National Centre for. Groundwater A RE-EVALUATION OF GROUNDWATER DISCHARGE FROM THE BURDEKIN FLOODPLAIN AQUIFER USING GEOCHEMICAL TRACERS

Transcription

1 National Centre for Groundwater Research and Training Australian Research Council A RE-EVALUATION OF GROUNDWATER DISCHARGE FROM THE BURDEKIN FLOODPLAIN AQUIFER USING GEOCHEMICAL TRACERS December 2011 P.G. Cook1,2, S. Lamontagne1,2, T. Stieglitz3, R. Cranswick1 and G. Hancock4 National Centre for Groundwater Research and Training, Flinders University, Adelaide CSIRO Land and Water, Adelaide 3 JamesCook University, Townsville 4 CSIRO Land and Water, Canberra 1 2

2 Published by: Authors: Copyright: Disclaimer: The National Centre for Groundwater Research and Training GPO Box 2100 Adelaide SA 5000 PG Cook, S Lamontagne, T Stieglitz, R Cranswick and G Hancock. The information contained in this document is the property of NCGRT. Use or copying of this document in whole or part without the written permission of NCGRT constitutes an infringement of copyright. NCGRT advises that the information contained in this publication comprises general statements based on scientific research. The reader is advised and needs to be aware that such information may be incomplete or unable to be used in any specific situation. No reliance or actions must therefore be made on that information without seeking prior expert professional, scientific and technical advice. To the extent permitted by law, NCGRT (including its employees and consultants) excludes all liability to any person for any consequences, including but not limited to all losses, damages, costs, expenses and any other compensation, arising directly or indirectly from using this publication (in part or in whole) and any information or materials contained in it. Citation: For bibliographic purposed this report may be cited as: Cook, PG, Lamontagne, S, Stieglitz, T, Cranswick, R and Hancock, G 2011, A Re- Evaluation Of Groundwater Discharge From The Burdekin Floodplain Aquifer Using Geochemical Tracers, National Centre for Groundwater Research and Training, Australia.

3 EXECUTIVE SUMMARY High water tables following an extensive wet season provided a unique opportunity to evaluate groundwater baseflow to rivers and coastal zones from the Burdekin River Delta aquifer. This region is one of the largest irrigation districts in Australia and relies, in part, on groundwater extraction. However, because of the complex nature of the aquifer, it is difficult to evaluate groundwater discharge to rivers and coastal zones using traditional hydraulic techniques. In , Cook et al. (2004) evaluated groundwater baseflow to rivers in the Burdekin region using radon. In addition, they evaluated submarine groundwater discharge to Bowling Green Bay, an important outlet for the Burdekin aquifer, using radon and the radium quartet ( 223 Ra, 224 Ra, 226 Ra, 228 Ra). Radon and radium are naturally occurring radioisotopes that tend to be enriched in groundwater relative to receiving surface waters. This property makes them ideal tracers to quantify groundwater discharge. In this study, the sampling survey by Cook et al. (2004) was repeated to compare groundwater discharge to surface water and the marine environment from the Burdekin aquifer at the end of the 2004 and 2011 wet seasons. This included an extensive field campaign from 9 12 May 2011 to collect longitudinal river profiles for radon from the Burdekin and Haughton rivers and from Plantation and Barratta creeks. In addition, radon and radium surveys were conducted in Bowling Green Bay. An effort was made to better characterize the radon and radium signatures for key water sources, such as radon activity in hyporheic water, radium activity at river mouths, and radium activity in the more saline sections of the Burdekin aquifer. Building on knowledge gained since the 2004 studies, the radon and the radium data from 2004 and 2011 were interpreted with modified models. For rivers, the potential contribution of hyporheic exchange to the river radon budget was quantified. For Bowling Green Bay, a different model was used that explicitly incorporated the effects of changes in bathymetry in the bay. Riverine groundwater baseflow The river sampling campaign was repeated between 9 12 May In the Burdekin River, radon activity in May 2004 and May 2011 were similar, generally ranging from 200 to 500 mbq/l. This contrasted with river samples collected in December 2003 and September 2004, which were generally below 200 mbq/l. This pattern is consistent with greater groundwater discharge at the end of the wet season. River EC was highest in May 2011 relative to other sampling periods, also consistent with high groundwater baseflow at that time. In addition, zones with elevated groundwater discharge along the banks were evident in May 2011 and had more elevated radon (>3000 mbq/l) and EC relative to well-mixed sections of the river. The groundwater inflow over the 62 km of river downstream from Clare Weir was estimated at 248,000 m 3 /day. In the Haughton River, river radon activities were larger in May 2011 than May In the section km upstream from the mouth, radon was mbq/l in May 2011 and <250 mbq/l in May Groundwater discharge was estimated to be larger in May 2011 (~138,000 m 3 /day) than in May 2004 (15,500 m 3 /day). In the Barratta Creek, radon activities in May 2011 were variable ( mbq/l) but substantially higher than in 2004 (<500 mbq/l). Like in the Burdekin, in May 2011 areas of groundwater discharge along some banks were evident and had 3

