ca ratio and
higher Na/Cl ratio than water contributing from the upper catchment. Furthermore, the large
decrease in Mg/Ca is consistent with dilution of Mg/Ca enriched waters associated with basalts
in the upper catchment (e.g. gauge 116004 in Figure 7-15) by lower catchment alluvial
groundwaters. The increase in estimated groundwater flux below the gorge is thus consistent
with the observed hydrochemistry. Similarly, in the middle reaches of the river, changes in ion
ratios (particularly Mg/Ca) coincide with changes in groundwater flux. As noted in Section
7.4.3.4 these hydrochemical variations are indicative of different contributing aquifers. The
corresponding groundwater flux estimates further indicate that there is a change in the river-
aquifer hydraulics such that discharge from groundwater increases downstream of reach 6. Note
that in Chapter 6 (Section 6.3.2.3) it was also established from baseflow filtering that there is an
enhanced contribution of baseflow between the two gauges in the study area. These collective
observations are consistent with a switch from HSd to HSs as the dominant contributing aquifers.
Chemical River-Groundwater Interactions
221
0.8
1.0
1.2
1.40 20 40 60 80 100
Distance downstream from gorge (km)
Mg/
Ca (m
eq/L
)
0
1
2
3
4
5
Gro
undw
ater
flux
(m3 /d
ay/m
)22
2 Rn a
ctiv
ity (B
q/L)
Mg/Cagroundwater fluxRn (Oct)
tidal limit
seawater mixing
15 17 19 249101221 5 6 83 21
a
0.5
1.0
1.5
2.0
2.5
3.0
0 20 40 60 80 100
Distance downstream from gorge (km)
Na/C
l (m
eq/L
)
0
1
2
3
4
5
Gro
undw
ater
flux
(m3 /d
ay/m
)22
2 Rn a
ctiv
ity (B
q/L)
Na/Clgroundwater fluxRn (Oct)
tidal limit
seawater mixing
15 17 19 249101221 5 6 83 21
b
Figure 7-34 Ions ratios, estimated groundwater flux and radon along the lower Herbert River in October 2004. Note that in (a) the vertical scale is reversed for Mg/Ca and that only concentrations along the freshwater reaches are depicted (Mg/Ca increases dramatically in the zone of seawater mixing).
Based on major ion trends it is difficult to verify whether the estimated groundwater flux in the
tidal zone is plausible, as mixing with seawater dramatically alters the solute concentrations in
the river and hence masks the groundwater signal. However, in Chapter 4 (Figure 4-7) it was
established that the vertical head gradient between the aquifers reverses from downwards to
dominantly upwards in the vicinity of reach 17. Hence, the increase in estimated groundwater
discharge in the tidal zone is consistent with the change in inter-aquifer hydraulic relationships.
Chapter 7
222
7.5.4 Uncertainty in groundwater flux
Longitudinal analyses presented above for a variety of environmental tracers have highlighted
that the chemistry of the lower Herbert River changes markedly from the most upstream
location studied (Nash’s Crossing) to approximately 34 km downstream (reach 2), particularly
at the end of the dry season. Moreover, the above analyses have shown that the dominant
process leading to this shift in hydrochemistry is discharge of groundwater to the river. The
groundwater flux along reach 2 of the river was estimated in Section 7.5.3 based on a mass
balance approach using radon. A different mass balance, incorporating the concentrations of
major and minor elements, is examined in this section in order to: (1) quantitatively assess
whether groundwater can explain the observed changes in ion concentrations; (2) compare the
relative differences in baseflow contributions to the river at the extremes of the dry season; and
(3) compare the results of two independent mass balance approaches (involving solute
concentrations or radon), in order to examine the uncertainty in the estimated groundwater flux.
7.5.4.1 Solute mass balance
Applying the principle of conservation of mass and assuming conservative mixing between two
end-members (Section 7.1.1.4), the relationship between river and groundwater concentrations
for the same chemical constituent can be expressed as:
[solute]down = x [solute]gw + (1- x )[solute]up (7-6)
which can be rearranged to give:
[solute] - [solute] [solute] - [solute]
upgw
updown=x (7-7)
where [solute]down, [solute]up and [solute]gw are the respective downstream, upstream and
groundwater concentrations of the chemical constituent of interest. The variable x is defined as
the fraction of groundwater in the stream. Re-writing equation 7-7 gives:
b-c b- a
=x (7-8)
where down[solute]=a , up[solute]=b and gw[solute]=c
i.e. z
yx = (7-9)
where by −= a and bcz −=
Chemical River-Groundwater Interactions
223
Numerous dissolved chemical species were measured in groundwater and the Herbert River
during the sampling periods in May 2004 and June 2005 (beginning of dry season) and October
2004 (end of dry season) (Table 3-2, Chapter 3). Therefore, a mass balance can be applied to all
of the common constituents at the same sampling time. As the approach assumes conservative
mixing between two end-members, the reliability of the estimates is a function of how well the
end-member chemical compositions can be constrained (Herczeg and Edmunds, 2000). An
important aspect of the calculation is capturing the uncertainty in the measurements of
individual solutes and propagating these errors through to the estimation of the contribution
from groundwater. Uncertainty can be introduced at numerous stages throughout the process of
collecting and analysing water samples. Random and systematic errors in the field occur in the
collection, preparation, and preservation of samples, while in the laboratory there are errors
associated with sample processing and the analytical instruments. Superimposed on this is the
variability in the composition of groundwater, which creates an uncertainty in its average value.
While it is difficult to quantify all of the possible error sources, the mass balance presented here
allows for uncertainty in the concentrations of the solutes due to: (1) instrument limits of
detection; and (2) the spatial heterogeneity of groundwater. The limits of detection vary between
solutes and between different laboratories: these differences were taken into account in the
uncertainty for each constituent during the three sampling periods. The concentration of HCO3
was calculated from the alkalinity, which was determined by titration against an acid;
measurement uncertainties were propagated through the calculation to define a limit of detection
for this solute. Under the assumption of local baseflow contributions to the stream along the
reach of interest, the chemistry of nearby bores (deep and shallow) was averaged to give a
representative groundwater composition. Note that the river intersects the deep aquifer in this
area. As there were only three groundwater measurements, the uncertainty in the groundwater
composition was defined as the range (i.e. ± ½ range) between the maximum and minimum
concentrations for each chemical constituent. This uncertainty is a reflection of the spatial
heterogeneity of groundwater in the upper part of the catchment. Assuming that the errors
accounted for are random, the errors on y and z (equation 7-9) can be expressed as:
22 )(a)( by Δ+Δ=Δ and 22 )(c)( bz Δ+Δ=Δ (7-10)
where aΔ and bΔ are the detection limits of the analytical instrument for each chemical
constituent (or error calculated for HCO3) and cΔ is the greater of the two types of error in the
measurement of each solute in groundwater.
Chapter 7
224
The fractional uncertainty in the calculated proportion of groundwater is thus:
222
⎟⎠⎞
⎜⎝⎛ Δ+⎟⎟
⎠
⎞⎜⎜⎝
⎛ Δ=⎟
⎠⎞
⎜⎝⎛ Δ
zz
yy
xx
(7-11)
A best estimate of x based all chemical constituents can thus be computed as:
∑
∑
=
== N
ii
N
iii
best
w
xwx
1
1 (7-12)
for solutes i = 1, 2, ….., N where the weight is defined as 2
1
ii x
wΔ
= (Taylor, 1982).
Therefore, chemical constituents with the greatest amount of uncertainty contribute less to the
final estimate of the proportion of groundwater. The uncertainty in xbest can therefore be
calculated as:
2/1
1
−
=
⎟⎠
⎞⎜⎝
⎛=Δ ∑
N
iibest wx (Appendix C) (Taylor, 1982) (7-13)
where 2
1
ii x
wΔ
= as before.
7.5.4.2 Local baseflow contribution
Based on the approach presented above, a mass balance was undertaken over reach 2. The
approach assumes that other than local baseflow, there are no additional sources of solutes to the
stream. The results of these computations for the three sampling periods are summarised in
Table 7-4.
Table 7–4 Best estimates of the percentage of local baseflow that contributes to streamflow along reach 2 during months in the dry season, based on a solute mass balance (refer to text for details).
May 2004 October 2004 June 2005
% groundwater (xbest) 5.2 30 8.8
uncertainty (Δxbest) 0.8 2 0.9
Chemical River-Groundwater Interactions
225
Where available, the constituents included in the mass balance were F, Cl, SO4, HCO3, Ca, Mg,
Na, K, Si, Zn, NO3, δ18O, and δ2H; solutes with no concentration difference between
downstream and upstream in the river did not contribute to the calculations for that particular
time period. As shown in Table 7-4, the estimated local contribution of groundwater to
streamflow over reach 2 is less than 10 % during the beginning of the dry season and rises to
around 30 % by the end of the dry season. The solutes that contribute most to the estimates (as
determined by the weight wi) include HCO3, Ca, Mg, Na, Si and NO3 towards the beginning of
the dry season, with Cl becoming an important solute and Mg less so, at the end of the dry
season.
The fraction of the error,
i
ibest
xxx
Δ−
(7-14)
for the main contributing solutes is generally around 1 or slightly beyond the error range, except
for NO3, which has a much higher deviation in October 2004 and June 2005 (Table 7-5). This
either suggests that the uncertainty has been underestimated or that there is an additional source
that contributes NO3 to the stream. Given that there were only three bore measurements used to
represent the composition of the groundwater, it is possible that the spatial variability in the
aquifer was underestimated, particularly with respect to NO3. On balance, the generally low
fraction of the error for most constituents suggests that the mass balance approach and the
associated assumptions are reasonable for estimating the local baseflow contribution. While not
all possible water sources have been considered, the change in river chemistry along reach 2 in
each sampling period is consistent with discharge from a local groundwater source in the upper
reaches of the river. In the absence of chemistry data for inflowing tributaries over the reach, it
is not possible to differentiate between direct groundwater discharge versus inflow from
groundwater-fed tributaries. However, there is strong evidence that there is an increase in the
local contribution of groundwater to the stream from the beginning to the end of the dry season.
Given that estimates of groundwater flux using a radon mass balance could not be determined
along reach 2 in June (due to the flow event), results from the solute and radon mass balance
approaches are compared for October only. Along reach 2, the groundwater flux was estimated
at 0.85 m3/day/m using radon, which equates to approximately 17 % of total streamflow
(approximately 200 ML/day at gauge 116006) during that period in October 2004. In
comparison, using a solute mass balance, the proportion of baseflow contribution along the
same reach is around 30 %. Assuming that inflowing tributaries are supplied by groundwater in
October, a comparable proportion of groundwater input of 38 % was estimated along reach 2
from a mass balance of solute loads (Gowrie Creek, Table 7-2).
Chapter 7
226
Note that these estimates represent the local baseflow contribution along the reach, not the total
groundwater flux contributing to streamflow above reach 2 (as determined in Chapter 6). Whilst
instrument errors and the spatial heterogeneity in groundwater composition were considered in
the solute mass balance approach, errors were not considered in the radon mass balance due to
the number of parameters with unknown uncertainties. Therefore, the discrepancy in the
estimated flux between the two approaches can be attributed to an underestimation of the errors.
In relation to the radon approach, sensitivity analysis indicates that the estimated groundwater
flux is sensitive to the radon activity of groundwater inflow as well as the gas exchange velocity
and river width. Therefore, improvement in the measurement of these parameters and capturing
their uncertainty would be important for refining the estimation of groundwater flux using a
radon mass balance. Similarly, additional sources of error could be included in the solute mass
balance. Whilst refinements to both approaches would provide greater confidence in the
estimated groundwater flux, this is beyond the scope of the thesis and is an area for future work.
Table 7–5 Calculated values for xi, Δxi and the fraction of error (equation 7-14) during the beginning (May) and end (October) of the dry season in 2004 where xbest (May) = 5.2±0.8, xbest (October) = 30±2 and xbest (June) = 8.8±0.9 (refer to Table 7-4).
F Cl SO4 HCO3 Ca K Mg Na S Si Zn NO3 δ18O δ2H
May 2004
xi - -22 3.3 17 10 75 15 6.5 14 4.2 50 6.0 -41 5
Δxi - 44 120 6.7 3.6 260 11 1.6 1800 0.91 160 3.9 71 38
fraction of the error - 0.62 0.02 -1.7 -1.3 -0.27 -0.93 -0.78 0 1.1 -0.28 -0.20 0.66 0
Oct 2004
xi 44 15 120 48 49 34 110 32 6700 27 100 13 - -
Δxi 680 17 210 10 11 56 77 2.7 1100000 5.8 280 5.5 - -
fraction of the error -0.02 0.83 -0.43 -1.8 -1.7 -0.09 -1.0 -0.88 -0.01 0.41 -0.26 3.0 - -
June 2005
xi - 37 24 11 14 - 15 11 - 7.8 100 2.9 - -
Δxi - 66 100 3.6 4.2 - 11 1.4 - 1.7 200 1.9 - -
fraction of the error - 0.70 -0.15 -0.69 -1.3 - -0.58 -2.0 - 0.58 -0.46 3.1 - -
Chapter 7
228
7.6 TRANSPORT OF NITROGEN BY GROUNDWATER
The analyses presented in Sections 7.2-7.5 have demonstrated that:
• discharge of groundwater is the major process that influences the chemistry of the lower
Herbert River and its tributaries, particularly at the end of the dry season;
• groundwater discharges along the entire length of river, either directly or via
groundwater-fed tributaries;
• the flux of groundwater varies along the length of the river, with a considerably greater
flux between reaches 6-10 due to a switch from HSd to HSs as the dominant contributing
aquifer; and
• a large flux of groundwater is evident in the tidal zone at the end of the dry season.
In addition, based on temporal trends in species of N (Section 7.3.3):
• dry season months are dominated by dissolved forms, between 80-90 % of total N;
• DIN ranges from 70 % (start of the wet season) to less than 2 % (end of the dry season)
of total dissolved N, representing a maximum of 30% and down to 1% of total N; and
• NO3- is the dominant inorganic species, with a relatively constant concentration during
the dry season.
In this section the distribution of DIN along the lower Herbert River is examined. The above
analyses have provided evidence of groundwater discharge along the river throughout the year;
therefore, the aim of the following analyses is to establish whether groundwater is also a
transport vector for DIN. The concentrations and spatial distribution of DIN in groundwater
were analysed in Chapter 5 (Section 5.6), and summarised in Figure 5-33. Similarly, a
conceptual diagram highlighting the spatial distribution of NO3- and sources to the lower
Herbert River at the end of the dry season is presented in Figure 7-38 at the conclusion of this
section.
Chemical River-Groundwater Interactions
229
7.6.1 Longitudinal trends in DIN
In addition to field parameters and major ions, samples were collected as part of this study for
the three species of DIN along the river and its tributaries. Comparison of longitudinal sections
indicates that the concentration of NO3- does not vary markedly between months representing
the extremes of the dry season. However, speciation in regards to the other forms of DIN differs
between higher and lower flow periods (Figure 7-35). Based on the available data, it is evident
that NO3- and NH4
+ dominate at the beginning of the dry season (NO2- below the detection
limit), while NO3- and NO2
- are the main species at the end of the dry season. Whilst it is
beyond the scope of the thesis to examine N speciation in detail, the observation is consistent
with differences in the environmental conditions of the river, such as in the pH, redox,
temperature and microorganism activity. In addition, due to the transition from surface-
dominated to groundwater-dominated discharge, changes in the sources of N to the river are
plausible.