4 elevated radon activities. Groundwater inflow for the 60 km reach above the mouth was estimated at 55,900 m 3 /day. In Plantation Creek, radon activity and EC were generally two-fold higher in May 2011 than in May However, because of low sampling resolution, no estimates of groundwater discharge were made for this river. Submarine groundwater discharge to Bowling Green Bay In Bowling Green Bay, 19 radium samples were collected along three transects perpendicular to the shoreline on May Seven river mouth radium samples were collected at the Haughton River and Barratta Creek during the same period. Point radon samples were collected at the river mouths together with radium, and also at seven additional shoreline locations. In addition, radon was measured continuously along the radium transects and also for two additional transects parallel to the Bowling Green Bay shoreline. In general, radon and radium activity followed the expected trend of higher activities near the shoreline declining exponentially in the offshore direction. The short-lived radium isotopes ( 223 Ra and 224 Ra) were used to estimate offshore diffusivity (the tendency for solute to move away from the shoreline). These estimates of offshore diffusivity were then used to estimate the offshore flux for the long-lived radium isotopes and for radon. The total radon and radium loads were remarkably similar in 2004 and For example, the average 222 Rn loads were 84,000 and 81,000 Bq/s in 2004 and 2011, respectively; well within measurement error. This similarity in the offshore flux for Ra and Rn between years indicates either that submarine groundwater discharge from the Burdekin aquifer is constant or that it is small relative to other sources of Rn and Ra to Bowling Green Bay. It is conceivable that groundwater discharge is constant from year to year at the coastline because there is little scope for the water table to rise as it is already close to the land surface. In other words, when the water table rises near the coastline groundwater will tend to be discharged as surface flow to the river network. A radium and radon mass-balance was calculated for Bowling Green Bay to estimate groundwater discharge from the Burdekin aquifer (Q gw ). This involved using estimates for the likely magnitude of the recirculated seawater flux combined with measurements of radon and radium activity in groundwater, recirculated seawater and estuarine water. In both 2004 and 2011, total submarine groundwater discharge (including recirculated seawater and groundwater discharge from the Burdekin aquifer) was a much larger source of radon or radium to Bowling Green Bay than inflowing rivers. In May 2011, Q gw ranged from , and ML/day based on 228 Ra, 226 Ra and 222 Rn, respectively. These values are in the same range as estimates of river discharge to Bowling Green Bay at the time of the study (1300 ML/day). In April 2004, Q gw was generally within the same range as in 2011 for each isotope. However, river discharge was lower in 2004 (140 ML/day). The key limitation for these estimates of groundwater discharge from the Burdekin aquifer is that the recirculation flux (the largest source of radon and radium to Bowling Green Bay) could only be estimated based on other studies. Due to some additional uncertainties with the radium-based fluxes, the Q gw estimated from the radon budgets are probably more reliable than those derived from radium. The Q gw estimated in April 2004 and May 2011 for Bowling Green Bay are probably at the higher range on a seasonal basis because they both represent the end of the wet season when regional water tables tend to be highest. 4

5 Combined baseflow estimates Using the mid-range estimates from radon, the relative proportion of groundwater discharging as river baseflow or submarine discharge was different between 2004 and Groundwater discharge to rivers varied from 40 to 440 ML/day in 2004 and 2011, respectively, whereas groundwater discharge to Bowling Green Bay remained relatively constant ( ML/day). In other words, most of the groundwater discharge was as submarine groundwater in 2004 and as river baseflow in When extrapolated on an annual basis, the combined estimates of groundwater discharge represent a significant component of the Burdekin aquifer water balance, varying from 124,000 ML/year in 2004 to 260,000 ML/year in In contrast, recharge to the Burdekin aquifer has been estimated to vary from 430,000 to 850,000 ML/year and groundwater pumping from 440,000 to 830,000 ML/year. However, the true groundwater discharge is probably lower than indicated here because the measurements are only representative of the end of the wet season, when groundwater discharge should be highest. Conclusions Radon activities were substantially higher in rivers but not in the coastal zone in the Burdekin region in May 2011 relative to the same period in 2004; Groundwater discharge to rivers was higher in May 2011 relative to the same period in 2004 but groundwater discharge to the coastline was similar between years; Total groundwater discharge to the Haughton, Barratta and Burdekin rivers and directly to Bowling Green Bay was estimated to be 340 in May 2004 and 700 ML/day in May

6 Table of Contents EXECUTIVE SUMMARY 3 1. INTRODUCTION 7 2. METHODS General River Flows and Aquifer Hydraulics Surface Water and Groundwater Sampling Sample collection in Bowling Green Bay Radon Analyses Radium Analyses DISCHARGE TO STREAMS Introduction Theory Radon Activity in Groundwater Burdekin River Haughton River Barratta Creek Plantation Creek DISCHARGE TO THE OCEAN Introduction Theory Results Discussion DISCUSSION AND CONCLUSIONS Discharge to surface waters Submarine Groundwater Discharge Conclusions Recommendations REFERENCES ACKNOWLEDGEMENTS 82 APPENDIX 1: Effect of Sampling Distance on Inflow Uncertainty 84 APPENDIX 2: Results of Geochemical Analyses 86 APPENDIX 3: Comparison of surface water gauge heights and groundwater elevations. 92 6