0.001
0.010
0.100
1.000
10.000
0 20 40 60 80 100
Conc
entr
atio
n (m
g/L)
NO3 (May)
NH4 (May)
16 18 20 249 1121 5
tidal limit
3
seawater mixing
a
0.001
0.010
0.100
1.000
10.000
0 20 40 60 80 100
Distance downstream from gorge (km)
Conc
entra
tion
(mg/
L)
NO3 (Oct)
NO2 (Oct)
NH4 (Oct)
15 18 20 249 10 1221 5 6 8
tidal limit
3
seawater mixing
b
Figure 7-35 Speciation of DIN along the lower Herbert River during months representing the beginning (May) and end (October) of the dry season. The zone of seawater influence is indicated based on major ion chemistry.
Chapter 7
230
7.6.1.1 Comparisons with other tracers
Based on the analysis of major ions and radon in Sections 7.4 and 7.5, it was established that
groundwater discharges along the entire length of the river, with contributions to the upper
reaches from the deep aquifer and from the shallow aquifer downstream of reach 5. Figure
7-36a illustrates that the large increase in NO3- concentration between reach 1-2 in October
corresponds to a decrease in the Mg/Ca ratio of a similar relative magnitude, which is indicative
of increasing groundwater contributions below the gorge and hence evolution of the chemistry
of the river. In addition, oscillations in NO3- downstream of reach 5 are generally paralleled by
shifts in both the Mg/Ca and Na/Cl ratios, consistent with discharge from different units within
the shallow aquifer (Figure 7-36a and b). Mismatch between major ions and NO3- are observed
at approximately 58 km (reach 5) and between 64-68 km downstream (reaches 7-9), where the
pronounced spikes (declines) represent NO3- loss due to denitrification or in-stream biological
reactions. In addition to major ions, 222Rn concentrations along the river display a similar trend
to NO3- (Figure 7-36c). On balance, these observations provide evidence that NO3
-, major ions,
and radon share a common groundwater source, at least at the end of the dry season. Measured
concentrations of NO3- in the shallow and deep aquifers (Section 5.6.2, Chapter 5) are also
consistent with a groundwater origin.
Downstream of reach 10 and upstream of the seawater mixing zone (reaches 10-18), the decline
in both NO3- and 222Rn (Figure 7-36c) corresponds with a decrease in the estimated groundwater
flux (refer to Figure 7-33). The decline in radon can be explained by a combination of
radioactive decay and degassing at a higher rate than the supply of 222Rn in the inflowing
groundwater. Similarly, the decrease in NO3- can be attributed to losses (such as through
denitrification or other in-stream transformations) in excess of the declining flux from
groundwater and/or dilution from NO3--depleted groundwater. There are a range of
environmental factors that can potentially influence the fate of NO3- in inflowing groundwater,
as well in-stream. However, given the similarity in trends between NO3- and 222Rn, the decline
in groundwater flux is a plausible explanation for the decline in the concentrations of both of
these dissolved species.
Chemical River-Groundwater Interactions
231
0
1
2
0 20 40 60 80 100
Distance downstream from gorge (km)
NO
3- (mg/
L)
1
Mg/
Ca
(meq
/L)
NO3
Mg/Ca
9 10 12 14 17 20 2421 5 6 8tidal limit
seawater mixing
3 0.8
1.2
1.4
a
0
1
2
0 20 40 60 80 100
Distance downstream from gorge (km)
NO3- (m
g/L)
0
1
2
3
Na/C
l (m
eq/L
)
NO3
Na/Cl
9 10 12 14 17 20 2421 5 6 8tidal limit
seawater mixing
3
b
0
1
2
0 20 40 60 80 100
Distance downstream from gorge (km)
NO3- (m
g/L)
0
1
2
222 Rn
(Bq/
L)
NO3
Rn (Oct)
9 10 1214 17 20 2421 5 6 8tidal limit
seawater mixing
3
c
Figure 7-36 Comparison of NO3- with ion ratios and radon along the lower Herbert River
in October 2004. The zone of seawater influence is indicated based on major ion chemistry.
Chapter 7
232
Analysis of major elements showed that there is an abrupt change in the hydrochemistry of the
river in the zone of seawater mixing (Figure 7-25). Similarly, the concentration of NO3-
increases in this zone in October (downstream of reach 18). Given that the flux of groundwater
is also estimated to increase in this zone (at least in October, Figure 7-33), NO3- measured in the
river could be the result of direct groundwater discharge. NO3- concentrations in the shallow
aquifer, adjacent to the estuarine reaches of the river (Figure 5-30, Chapter 5), are consistent
with a groundwater source of NO3-. However, mixing with seawater cannot be discounted,
especially given that NO3- levels in seawater can be quite high e.g. 50 mg/L NO3
- has previously
been measured at Moreton Bay (QLD) (Cresswell, 2006, pers. comm.). Whilst the data are
inconclusive with regards to the source of NO3- in the zone of seawater mixing, previous studies
have found evidence of submarine discharge to the near-shore environment adjacent to the
catchment (Gagan et al., 2002; Stieglitz and Ridd, 2000). Therefore, an indirect groundwater
source of NO3- to the river, via mixing with seawater NO3
- sourced from submarine discharged
groundwater, is also possible.
Mass balance calculations can be used to test whether a potential groundwater contribution of
NO3- to the river is plausible (Table 7-6). Similar to the solute load balances in Section 7.4.3.3
(Table 7-2), a NO3- mass balance along reach 2, which displays the greatest increase in NO3
-
(Figure 7-36), highlights a surplus NO3- load in the river which cannot be accounted for by
Gowrie Creek alone. Assuming that other tributaries along the reach are supplied by
groundwater, the required groundwater discharge rate to contribute the NO3- load is
approximately 40 ML/day (given the concentration of NO3- in nearby HSd groundwater). This is
comparable with the estimated groundwater input required to account for the observed change
in solute load along reach 2 (Table 7-2). Hence, groundwater contributes to both the salinity and
NO3- load of the river.
Table 7–6 Required groundwater discharge to account for observed NO3- concentrations
in the Herbert River in October 2004
Tributary NO3- in
tributary (mg/L)
Estimated tributary NO3
-
load (106 mg/day)+
Δ NO3- load in
Herbert R (106 mg/day)*
Unaccounted NO3
- load in Herbert R
(mg/L)#
Required gw input
(ML/day)@
Gowrie Ck 2.7 54 300 250 40
Elphinstone Ck 1.3 110 100 0 0
Dalrymple Ck 1.0 10 -130 0 0
Stone R 1.9 110 30 0 0
Ripple Ck 12 50 -50 0 0 + Estimated tributary discharge from Table 7-2 * Defined as downstream minus upstream concentrations in the Herbert River # NO3
-concentrations in contributing aquifer (HSd upstream of Dalrymple Ck and HSs downstream) @ In addition to inflow from measured tributary (compare with Table 7-2)
Chemical River-Groundwater Interactions
233
Similarly, the other tributaries of Elphinstone Creek and Stone River, which are most likely
groundwater-fed in October, contribute to the NO3- load of the Herbert River. Consistent with
the solute load calculations, Dalrymple Creek does not contribute NO3- to the river; however,
while Ripple Creek contributes other solutes (Table 7-2), it is not a source of NO3- (Table 7-6).
The low Eh of Ripple Creek compared to the river is consistent with denitrification and hence
transformation of NO3- prior to inflow or upon mixing with the river.
7.6.1.2 Comparisons within the dry season
Comparison of longitudinal data at the beginning and end of the dry season indicates that in
general there is an increase in NO3- concentration downstream of reach 1 and a gradual decrease
downstream of reach 10, noting that as for major ions, the concentration of NO3- between
reaches 7-10 is diluted by overland flow during the June sampling period (Figure 7-37).
0.0
0.4
0.8
1.2
1.6
2.0
0 20 40 60 80 100
Distance downstream from gorge (km)
NO
3- (mg/
L)
May 2004
Oct 2004
June 2005
15 18 20 249 1221 5 6 83 10
tidal limit
seawater mixing
seawater mixing
Figure 7-37 Longitudinal plots of NO3- during months representing the beginning (May,
June) and end (October) of the dry season. The zone of seawater influence is indicated based on major element chemistry in the corresponding sampling periods.
Whilst the concentration of NO3- at reach 1 is similar in each sampling period, the rate of
increase for 45 km downstream of the gorge is markedly higher at the end of the dry season
(October) compared to the beginning (May/June). As established from major ions and EC, this
intra-seasonal variation can be attributed to differences in the dominant transport vectors at the
extremes of the dry season i.e. a combination of surface runoff and groundwater discharge at the
beginning of the dry season transitioning to groundwater dominance by the end of the dry
season. Whilst the concentration of groundwater NO3- is likely to remain fairly constant
throughout the dry season, inflow of tributaries between reaches 1-5, such as Gowrie Creek,
may dilute the NO3- concentration in the river during higher flow conditions. Further tributary
sampling during the beginning of the dry season, in the absence of rainfall events, would be
required to test this.
Chapter 7
234
Within the tidal zone, different NO3- trends are also observed during each of the sampling
periods. Disregarding the tidal reaches of the June data, which were possibly affected by the
flow event, the decline in NO3- downstream of the freshwater zone is observed at reach 14 in
May and at reach 18 in October. These differing trends within the estuary indicate that different
processes influence the concentration of NO3- during the extremes of the dry season. There is
insufficient data to explain why these discrepancies are observed; however, it is noted that the
chemistry in the tidal zone at different times of the year is a reflection of the complex balance
between competing hydraulic pressures imposed by the river, the groundwater, and the sea. The
collection of complementary radon data and hence groundwater flux modelling at the beginning
of the dry season (in the absence of flow events) could be an avenue for future research.
The analyses in this section highlight that there are many unanswered questions regarding the
distribution of DIN in the river. However, the analyses have shown that the concentration of
NO3- at the downstream extent of the freshwater zone is distinctly elevated over the
concentration measured below the gorge during both the beginning and end of the dry season.
This demonstrates that NO3- originates from a source in the lower catchment. Furthermore, the
results provide strong evidence that groundwater is an important transport vector for NO3- to the
lower Herbert River, at least during months of the dry season. In addition, the concentration of
NO3- is highly variable along the river; therefore, NO3
- data must be interpreted with regard to
the location of sample collection as well as the timing. The data presented above clearly
demonstrate that measurements at the river mouth do not necessarily represent the maximum
levels of NO3- attained in the river system. Therefore, while end-of-river sampling can be useful
for monitoring downstream impacts, such as the potential transport of nutrients to the Great
Barrier Reef, longitudinal sampling is important for river management and hence identifying
hotspots of high nutrient availability.
7.6.2 Environmental significance
Water quality guidelines have been developed specifically for different regions and water types
within Queensland (EPA, 2006) as an extension of the National Water Quality Management
Strategy and the ANZECC8 guidelines for fresh and marine water quality in tropical Australia
(ANZECC and ARMCANZ, 2000). The regional guideline values for the defined Wet Tropics
drainage division, which includes the Herbert River catchment, summarise the trigger values for
oxidised N (NO3- + NO2
-) and ammonium N to be 30 μgN/L and 10-15 μgN/L, respectively, in
lowland streams and mid-estuarine/tidal canals in slightly-moderately disturbed systems.
Comparison of these trigger values with measurements in the lower Herbert River (Table 7-7)
indicate that concentrations of NH4+ are close to the guidelines values for aquatic ecosystem
protection. However, downstream of Nash’s Crossing (reach 1), the concentration of oxidised N
8 Australian and New Zealand Environment and Conservation Council
Chemical River-Groundwater Interactions
235
far exceeds the recommended trigger value at both the beginning and end of the dry season.
This suggests that N contributed from the lower catchment is of greater environmental concern
than upstream contributions, at least during the dry season. The concentration of NO3-, the main
form of oxidised N, is particularly high within the first 70 km downstream of the gorge.
Table 7–7 Concentration range for species of DIN for all samples collected in the lower Herbert River during selected months of the dry season. EPA and ANZECC guideline values for tropical Australia are also indicated.
Sampling time NO3- (μg/L) NO2
- (μg/L) NH4+ (μg/L)
May 2004 66 - 1549 - 3 - 10
October 2004 35 - 1709 7 - 16 8 - 12
June 2005 18 - 1062 - 8 - 22
EPA+ 133 13 - 19
ANZECC+ 44-133 13 - 19 + the oxidised N trigger value includes both NO3
- and NO2-
Importantly, longitudinal analyses highlight that the concentration of NO3- can vary
considerably along the river. Therefore, depending on where samples are collected, comparisons
against guideline values may or may not reflect the magnitude of potential impacts on
ecosystem health. For instance, as illustrated for measurements in October 2004 (Figure 7-37),
if samples were collected just below the gorge and downstream at the end of the freshwater
zone, this would not reveal that much higher concentrations are obtained in the stream in
between. As reviewed in Chapter 3 (Section 3.2.1.1), a previous study suggested that nutrient
concentrations in the Herbert River are below ANZECC (2000) target levels for the protection
of freshwater ecosystems, except in high flow conditions when the trigger values are exceeded.
Furthermore, it was concluded that nutrient loss in the catchment is event based and
insignificant outside the wet season months (Bramley and Johnson, 1996). Whilst peak wet
season events dominate the annual riverine export of nutrients (Bramley and Muller, 1999), the
results presented in this chapter clearly demonstrate that the dry season is also characterised by
elevated in-stream concentrations of NO3- within the freshwater reaches of the river.
Measurements in the estuary during the extremes of the dry season also highlight that the
concentration of NO3- is well above the Wet Tropics trigger value. This indicates the potential
for transport of high NO3- waters to the near-shore environment during the dry season, which
may also have implications for ecosystem health in the Great Barrier Reef.
Chapter 7
236
Figure 7-38 Conceptual diagram summarising the movement of water and N between the aquifers and the lower Herbert River at the end of the dry season
Chemical River-Groundwater Interactions
237
7.7 CHAPTER SUMMARY
Based on the availability of temporal water quality data in the lower Herbert River, the analyses
in this chapter have shown that there are distinct hydrochemical differences between the wet and
dry seasons, as well as between the extremes of the dry season. Although the availability of fine
resolution temporal data is limited, daily EC measurements clearly illustrate a progressive
increase in salinity of the river throughout the dry season, followed by dilution during months of
the subsequent wet season. Whilst evaporative effects are not discounted, a progressive increase
in baseflow contributions is the dominant mechanism driving the increase in river EC as flow
decreases. Hence, there is a seasonal shift from runoff-dominated to baseflow-dominated waters
contributing to the Herbert River. A decrease in the Cl/HCO3 ratio of surface waters between
the wet and dry seasons, coupled with similarities between the hydrochemical signature of
surface waters and Ca-Mg-HCO3 enriched groundwaters, provide further evidence of aquifer
discharge to the Herbert River and its tributaries during the dry season.
Consistent with the temporal analyses, longitudinal analysis of a range of parameters in the dry
season has demonstrated that groundwater discharge is the major process that influences the
chemistry of the lower Herbert River, particularly during lowest flow conditions. Although
tributary inflow is also important within some river reaches, these tributaries are supplied by
groundwater, especially at the end of the dry season. Therefore, groundwater discharge to the
river represents a combination of direct contributions from an aquifer and inflow from
groundwater-fed tributaries. Whilst there is evidence to support groundwater discharge to the
stream along the entire freshwater section throughout the dry season, the flux of groundwater
varies along the river. Considerably greater discharge is evident between approximately 60-70
km downstream from the gorge in October and June. Supported by major element
concentrations, this change in flux is due to a switch from HSd to HSs as the dominant
contributing aquifer. Whilst discharge from the deep aquifer is the main influence on river
chemistry in the upper reaches, discharge from the shallow aquifer dominates the lower reaches.