7 1. INTRODUCTION The Burdekin River Delta covers an area of approximately 850 km 2, and represents one of the largest unconfined coastal aquifer systems in eastern Australia (McMahon et al., 2002). The Burdekin River, Haughton River, Barratta Creek and a number of minor watercourses drain the delta and surrounding floodplain area, which together cover an area of approximately 2760 km 2. The Burdekin River has a mean annual flow volume of approximately 9.3 million ML, almost 80% of which occurs between January and March. The mean annual flow volume of the Haughton River is ML, and for Barratta Creek it is ML. Flow of the Burdekin River is regulated at Burdekin Falls Dam, and at Clare Weir. Two weirs have also been constructed on the Haughton River to provide storages for irrigation water, and the river is also supplemented by water pumped from the Burdekin River. There are no weirs on Barratta Creek. Both surface water and groundwater are extensively used for irrigation within the floodplain region. Currently more than 350 km 2 of land is irrigated, with sugarcane the major crop. Recharge to the unconfined aquifers of the Burdekin River floodplain is from a combination of natural and artificial processes, including natural infiltration of rainfall, leakage from beds and banks of the Burdekin River and other watercourses, overland floods, inflow from bedrock and adjacent areas, irrigation return flows and artificial recharge pits and channels. McMahon et al. (2002) estimated total recharge from these various sources for the Burdekin River Delta to be between and ML/year. Groundwater pumping is estimated to be between and ML/year. Because of large uncertainty surrounding estimates of groundwater extraction and aquifer recharge, estimation of the volume of groundwater discharge to streams and to the ocean from a water balance is difficult. Recently, Kalbus et al. (2006) reviewed the various methods available for quantifying surface water groundwater exchange. The available methods operate over a range of temporal and spatial scales. Many methods provide only point measurements of the exchange flux (e.g., seepage meters), which, although important for understanding processes, are difficult to upscale due to spatial variability of streambed properties (Calver, 2001; Kennedy et al., 2008). The direction of groundwater surface water flow (i.e. distinguishing losing from gaining streams) can usually be obtained from a comparison of surface water levels and hydraulic heads in piezometers adjacent to streams. However, it is often difficult to accurately quantify the flow rate from this data because hydraulic conductivity is difficult to characterise on this scale. Environmental tracer methods have been used to quantify groundwater discharge to rivers for the past few decades. They offer advantages over physically-based methods in that they can potentially provide more accurate information on the spatial distribution of groundwater inflows over large regions with a much lower investment of resources. An environmental tracer is useful for estimating groundwater inflows to rivers when the concentration of the tracer in groundwater is relatively uniform and significantly different to that in the river. One of the most powerful tracers for this purpose is radon ( 222 Rn). With a half-life of 3.8 days, radon is produced in the subsurface by the radioactive decay of uranium-series isotopes. After groundwater containing radon discharges to surface water bodies, radon activities decrease due to gaseous exchange with the atmosphere and radioactive decay (Ellins et al., 1990; Lee and Hollyday, 1993). High radon activities are therefore present in surface waters only in the immediate vicinity of points of groundwater inflow, and for relatively short 7

8 distances downstream of such locations. In some cases, ion chemistry can also be used to quantify groundwater inflow (Genereux et al., 1993; Cook et al., 2003), although changes in groundwater inflow rates will not always cause changes in ion concentrations. One of the difficulties of using radon to estimate groundwater discharge is that radon can be contributed to streams by both groundwater discharge and hyporheic exchange. Hyporheic exchange is the process by which stream water continually flows into and out of the river bed sediments. Although hyporheic exchange does not represent a net flux of water to the river, water will accumulate radon through radioactive decay of radium as it moves through sediments. This radon is then transported back to the river. The thickness of the hyporheic zone is usually a few centimeters to tens of centimeters, but Cook et al. (2006) showed that failure to consider hyporheic exchange can result in significant errors in groundwater discharge estimates. Quantification of groundwater discharge to the ocean is more difficult. In addition to radon, radium isotopes are commonly used in marine environments (Moore, 1996; Hancock et al., 2006). Radium is more useful in marine than freshwater environments because groundwater tends to partially mix with seawater near coastlines. This is especially true in the case of aquifers undergoing seawater intrusion. Unlike radon, most of the radium pool in aquifers is bound to sediments, but a portion of this pool is released in more saline groundwater. Like radon, with radium the flux of groundwater is determined by closing the mass balance, which requires knowledge of the groundwater activities, the flushing rate of the coastal water, surface water fluxes, and the removal pathways of the tracer within the marine environment. The most challenging aspect of the radium mass-balance in coastal environments is the estimation of the recirculated seawater flux (Colbert and Hammond, 2008; Lamontagne et al. 2008). This is seawater that moves in and out of shallow sediments and beachfaces through water motions produced by tides, waves and currents, picking-up small amounts of radium and radon in the process. An amendment to the Water Resource (Burdekin) Plan 2007 (WRP) to deal with the provision of groundwater is expected to be finalised in The amendment will identify sustainable allocation and management arrangements for groundwater resources in the Lower Burdekin. As a complement to this work, a National Water Commission (NWC) project has also been initiated to develop an integrated and holistic package of modelling tools to support the decision making process for water management in the Lower Burdekin. The proposed enhancements will promote the level of integration of natural and anthropogenic processes by providing a single modelling framework for data synthesis, informing data acquisition strategies and improving the reliability of model predictions. The work described in this report was commissioned by Queensland Department of Environment and Resource Management (DERM) to support the Lower Burdekin modelling toolkit. It aims to refine existing estimates of groundwater discharge from the Burdekin River floodplain aquifers using environmental tracer methods. Measurement of environmental tracer concentrations was carried out in rivers and creeks of the Burdekin River floodplain during an extensive sampling campaign in May Groundwater discharge has been estimated from comparison of electrical conductivity and radon activity in surface water and groundwater samples. Submarine groundwater discharge has been estimated from comparison of radon and radium isotope activities in groundwater and ocean water. Although these methods only provide fluxes at one point in time, and other methods must be used for 8