This shift in river-aquifer relationship, based on hydrochemical evidence, is consistent with
spatial variations in the incision depth of the river into the aquifers (Chapter 4), together with
baseflow estimates that indicate an enhanced contribution of baseflow between the two gauges
in the study area (Chapter 6). Whilst the groundwater flux declines towards the estuary, a
dramatic increase is estimated in the zone of seawater mixing. Although strong discharge of
groundwater is plausible in the tidal zone, there is lower confidence in the flux as mixing with
seawater masks the groundwater signal.
Uncertainty analysis of the estimated groundwater flux using radon, by comparison with a
solute mass balance, provides reasonable confidence in the estimated local baseflow
contribution along the reach characterised by the greatest change in solute concentrations.
Chapter 7
238
Furthermore, comparison of the estimated groundwater discharge between the two gauges along
the river, using a radon mass balance, is of a similar magnitude to the difference in recorded
discharge between the gauges during the same period. Improvements to measurements of the
most sensitive parameters and their uncertainties would provide greater confidence in the
estimated groundwater fluxes determined by each of the mass balance approaches. This is an
area for future work.
The analyses presented in this chapter provide strong evidence that groundwater is an important
transport vector for NO3- to the lower Herbert River during months of the dry season. NO3
- and
NH4+ dominate the river at the beginning of the dry season, whilst NO3
- and NO2- are the main
species during lowest flow conditions. These speciation differences are due to the transition
from surface-dominated to groundwater-dominated discharge to the river through the course of
the dry season. An increase in NO3- below the gorge is attributed to discharge from the deep
aquifer and inflow from groundwater-fed tributaries, while in the middle reaches of the river
NO3- is sustained by groundwater fluxes from the shallow aquifer. Although some tributaries
influence both the salinity and NO3- load of the river, others contribute to the solute load alone
with no input of NO3-. Therefore, the interpretation of non-conservative tracers such as NO3
- is
strengthened by complementary analysis of conservative tracer data. Towards the estuary, a
decline in NO3- concentration corresponds with a decrease in groundwater flux, coupled with
losses due to in-stream processes and/or denitrification. Whilst there is an estimated increase in
groundwater flux in the zone of seawater mixing, the observed concentrations of NO3- in the
river could be from groundwater or seawater sources.
Analysis of particulate and dissolved species of N has shown that dissolved forms are an
important component of total N in the stream during the dry season. Although the proportion of
DIN in the dry season is relatively minor compared to organic forms, longitudinal analysis
indicates that NO3- is present at concentrations that markedly exceed recommended guideline
values for aquatic ecosystem protection. Moreover, the analyses highlight that the concentration
of NO3- varies considerably along the river. Therefore, depending on where samples are
collected, comparisons against guideline values may or may not reflect the magnitude of
potential impacts on ecosystem health. The data clearly demonstrate that measurements at the
river mouth do not necessarily represent the maximum levels of NO3- attained in the river
system. Hence, while end-of-river sampling can be useful for monitoring downstream impacts,
such as the potential transport of nutrients to the marine environment, longitudinal sampling is
important for river management, including identification of hotspots of high nutrient
availability. Implications of the results for nutrient monitoring and management are discussed
further in the concluding chapter of the thesis.
239
Chapter 8 Research Conclusions
8.1 INTRODUCTION
The thesis commenced with a depiction of an integrated catchment management framework that
highlighted the significance of water resources as a link between land management practices
and outcomes for environmental, social and economic systems (Figure 1-1, page 2). The
management of water relates to both quality and quantity. Both of these elements can be
significantly influenced by the interactions between groundwater and surface water. This
research has focussed on the quality implications of connected water resources, motivated by
growing worldwide concern over the pollution of surface and groundwater resources by
nutrients, which can have damaging effects for both ecosystem and human health. Water is a
vehicle for mobilising dissolved constituents, including nutrients, between surface and
subsurface waters, and between terrestrial and marine systems. Therefore, an understanding of
river-aquifer connectivity is a key aspect of nutrient management in landscapes where nutrient-
bearing surface and groundwaters interact. In particular, ascertaining the significance of
groundwater fluxes for river nitrogen budgets is an important application of river-groundwater
linkages. This overarching concept has been explored and developed through the course of the
thesis.
Due to the ecological significance of the Great Barrier Reef World Heritage Area, an adjacent
coastal agricultural catchment, the Herbert River, was chosen as the target area for this research.
The selected case study catchment, located in the tropical climate zone of northeastern
Australia, also provided a unique opportunity to study groundwater contributions to the river
with minimal overland flows. Characterising the dynamics of water movement between river
and aquifer storages is a crucial step to understanding the mobility of dissolved N between
them. Accordingly, a major component of the thesis has been to investigate river-groundwater
interactions at the catchment scale, which is underpinned by characterisation of the
hydrogeological system. The role of groundwater as a vector for N is hence evaluated in light of
this conceptual understanding.
The major contributions of this thesis have been to:
(1) generate new knowledge about the Herbert River catchment, particularly in regards to
the alluvial aquifer system and surface-groundwater interactions;
Chapter 8
240
(2) establish the significance of groundwater as a vector for dissolved N to the river and
hence raise important implications for nutrient management, particularly in tropical
rivers; and
(3) demonstrate the value of combining analytical techniques, not provided by any one
method, to unfold different layers of a complex water resource problem involving both
surface and groundwater systems.
These research contributions are discussed in the following sections, with avenues for future
work also identified.
8.2 KEY FINDINGS
The analytical findings from this research are important because they provide a new
understanding of the sources of water that sustain flow in the lower Herbert River. The analyses
demonstrate that discharge of groundwater from the alluvial aquifers is a dominant influence on
both the flow and chemistry of the river in the dry season. In particular, groundwater is a key
vector for the delivery of nitrate (NO3-) to the river during low flow conditions. This provides a
new perspective for monitoring and management of nutrients in tropical rivers where there is
good connectivity with the underlying groundwater system.
The main analytical components of the thesis include:
(1) characterising river-aquifer interactions; and
(2) ascertaining the role of groundwater for the nitrogen budget of river systems.
A major element of (1) was also to conceptualise the hydrogeological system. The key
analytical findings from each of these components are summarised below.
8.2.1 Hydrogeological framework
The alluvial sequence of the lower Herbert River catchment is conceptualised as a two-aquifer
system comprised of a shallow unconfined aquifer (HSs) and semi-confined deep aquifer (HSd).
Although the aquifers are distinct, the extent of vertical hydraulic connection varies spatially.
The western half of the alluvial system is characterised by good/strong downwards connectivity.
In contrast, poor inter-aquifer connection is evident in the east, except towards the coast, where
strong upward heads and good connectivity results in vertical discharge from HSd to HSs.
Groundwater in each aquifer flows from the upland lateral recharge areas in the northwest and
southwest towards the lower Herbert River, and then eastward in the direction of the coastal
discharge zone. In addition to lateral sources, diffuse rainfall is an important component of
recharge to HSs.
Research Conclusions
241
Shallow groundwaters are dominated by Na-HCO3-Cl facies, with both Ca-Mg enriched and Na
enriched groups relating to clay content of the overlying soils. The exception is towards the
coast, where Na-Cl ± HCO3 facies groundwaters are associated with contributions from the deep
aquifer and mixing with the Herbert River estuary. The western half of the deep aquifer is
dominated by Na-HCO3 and Na-HCO3-Cl facies waters. Lateral hydrochemical evolution is
suppressed by enhanced vertical leakage from the shallow aquifer and the convergence of
flowpaths contributing groundwater of different compositions. The eastern half of the deep
aquifer is represented by Na-Cl facies waters with high salinities, influenced by seawater from
past and/or present-day intrusion via a preferential pathway in the northeast.
Despite good vertical connectivity between the aquifers in some areas, dissolved inorganic
forms of N (DIN) are very different between aquifers due to redox controls. Therefore, whilst
oxidising conditions in the main recharge area favour NO3-, the concentration of DIN in the
deep aquifer dramatically declines down the flowpath due to N reduction processes. Hence,
elevated NO3- in HSd is restricted to the high country in the northwest, the main source region
for DIN leached to the deep aquifer. In contrast, relatively high concentrations of NO3- are
found throughout the shallow aquifer, with the distribution and speciation of DIN influenced by
aquifer composition, including soil type and redox state. Oxidising Ca-Mg-enriched waters are
associated with high concentrations of NO3-, while more reduced Na-enriched waters are
dominated by NH4+. Importantly, high concentrations of NO3
- are present in shallow and deep
groundwaters adjacent to the Herbert River. Therefore, given the hydrogeological framework,
there is potential for subsurface N to contribute to the river system. Ultimate discharge of
groundwater from both aquifers to the sea indicates an additional pathway for the movement of
N offshore. A conceptual diagram summarising water and N movement in the alluvial aquifer
system was presented in Chapter 5 (Figure 5-33). This figure is repeated below as Figure 8-1.
8.2.2 River-groundwater interactions
Groundwater discharge is the major process that influences the chemistry of the lower Herbert
River during the course of the dry season, particularly during the lowest streamflow conditions.
Whilst other processes such as evaporation are not discounted, a progressive increase in
baseflow contributions is the dominant mechanism that drives an increase in river salinity as
flow decreases. Hence, there is a seasonal shift from runoff-dominated to baseflow-dominated
flow in the river. Tributaries are also an important source of water and dissolved constituents
along some reaches; however, many of these tributaries are supplied by groundwater, especially
at the end of the dry season. Therefore, groundwater discharge to the river in the dry season
represents a combination of direct and indirect (via tributary inflow) aquifer contributions.
Chapter 8
242
Year-round, the dominant potential direction of groundwater flow is from the aquifers to the
river; however, towards the coast, the direction of potential flux reverses during short periods of
high river stage in the wet season. Hence, the upper reaches of the river are dominantly gaining,
whilst the lower reaches are characterised as variably gaining/losing. Whilst groundwater
discharges to the river along the entire freshwater section throughout the dry season, the flux is
variable along the river. Considerably greater flux is evident between approximately 60-70 km
downstream from the gorge in October and June, attributed to a transition from HSd to HSs as
the dominant contributing aquifer. Hence, discharge from the deep aquifer is the main influence
on river chemistry in the upper reaches, while discharge from the shallow aquifer dominates the
lower reaches. Groundwater inflow to the river declines towards the estuary, but with a dramatic
increase estimated in the zone of seawater mixing. Although strong discharge of groundwater in
the tidal zone is plausible, there is lower confidence in the flux, as mixing with seawater masks
the groundwater signal.
Uncertainty analysis, using independent mass balances of radon, and solute loads, provides
reasonable confidence in the estimated local baseflow contribution along the freshwater reach
characterised by the greatest change in ionic concentrations. In addition, the estimated discharge
between the two gauges along the river is consistent with the difference in gauged discharge
during the same period. Improvements to measurements of the most sensitive parameters and
their uncertainties would provide greater confidence in the estimated groundwater fluxes
determined by each of the mass balance approaches.
8.2.3 The significance of groundwater for river N budgets
This research demonstrates that groundwater is an important transport vector for DIN to the
lower Herbert River during months of the dry season. Although NH4+ and NO2
- also comprise
the inorganic component, NO3- is by far the dominant species throughout the dry season. An
increase in NO3- below the Herbert River gorge is attributed to discharge from the deep aquifer
and inflow from groundwater-fed tributaries, whilst in the middle reaches of the river, NO3- is
sustained by groundwater fluxes from the shallow aquifer. Importantly, whilst some tributaries
influence both the salinity and NO3- load of the river, others contribute to the solute load alone
with no input of NO3-. Towards the estuary, a decline in NO3
- corresponds with a decrease in
groundwater flux, coupled with losses due to in-stream processes such as denitrification. Whilst
there is an estimated increase in groundwater flux in the zone of seawater mixing, the observed
concentrations of NO3- in the river could be from groundwater or seawater sources. A
conceptual diagram summarising the movement of water and N between the aquifers and the
lower Herbert River is presented below in Figure 8-2 (equivalent to Figure 7-38, Chapter 7).
Figure 8-1 Conceptual diagram summarising the movement of water and N in the alluvial aquifer system and potentially to the Herbert River (HR)
Chapter 8
244
Figure 8-2 Conceptual diagram summarising the movement of water and N between the aquifers and the lower Herbert River at the end of the dry season
Research Conclusions
245
The N budget of the lower Herbert River varies both temporally and spatially. Particulate forms
dominate in the wet season, whilst dissolved forms comprise 80-90% of total N during months
of the dry season. Although the proportion of DIN in the dry season is relatively minor
compared to organic forms, the concentration of NO3-, particularly within 70 km downstream of
the gorge, markedly exceeds the recommended Wet Tropics and ANZECC guideline values for
aquatic ecosystem protection. An important result arising from this research is that the
concentration of NO3- is highly variable along the river during the dry season. Measurements
immediately below the gorge and towards the river mouth do not reflect the much higher
concentrations of NO3- present in the intervening reaches. Therefore, water quality data must be
interpreted with regard to the location of sample collection as well as the timing. This is of
particular importance when comparing water quality indicators such as NO3- against trigger
values to assess the magnitude of potential ecological impacts. Hence, whilst end-of-river
sampling can be useful for monitoring downstream impacts, such as the potential transport of
nutrients to the Great Barrier Reef, longitudinal sampling is important for river management,
including identification of hotspots of high nutrient availability.
8.3 RESEARCH APPROACH
The key findings were determined through an integrated research approach comprising a data
collection component and an assessment methodology. These elements were guided by the
fundamental research questions, together with regard to existing available information,
resources to collect additional data, and the specific attributes of the case study catchment. The
water quality sampling program took advantage of the distinct seasonal nature of tropical
catchments by sampling during the period characterised by minimal overland flows. Hence the
pure baseflow component in the river has been captured with reasonable confidence. The
analytical approach applied in this thesis clearly demonstrates that methods traditionally used by
separate disciplines, both qualitative and quantitative, can be combined to provide a powerful
toolbox of complementary techniques that enrich the conceptual process understanding of the
system.
8.3.1 Data collection
In order to assess large scale river-groundwater interactions and the application to nutrient
transport, a catchment scale investigation was undertaken. Given the broad scale of the study,
spatial relationships, rather than temporal dynamics, were considered to be of greatest interest.
Hydrological processes in the tropics are generally less dynamic during the dry season, as
overland flow is minimal and baseflow dominates streamflow. Accordingly, an extensive water
quality sampling program was instigated in the case study catchment during low flow
conditions. Grab samples of groundwater and surface waters at the beginning and end of the dry
Chapter 8
246
season were considered to be adequate for comparing differences in river-aquifer connectivity
relationships between the extremes of the dry season at a catchment scale. The results highlight
that the end of the dry season is particularly useful for isolating the groundwater signal in the
river.
A key element of the research approach was the use of physical datasets sourced primarily from
existing government department records. These provide a backbone for understanding the
hydrogeology and hydrology in the case study catchment. Whilst it is noted that these datasets
generally have gaps in historical records and issues with data quality, they are nonetheless
considered to be a valuable source of high temporal resolution information, at least during
particular monitoring periods. The collection of time series data was not feasible within the
scope and resources of the project. Existing datasets were therefore an important supplement to
the hydrochemical data collected specifically for this investigation. As this research
demonstrates, verification and extension of physical concepts can be accomplished through
analysis of water quality data. The extensive database of hydrochemical information is of
reasonably high spatial resolution and low temporal resolution, thus complementing the
physical datasets.