9 temporal extrapolation, they provide data on spatial scales useful for water management. It is important to note that estimates of groundwater discharge produced by Cook et al. (2004) did not consider hyporheic exchange, and so may have over-estimated discharge rates. Importantly, groundwater levels in 2011 are significantly higher than they were in 2004 (Figure 1.1), and so comparison of data obtained at these two times provides information on how groundwater discharge rates have changed over time, but also allows effects of groundwater discharge to be separated from hyporheic exchange. Figure 1.1. Groundwater levels for three representative bores within the Burdekin floodplain aquifer, and comparison with daily rainfall since

10 An extended sampling campaign was undertaken in May 2011 to sample submarine groundwater discharge in Bowling Green Bay (BGB) using radon, radium and salinity surveys. The total offshore flux for radium and radon was estimated using the Hancock et al. (2006) model, which was originally derived for similar work in the Great Barrier Reef (GBR). The main difference with the approach used by Cook et al. (2004) is that changes in the bathymetry can be accounted for using the Hancock et al. (2006) model. The radium and radon data collected in April 2004 were reanalysed using the Hancock et al. (2006) model to provide comparable estimates to the May 2011 sampling campaign. To help constrain the radium signature of the Burdekin aquifer, a subset of piezometers was sampled covering a greater range in groundwater salinity than the ones sampled by Cook et al. (2004). In addition, radium samples were collected at the Haughton River and Barratta Creek mouths over a tidal cycle to estimate the riverine radium signature entering BGB. As in the case of submarine groundwater discharge, river water tends to become enriched in radium in estuaries where fresh and salt water mix (Hancock and Murray, 1996). In the following, we present the results of the May 2011 sampling campaign for rivers and Bowling Green Bay and estimate groundwater discharge using radon and radium for that period. These results are then compared to the surveys undertaken with the same environmental tracers in

11 2. METHODS 2.1 General Groundwater discharge to rivers and creeks within the Burdekin River Delta has been estimated from comparison of radon and major ion chemistry in surface water and groundwater samples. Submarine groundwater discharge has been estimated from comparison of radon and radium isotope activities in groundwater and ocean water. 2.2 River Flows and Aquifer Hydraulics Flow rates for the major rivers at the time of sampling were obtained from gauging stations with automatic stage recorders located on the Burdekin River at Clare (Station B), Barratta Creek at Northcote (Station A) and Haughton River at Powerline (Station A; Figure 2.1). Stage height information was also obtained from 10 surface water monitoring sites. For these sites, stage heights are manually recorded at approximately monthly intervals, although none of these sites are currently operational. Stage heights for all stations were converted to AHD, and compared with groundwater level observations at nearby monitoring bores. The five closest bores were selected for this comparison, up to a maximum distance of 5 km from the surface water monitoring site. (Bores that were not considered representative of the regional aquifer were excluded from this comparison.) Differences between the water table and surface water elevation were used to determine the direction of surface water groundwater flow. Locations of the surface water monitoring sites, and the bores used for the comparison are shown in Figure Surface Water and Groundwater Sampling Groundwater samples were collected from 40 monitoring wells as part of the study by Cook et al. (2004). Between December 2003, samples were collected from 30 wells for analyses of 222 Rn. Between April 2004, a further eight wells were sampled for 222 Rn and five wells were sampled for radium isotopes. Results of electrical conductivity (EC), 222 Rn and radium isotope analyses are described in Cook et al. (2004). As part of the current study, a further five wells were sampled on 28 August 2011 for radium isotope analysis. Stream water samples were collected from the Burdekin River, Haughton River, Barratta Creek, and Plantation Creek. Samples from the lower reaches of the Burdekin River were collected using a small boat, whereas samples from the upper reaches were collected where road access permitted. In both cases, a small pump was used to pump water samples from between 0.3 and 0.5 m depth below the water surface. Field measurements were made of electrical conductivity (EC) using a WTW electrical conductivity meter. Samples for 222 Rn analysis were collected in 1250 ml bottles, and radon was extracted as described below. Seventy five sites were sampled between 9-14 May Locations of sampling sites are shown in Figures 2.2, 2.3 and 2.4. Radon and electrical conductivity measurements were also carried out on samples collected from these rivers between December 2003 and May 2004, and these have previously been described by Cook et al. (2004). Comparison 11