8.3.2 Characterising river-aquifer interactions
A combination of hydrogeological, hydrometric, hydrological and hydrochemical techniques
were applied in the thesis to characterise the interaction between the alluvial aquifers and the
lower Herbert River. An important aspect of establishing the relationships between surface and
subsurface waters was to build a conceptual understanding of the hydrogeological system. This
included:
(1) classifying distinct aquifers based on lithostratigraphic interpretation and assessment of
vertical hydraulic behaviour;
(2) establishing lateral groundwater flow patterns in each aquifer; and
(3) identifying aquifer recharge and discharge zones and potential interactions with the
Herbert River
Whilst the analysis of physical attributes alone enabled a reasonable framework for the
groundwater system to be developed, the complementary analysis of hydrogeochemical data
allowed verification and enhancement of the conceptual model. For example, bore hydrograph
analysis indicated the potential direction and degree of vertical flow between aquifers, while the
analysis of hydrogeochemical data provided evidence of actual exchange of water and the
relative extent of inter-aquifer mixing. Furthermore, hydrogeochemical analysis highlighted the
importance of vertical recharge processes over lateral hydrochemical evolution in each aquifer,
Research Conclusions
247
which could not be inferred from the available hydraulic data. Conversely, the interpretation of
hydrogeochemical information would have limited scope in the absence of a physical process
understanding of groundwater flow. The analyses presented in Chapters 4 and 5 demonstrated
that the hydrogeological conceptualisation was strengthened by the coupling of physical and
chemical datasets and the corresponding hydrogeological and hydrochemical assessment
techniques.
Similar to the approach for characterising the hydrogeological system, a combination of
physical-and chemical-based methods were utilised in Chapters 6 and 7 to examine the
connectivity between groundwater and the lower Herbert River. The potential for hydraulic
connection and the direction of flux between the aquifer system and the river were evaluated
through qualitative hydrometric approaches, including:
(1) depth relationships of the river channel with that of the underlying alluvial sediments;
(2) historical groundwater elevation-stream stage relationships; and
(3) groundwater flow patterns around the river
Hydrological techniques such as the analysis of (4) stream hydrographs and (5) flow duration
curves were also applied to determine the temporal characteristics of flow in the river, while the
groundwater flux to the river was quantified during the wet and dry seasons by (6) hydrograph
separation. Hence, the hydrometric methods provided an indication of the theoretical direction
of interaction along particular reaches of the river, whilst examination of flow characteristics
provided evidence of actual volumetric flux over a broad area.
The physical understanding of river-aquifer linkages was verified and enriched through the
assessment of surface water chemistry data, in conjunction with the conceptual hydrogeological
model incorporating the hydrochemical signature of groundwater. Surface water integrates
across the entire catchment as it is the cumulative result of hydrological processes upstream of
the sampling point. Therefore, water quality samples analysed along the river reflect upstream
inputs, such as the discharge of groundwater. The utility of different environmental tracers
(electrical conductivity, temperature, major ions, radon, stable isotopes) for distinguishing
between the key processes that influence river chemistry was illustrated, including the
application of radon as a tracer of groundwater discharge to the river. The analyses
demonstrated how the combination of tracers, at varied spatial and temporal resolution, can
provide a powerful toolbox for characterising the chemistry of the river in space and time.
Consistency between tracers provided confidence that the observed longitudinal trends were
important for identifying relative changes in water sources to the river, even though the
magnitude of hydrochemical variation was only minor for some dissolved components.
Chapter 8
248
Furthermore, the spectrum of tracer techniques yielded both qualitative and quantitative
information regarding the flux of groundwater along the length of the lower Herbert River.
Whilst the absolute groundwater fluxes determined have a degree of uncertainty, the various
mass balance approaches using radon and solute loads highlighted the general methodology and
efficacy of quantitative estimates in combination with visual qualitative trends to characterise
river-aquifer relationships.
8.3.3 Characterising nutrient mobility
As an application of river-aquifer interactions, the potential for movement of N from the
aquifers to the river was assessed. Approaches for characterising nutrient mobility range from
qualitative field-based studies to quantitative nutrient modelling. Given the focus on processes
and the extensive data requirements which could not be met with existing data, a modelling
approach was considered beyond the scope of the thesis. However, it is noted that the
interpretation of model outputs can be enhanced with process knowledge of key components
such as physical and chemical river-aquifer linkages. The field-based hydrochemical approach
for characterising the movement of N between surface water and groundwater comprised three
key components, involving:
(1) examination of the spatial distribution and speciation of DIN in the alluvial aquifers;
(2) appraisal of river-groundwater interactions and hence whether transport of DIN to
surface waters via groundwater is a plausible mechanism; and
(3) assessment of whether there is hydrochemical evidence of N transport to the lower
Herbert River via this mechanism
Investigating the significance of groundwater N fluxes to the river was founded therefore on an
understanding of the hydrogeology and the dynamics of water movement between the river and
alluvial aquifers. Furthermore, given that NO3-, the dominant form of DIN in the river, is non-
conservative under reduced chemical conditions, the interpretation of NO3- data was
strengthened by complementary analysis of conservative tracer data such as major ions and
solute loads. In addition, comparison of NO3- data with radon was a key diagnostic for a
groundwater source of NO3-. The hydrochemical analysis thus underlines the importance of a
multi-tracer approach for verifying and enriching a conceptual process-based model.
Research Conclusions
249
8.4 IMPLICATIONS FOR THE USE OF NUTRIENT BUDGETS AND MODELS
This thesis illustrates a methodology for undertaking a catchment scale study concerned with
characterising river-groundwater interactions and evaluating the significance of these
interactions for river nutrient budgets. The analyses have provided an understanding of the
dynamics of water movement between the alluvial aquifers and the lower Herbert River during
the dry season and shown that groundwater is a dominant source and vector for the transport of
NO3- to the river. The results highlight that there is considerable spatial variability in the system.
For example, the sources and fluxes of groundwater vary along the river, as do the
concentrations of NO3- in the groundwater and measured in-stream. In the absence of the multi-
tracer approach presented in the thesis, this variability would be difficult to identify in a nutrient
budget or other quantitative modelling approaches. Therefore, the results demonstrate the value
of on-ground research in order to gain process understanding, which could inform the inputs
into and/or design of a nutrient model and hence be used for nutrient management purposes.
However, it is acknowledged that resource constraints may inhibit comprehensive data
collection for a particular study. Hence, an important consideration arising from the thesis is the
minimal dataset requirements to undertake a meaningful nutrient budget or modelling approach,
in a system with high river-aquifer connectivity.
Based on the range of analytical techniques applied in this thesis, a basic understanding of river-
groundwater interactions can be gained through baseflow filtering and qualitative hydrometric
methods such as assessment of bore and stream hydrographs. Coupled with these tools,
literature values can be used as an indicator of the typical concentrations of DIN expected in
surface waters that drain catchments under a particular land use. Although these desktop
analytical methods are relatively simple, they rely on reasonably high temporal resolution river
and/or groundwater level monitoring data for examining connectivity. Furthermore, surrogate
literature values of DIN fail to provide information on the spatial variability of nutrient
concentrations found in the stream. In the absence of such datasets or appropriate monitoring
networks, simple and relatively inexpensive hydrochemical data can be collected in surface
waters, such as for the analysis of radon and NO3-. Radon data, coupled with other
measurements (depth, width, discharge) can identify sites of groundwater discharge and
potential changes in the flux along the river, while NO3- samples can indicate longitudinal
variability and the actual concentrations of NO3- reaching the stream. Ideally, an understanding
of the hydrogeological system would be necessary to capture the variability and complexity of
river-groundwater interactions. Nonetheless, for the purposes of a broad scale nitrogen budget
or other nitrogen models, a combination of simple desktop tools and targeted water quality
sampling are the minimum requirements. It is noted however, that the reliability of nutrient
model outputs and their interpretation in a connected river-aquifer system can be enhanced with
Chapter 8
250
sound process understanding of the physical and chemical relationships between surface water
and groundwater resources.
8.5 MONITORING AND MANAGEMENT IMPLICATIONS FOR THE TROPICS
This research raises important implications for nutrient monitoring, management and policy,
particularly in tropical rivers. A defining characteristic of tropical systems is the marked
seasonality of surface water flows, driven by the distinct wet season-dry season rainfall patterns.