12 between 2011 samples and earlier samples provides some information on how groundwater discharge to the rivers and creeks changes over time. Figure 2.1. Locations of surface water gauging sites and bores used for comparison of hydraulic heads. 12

13 Figure 2.2. Locations of surface water sampling sites. (See Figures 2.3 and 2.4 for more detail.) Seven surface water samples for radium isotope analysis were collected from the lower reaches of the Haughton River and Barratta Creek on May River widths at sampling sites were measured using a laser range finder (Cook et al., 2004), or estimated from aerial photography. Samples of water within river bed sediments were collected from the Haughton River using a drive point piezometer. Three vertical profiles were constructed across the river approximately 41 km upstream of the mouth. Samples were taken at depths of 2, 4, 7, 10, 20, 40, 60, 80 and 100 cm below the sediment river interface using a hand driven Masterflex peristaltic pump. Samples for radon analysis were collected and field measurements made of electrical conductivity and temperature at each depth. Measurements of vertical hydraulic gradients between the river and 100 cm depth were made for each vertical profile using an air-water manometer. Sediment core samples were collected using a Dormer Undisturbed Wet Soil Sampler at each of the three vertical profile locations. 13

14 Figure 2.3. Locations of surface water sampling sites on the Haughton River and East and West Barratta Creek, and river distances (kilometres) upstream of the mouth. 14

15 Figure 2.4. Locations of surface water sampling sites on the Burdekin River and Plantation Creek. For the Burdekin River, distances (kilometres) upstream of the mouth are shown. 2.4 Sample collection in Bowling Green Bay Offshore samples were collected from Bowling Green Bay, using the Australian Institute of Marine Science (AIMS) RV Scorpio and RV James Kirby. The general sampling strategy was to collect both point and continuous measurements for salinity and radon along transects perpendicular to and parallel to the BGB shoreline (Fig. 2.5) as well as point radium samples perpendicular to the shoreline and at river mouths. Sampling took place on 9 May 2011 for some of the inshore Rn transects using the RV Scorpio and on May 2011 for the perpendicular and offshore Rn and Ra transects using the RV James Kirby. For the perpendicular transects, Transect 1 (T1) was located between the mouth of Baratta Creek and Haughton River, Transect 2 (T2) was off Cape Bowling Green and Transect 3 (T3) was located between T1 and T2. 15

16 T1 T3 T2 Figure 2.5. Location of the radon and radium point samples (circles) and continuous measurement transects (blue line) in May To help constrain the radon signature of recirculated seawater, sediment porewater profiles were collected near the shoreline in ~0.5 m water at T1 and T3 using a drive point. The drive point had a 5 cm screen and was inserted in the sediments by gentle percussion with a butyl rubber hammer. At T1, samples were collected at 37 cm and 58 cm depth in the sediments. At T3, samples were collected at 20 cm intervals down to 1 m. Radon was collected following the direct method of Leaney and Herczeg (2006). 2.5 Radon Analyses Measurements of radon activity in groundwater and river sediment water were made on 14 ml samples that were collected directly from the pump outlet using a syringe. The water sample was transferred to a pre-weighed teflon-coated PTFE scintillation vial containing 6 ml Packard NEN mineral oil cocktail. The radon activity was counted in the laboratory 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). Corrections were made for radioactive decay that occurs between the time of sampling and time of analysis in the laboratory. Radon activities in rivers and creeks were measured using liquid scintillation counting. Samples for liquid scintillation counting were collected in 1250 ml plastic bottles, which were filled without headspace. Within 24 hours of sample collection, radon was extracted from these water samples. Approximately 50 ml of water was first removed from the bottles, and then 20 ml of mineral oil scintillant was added from a pre-weighed scintillation vial. The bottle was shaken for four minutes to equilibrate the radon between the water-air-scintillant phases. After allowing the scintillant to settle to the top of the bottle (about 1 minute), the scintillant was returned to the vial, which was sealed. The vials were returned to Adelaide by courier 16