Accordingly, there is a shift in the dominant sources of water contributing to the lower Herbert
River. As a consequence of the shift from overland flow to baseflow dominance, the chemistry
of the river changes, including with respect to dissolved inorganic forms of N. In general, N
monitoring and management of tropical rivers has traditionally focussed on high intensity flow
events that result in high total N loads in surface waters, especially of particulate forms.
However, the results of this thesis demonstrate that significant concentrations of dissolved N are
also present in surface waters of the Herbert River throughout the dry season, due to N inputs
from groundwater sources. Even though wet season flows dissipate rapidly, longer residence
times of water during low flow conditions provide greater opportunity for nutrient contributions
to impact on river health. In addition, the analyses highlight that the concentration of DIN varies
within the alluvial aquifers contributing NO3- and that there are hotspots of high nutrient
availability in the groundwater system. The results clearly illustrate that there is considerable
variation in measured concentrations of DIN, particularly NO3-, along the entire length of the
lower Herbert River. Importantly, NO3- concentrations along particular reaches of the river are
at potentially harmful levels for the health of riverine ecosystems.
The results from this research in the lower Herbert River catchment, combined with the general
characteristics of tropical rivers, underline the following key points and recommendations in
regards to nutrient monitoring, management and policy guidelines in connected river-aquifer
systems within the tropics.
• Water quality sampling should be undertaken at recognised periods on the
stream/groundwater hydrograph, with an understanding of temporal and spatial river-
aquifer connectivity relationships. Sampling during low flow conditions, particularly at
the end of the dry season, is the most effective time for isolating the groundwater signal
and hence evaluating potential nutrient contributions in gaining stream situations.
• Surface and subsurface sources of water and dissolved nutrients must be considered.
Hence, nutrient hotpots in both surface water and groundwater systems should be
identified and targeted for land management actions.
Research Conclusions
251
• Sampling location, as well as timing, should be carefully chosen. End-of-river
monitoring does not necessarily capture the maximum nutrient concentrations along
upstream reaches and hence may not reflect the magnitude of potential impacts on
ecosystem health. Longitudinal sampling is considered to be of greatest relevance for
assessing river health.
• Appropriate water quality guideline values must be set that account for seasonal
changes in both the sources that deliver nutrients and the forms of N transported to
surface waters. Single water quality targets are not appropriate in distinctly seasonal
catchments.
8.6 FURTHER RESEARCH
Whilst the aims of the thesis were adequately addressed, further work would improve spatial
and temporal process understanding of the dynamics of water and N movement in the river-
aquifer system of the Herbert River catchment. Furthermore, this thesis provides a platform for
undertaking a catchment-scale nutrient budget or other modelling approach, which could
potentially be applied to other case study areas.
This research identifies a number of areas for future work:
(1) Improvements to the hydrogeological framework
This could be achieved by:
• determining aquifer hydraulic properties over a greater spatial extent, both laterally and
vertically, including measurements in multiple sandy units;
• measuring groundwater levels and groundwater quality at close proximity to the river
during the wet and dry seasons;
• analysing deep aquifer samples for isotopes such as 36Cl, SF6 and CFC’s to estimate
groundwater residence time; and
• strategic development of additional monitoring bores coupled with further enhanced
surveying techniques (e.g. ground and airborne geophysics).
Chapter 8
252
(2) Improvements to the spatial and temporal understanding of river-groundwater fluxes
Refinements could include:
• collection of stream hydrograph data at multiple gauges along the river;
• measurement of groundwater elevations close to the river and at a greater number of
sites adjacent to and within the river;
• installation of automatic piezometers in the river and along transects away from the
river, as well as temperature and salinity probes;
• repetition of radon sampling in the river and its tributaries at the beginning, end and part
way through the dry season, including during different tidal cycles in the estuary; and
• accurate determination of river parameters such as river depth, width and stream
discharge at all radon sampling sites, as well as increasing the spatial extent of radon
measurements in the aquifers. These improvements could provide greater confidence in
groundwater flux estimates.
(3) Improvements to inputs of a N mass balance
Not all forms of N were measured in surface water and groundwater in this study. Therefore, to
complement the existing data, additional samples could be collected at the beginning and end of
the dry season for analysis of:
• gaseous forms of N such as nitrous oxide and nitrogen gas and dissolved organic N in
the river and in groundwater proximal to the river;
• particulate N in surface waters; and
• dissolved and particulate N in adjacent seawater.
In addition, further work could include:
• examination of N transformations in groundwater through N isotopes and other redox-
sensitive tracers;
• geochemical modelling to identify where changes in groundwater chemistry can
account for changes in N; and
• improvement to the quantification of water movement between the aquifers and the
river in order to quantify the flux of N between them.
Research Conclusions
253
8.7 CONCLUDING STATEMENT
The results from this thesis, based on a case study catchment in the tropics of northeastern
Australia, demonstrate that groundwater can be an important vector for the delivery of dissolved
forms of nitrogen to surface waters during low flow conditions. Therefore, characterising the
interaction between surface water and groundwater resources, through a combination of
complementary analytical techniques, is crucial for effective management of water quality. It is
hoped that this research also provides a new perspective for monitoring and management of
nutrients, particularly in tropical river systems.
255
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Appendix A Laboratory analyses and field data
Raw data collected for this research is provided electronically in a spreadsheet labelled
Herbert_River.xls (CD in back pocket of the thesis). GPS location data for surface water and
groundwater samples are in map datum GDA94. Other units are as specified in the spreadsheet.
273
Appendix B Radon sampling procedure
The procedure used for sampling radon in this study is a modification of the method described
in Cook et al. (2004). Outlined below are the key steps (adapted from the method of F. Leaney
2004, pers. comm.).
(1) Fill a 1250 mL plastic bottle with the required river/groundwater sample, ensuring minimal
air bubbles. To minimise agitation, attach a short length of clear tubing to the submersible
pump; insert the tubing into the collection bottle and allow it to gradually fill (at a low flow
rate) and overflow prior to capping. Record the time and location of sampling; for river samples
also record the width and depth of the stream.
(2) To extract radon from the river/groundwater sample, remove the lid of the bottle and syringe
out 50 ml of water to discard. Pour approximately 20 mL of scintillant9 from a pre-weighed vial
into the bottle and recap.
(3) Mix the water and scintillant by inverting the bottle every 2 seconds for a period of 4
minutes. Allow the bottle to stand for one minute to allow the scintillant to move to the top.
(4) Remove the lid from the bottle and insert a glass nozzle onto the top of the bottle.
(5) Remove the lid from the empty scintillant vial (from step 2) and hold at the outlet of the
glass nozzle. Slowly squeeze the bottle to push the scintillant out of the bottle into the vial. Stop
squeezing when the scintillant/water interface reaches the start of the capillary tubing at the start
of the nozzle, avoiding water from being transferred to the vial.
(6) Put the cap firmly back on the scintillant vial and tighten using pliers. Record the sampling
location, time and corresponding scintillant vial number for sending to the laboratory.
(7) Rinse the flask and nozzle with methylated spirits and dry in preparation for the next sample.
(8) Courier bottles to the laboratory within 2-3 days after sampling for radon counting.
9 Scintillant vial contains Packard NEN mineral oil cocktail
275
Appendix C Uncertainty analysis
In equation 7-13, the uncertainty in xbest was expressed as:
2/1
1
−
⎟⎠
⎞⎜⎝
⎛=Δ ∑
N
ibest wx
The derivation of this is as follows:
∑∑=Δ
i
iibest w
xwx (7-12)
i.e. }{ 2211 )(.....)(1
NNi
best xwxww
x Δ++Δ=Δ∑
i.e. ⎪⎩
⎪⎨⎧
⎭⎬⎫
Δ++
Δ=Δ∑
22
1
)1(.....)1(1
Nibest xxw
x where 2
1
ii x
wΔ
=
Therefore,
}{ ∑∑=Δ i
ibest w
wx 1
or 2/1
1
−
⎟⎠
⎞⎜⎝
⎛=Δ ∑
N
ibest wx as in equation 7-13.