17 for counting within 7 days of sample collection, and counted by liquid scintillation. The analytical technique is described in more detail in Leaney and Herczeg (2006). Measurements of radon emanation were made on sediments collected from the bed of the river. Approximately 40 g of oven-dried sediment was sealed in 60 ml brass containers, with 20 ml of mineral oil scintillant. The balance of the volume (~ 20 ml) was filled with distilled water. After a period of several weeks, the radon activity within the chamber will reach a constant value as the radon production rate from the sediment will be exactly balanced by the radon lost by radioactive decay. After allowing six weeks for this secular equilibrium condition to be reached, the mineral oil was sampled and its radon concentration was measured. By using a series of radium standards, the efficiency of this process (percentage of emanated radon that is captured in the scintillant) was determined to be approximately 60%. The radon activity in the mineral oil is used to calculate the total radon emanation rate, E (Bq/kg), which is related to the radon production rate, (Bq/L/day) by: E (1 ) s [2.1] where s is the density of the solid phase (kg/cm) and is the porosity (cm/cm). Estimates of sediment porosity and density were made by adding sediments to a container of known volume and weight, adding water while agitating the samples such that entrapped bubbles were minimised, weighing the saturated sediments, followed by placing samples in an oven to then measure a dry weight. Radon activities in seawater were measured using a commercial radon-in-air monitor (Burnett and Dulaiova, 2003). Seawater was pumped directly through an air-water exchanger, which removes 222 Rn from the water by evasion into the chamber. The 222 Rn-enriched air was circulated in a closed air-loop connected to the monitor. The monitor counts -decays of radon daughters, and discriminates different decays in energy-specific windows. With this method, continuous measurements of 222 Rn can be made. 2.6 Radium Analyses Radium isotopes were extracted from large volume water samples by adsorbing the isotopes onto MnO 2 -coated acrylic fibers (Moore, 1976). Water was gravity-fed or slowly pumped past the MnO 2 -fibers, thereby removing nearly all of the dissolved radium. Activities of the short-lived isotopes 223 Ra and 224 Ra are measured in the laboratory shortly after sample collection by the detection of the -decay of their respective daughter nuclides ( 219 Rn and 220 Rn) with photomultiplier tubes, and their identification with a delayed coincidence circuit (Moore and Arnold, 1996). Activities of the long-lived radium isotopes, 226 Ra and 228 Ra, are determined by -ray spectrometry (Moore, 1984). 17

18 3. DISCHARGE TO STREAMS 3.1 Introduction Although the direction and relative magnitude of groundwater surface water exchange can be assessed from a comparison of surface water and groundwater heads, the dynamic nature of the aquifer and the uncertainty in hydraulic parameters mean that it is difficult to quantify the flux using hydraulic methods. In this chapter, inflows of groundwater to the Burdekin River, Haughton River, Barratta Creek and Plantation Creek have been estimated based on comparison of radon activities in surface waters and groundwater. Surface water sampling took place between 9 12 May 2011, and was designed to coincide with expected maximum groundwater inflow. Results are compared with directions of flow determined from comparison of surface water and groundwater levels. Results are also compared with samples collected between 8 16 December 2003, 3 6 February 2004, and between 26 April 4 May Figure 3.1 depicts flow rates of Burdekin River, Haughton River and East Barratta Creek between 2003 and 2011, and rainfall at Giru over the same period. The mean annual flows of the three rivers over this period are 14100, 250 and 570 GL, respectively. An increase in flow over this time is also apparent, and consistent with the increase in rainfall. Figure 3.1. Average daily flow rates of Burdekin River at Clare (Station B), Haughton River at Powerline (119003A) and Barratta Creek at Northcote (119101A), and daily rainfall at Clare between January 2003 and June

19 3.2 Theory In the absence of surface water inflow or direct rainfall input, change in streamflow with distance is simply given by: Q x I Ew [3.1] where Q is the stream flow rate (m 3 /day), I is the groundwater inflow rate per unit of stream length (m 3 /m/day), E is the evaporation rate (m/day), w is the width of the river surface (m), and x is distance in the direction of flow. For a conservative tracer (such as chloride), the solute mass balance is given by: dqsc dx s Ic [3.2] i where c s and c i are the concentrations of the stream and groundwater inflow respectively. Changes in stream flow chemistry are therefore related to groundwater inflow rate by: dcs Qs I( ci cs) wecs [3.3] dx Groundwater inflow rates can therefore be calculated if the other parameters are known. However, radon is also affected by radioactive decay, gas exchange and hyporheic exchange, so additional terms are required in the mass balance. The radon mass balance is given by (Cook et al., 2006): Qc wh c Ici kwc dwc x 1 t h [3.4] where c is the radon activity within the stream (Bq/L), c i is the activity in groundwater inflow (Bq/L), Q is the stream flow rate (m 3 /day), I is the groundwater inflow rate per unit of stream length (m 3 /m/day), k is the gas transfer velocity across the water surface (m/day), is the radon production rate within the hyporheic zone (Bq/L/day), is the radioactive decay constant (day -1 ), w is the width of the river surface (m), d is the mean stream depth (m), and x is distance in the direction of flow. The equation for radon activity as a function of distance therefore becomes: c Q I( c x i wh c c ) wec kwc dwc 1 t h [3.5] The decay coefficient for radon is = 0.18 day -1. The terms on the right hand side of Equation 3.5 represent changes in activity due to groundwater inflow, evaporation (which increases the radon activity in the remaining water), gas exchange (which 19

20 decreases the radon activity), radioactive decay (which decreases the activity) and hyporheic exchange. The relative magnitudes of the second third and fourth terms are proportional to E, k and d, respectively. Since E is usually in the range m/day (1-10 mm/day), and k is usually in the range m/day (Wanninkof et al., 1990), it is clear that the evaporation term will usually be negligible. The relative magnitudes of the gas exchange and radioactive decay terms will depend on the value of k and the river depth. For a gas exchange velocity of k = 1 m/day, the radioactive decay term will dominate when the river depth exceeds approximately d = 5.5 m. For shallow streams, gas exchange is the main process controlling radon loss. It is apparent from Equation 3.5, that radon can be contributed by hyporheic exchange as well as groundwater inflow. Figure 3.2 depicts the relationship between c and t h, for typical aquifer parameters for the case of negligible groundwater inflow (I = 0). It shows that, under suitable conditions, significant radon activities in rivers can arise due to hyporheic exchange. Also apparent from Figure 3.2 is the relative insensitivity of radon activity in the river to hyporheic zone residence time, provided that the residence time is less than a few days. This is because when residence times are short, almost all of the radon produced in the hyporheic zone is transported into the river. However, at residence times longer than a few days, a significant fraction of the radon that is produced is lost through radioactive decay and never enters the river. In the case where the hyporheic zone sediments are mineralogically equivalent to groundwater sediments, then / = c i, and so Equation 3.5 can be written c Q x I i ( ci c) wec kwc dwc [3.6] 1/ th wh c c It is apparent from Equation 3.6 that there is an similarity between groundwater inflow and hyporheic exchange, so that I wh 1/ t h [3.7] Thus these two processes will cause identical changes to the radon activity in the river, and hence can be indistinguishable unless hyporheic exchange is independently constrained. Groundwater discharge rates have been quantified by solving Equations 3.1 and 3.5 numerically, using an EXCEL spreadsheet and an explicit finite difference approach. The model uses a spatial discretisation of between 51 and 62 m (one thousandth of the simulated river length). Values for most model parameters are based on measured values, and held fixed during model calibration (Table 3.1). The model is calibrated by manually adjusting groundwater inflow rates until the model matches the observed radon activities in the river. 20

21 Figure 3.2. Relationship between the residence time of water within the hyporheic zone and radon activity within a river receiving no groundwater inflow (and with negligible evaporation). Curves are based on a radon production rate of = 2500 mbq/l/day, gas exchange rate of k = 1 m/day, river depth of d = 1 m and hyporheic zone porosity = 0.4 and thicknesses of h = m. Table 3.1. River radon mass balance model parameters. Parameter Units Description Q (m 3 /s) River flow rate Q 0 (m 3 /s) Upstream river flow rate Q ti (m 3 /s) Inflow rate from tributary i I (m 3 /s) Groundwater inflow rate C (mbq/l) Concentration in river c 0 (mbq/l) Concentration in groundwater inflow c t1 (mbq/l) Inflow concentration from tributary i c i (mbq/l) Concentration in groundwater inflow k (m/day) Gas transfer velocity w (m) River width d (m) River depth h (m) Hyporheic zone depth (m) Hyporheic zone porosity (mbq/l/day) Radon production rate in hyporheic zone t h (days) Hyporheic zone residence time 21

22 Cook et al. (2006) defined the scale length for changes in radon activity in rivers, based on a simplified analytical solution to Equation 3.5. The scale length is the distance after which concentration decreases to 37% (1/e) of its initial value, when inflow, evaporation and hyporheic exchange are all negligible, and k, w and d do not change with distance. The scale length is equal to: Q L [3.8] kw dw Significant errors in groundwater inflow rates can occur if the spacing between surface water sample locations is greater than the scale length. When estimating groundwater inflow from longitudinal sampling of stream chemistry, an assumption must be made about where between sampling points the groundwater inflow occurs. Here, it is assumed that the rate of groundwater inflow is constant between sampling points. However, if inflow is concentrated at the upper or lower ends of the reach, then errors will occur. Figure 3.3 depicts the potential error due to assuming that groundwater inflow is constant between sampling points. (See Appendix 1 for derivation.) Where the distance between samples is much less than the scale length, the potential error is relatively small, but when the distance increases beyond the scale length the potential error increases rapidly. Figure 3.3. Effect of sample spacing on potential errors in estimated groundwater inflow rates from radon activities in the river. I ˆ / I is the ratio of the estimated value to the actual value. 3.3 Radon Activity in Groundwater The distributions of radon activities and electrical conductivities in groundwater and surface water are shown in Figure 3.4. In the case of radon, surface water and groundwater activities are distinct, and the distribution of groundwater is narrow. These two qualities make radon a useful tracer for estimating groundwater discharge to surface water. Electrical conductivity distributions are much broader, but also show 22

23 distinct overlap. For this reason, electrical conductivity is not a useful tracer of surface water groundwater interaction in the Burdekin. For the 38 sampled bores, radon activities ranged between 2735 and mbq/l, with 70% of sampled bores having activities between 5000 and mbq/l. There is no obvious spatial patterning to these observed variations. The mean activity of all bores is approximately mbq/l. Figure 3.4. Distribution of radon activities and electrical conductivities in groundwater and surface water. 3.4 Burdekin River Records of Burdekin River levels are available for the gauging station (Burdekin Clare B), and for eight other sites (see Appendix 3). At Clare weir, the most upstream site ( ), groundwater levels were below river levels until approximately 2005, after which there was a significant rise in groundwater levels, and since which time levels observed in some bores have been above the river level. At Clare A Pump Station ( ), located approximately 7 km below Clare Weir, river levels have not been measured since 1997, but comparison between current groundwater levels and river levels during the measurement period suggests that the 23

24 river would be gaining at this point. The river is also clearly gaining at B, where groundwater levels are approximately 5 m above river levels. Groundwater levels also appear to be significantly above the river level at The Rocks Pumping Station ( ). At SBWB Pumping Station ( ), groundwater levels appear to be similar to river levels, although measurement of river levels was discontinued in the late 1980s. The river appears to be losing at the Burdekin Ridge Bridge at Ayr ( ). Figure 3.5. Radon activity and electrical conductivity of the Burdekin River upstream of the mouth. (Flow direction is from left to right.) Samples collected in May 2011 are compared with those collected in 2003 and 2004 (Cook et al., 2004). Radon activities and electrical conductivities measured in the Burdekin River in May 2011 are shown in Figure 3.5. Immediately upstream of Clare Weir, the radon activity was 214 mbq/l, but this reduced to 124 mbq/l immediately downstream of the weir. The reduction in radon activity is due to degassing as water plunges over the weir. The radon activity between the weir (at 62 km) and 34.2 km varies between 124 and 316 mbq/l. At 41.3 km, the radon activity immediately upstream of a set of small rapids is 171 mbq/l, but this increases to 241 mbq/l immediately downstream of the rapids. This increase is likely due to hyporheic exchange, with water upwelling at the base of the rapids. Variations in radon activity in the upper part of the river are also 24

25 therefore likely due to variations in hyporheic exchange rate. Radon activity increases downstream of 27.8 km, and is between 340 and 590 mbq/l between 28.7 and 25.5 km. A number of radon samples were also collected from locations close to the bank, where groundwater appeared to be entering the river. Thus, for example, a sample collected from the west bank of the river at 28.7 km had a radon activity of 3420 mbq/l and electrical conductivity of 308 S/cm. Similarly, a sample collected from the near the bank of the river at 25.5 km had a radon activity of 4250 mbq/l and electrical conductivity of 201 S/cm. These samples (which are not shown on Figure 3.5, but included in Appendix 2) likely represent groundwater which was flowing into the river, but which had not yet fully mixed with river water. Figure 3.6 shows the variation in radon activity and electrical conductivity in a transect across the river, located 22.6 km upstream of the river mouth. This crosssection is located within a broad meander in the river, and the river bottom slopes gently away from the bank on the inside of the meander, but is much steeper on the northern bank. Theoretical studies have shown that groundwater discharge to streams should be concentrated close to the banks (Genereux and Bandopadhyay, 2001), and on the insider of meander bends. Figure 3.6 shows significantly lower electrical conductivity and higher radon activity close to the bank on the inside of the meander, which is consistent with groundwater discharge occurring close to the bank. Figure 3.6. Radon activity and electrical conductivity in a transect across the Burdekin River, 22.6 km upstream of the mouth. Radon activity is higher, and electrical conductivity is lower near the bank on the inside of the meander, indicating groundwater discharge. 25

26 Further downstream, radon activities are between 209 and 285 mbq/l, although a slightly elevated activity of 383 mbq/l was measured at the Groper Creek boat ramp. Radon activities observed in are mostly lower than those observed in Highest values in were measured in May and Sept. However, at these times, flow rates were much lower than in May 2011 (9.2 and 10.4 m 3 /s, respectively, compared with 140 m 3 /s in May 2011), and so the results are not directly comparable. Cook et al. (2004) measured the radon activity of the Burdekin River at Pump Station No. 3 on six separate occasions between December 2003 and September At the time of sampling, the flow rate of the Burdekin River at Clare Weir varied between 9 m 3 /s and 190 m 3 /s. This data is depicted in Figure 3.7, along with the radon activity measured in May 2011, when the flow rate at Clare Weir was 140 m 3 /s. For 2003 and 2004 data, radon activities increase as river flow decreases, which is consistent with a greater proportion of the flow being derived from groundwater during low flow periods. However, data for May 2011 does not fit the trend of the 2003 and 2004 data. Also shown in Figure 3.7 is the relationship between flow rate at Clare and the total radon flux at Pump Station No. 3 (calculated as the radon activity multiplied by the river flow rate). Although river flow rate varies by more than a factor of 20, the radon flux for the 2003 and 2004 data varies only by a factor of two, which is consistent with a radon input upstream of this location that is relatively stable over time. However, the radon flux in May 2011 is much greater than in 2003 and 2004, which is consistent with a much greater groundwater input at this time. Figure 3.7. Relationship between flow rate of the Burdekin River at Clare and radon activity and radon flux at Pump Station No. 3 (28.7 km river distance) in (open circles) and May 2011 (closed circle). Points represent observations on 9 December 2003, 3 February 2004, 26 April 2004, 27 April 2004, 2 May 2004, 8 September 2004 and 10 May (On 2 May 2004, the radon activity for Pump Station No. 3 has been interpolated from measured activities at Sites 53 and 54.) Groundwater inflows are estimated from radon activities using a numerical model. The river width, river depth and hyporheic zone depth used in the modelling are shown in Figure 3.8. Constant values of river depth and hyporheic zone depth have been used for the entire length of river. Hyporheic zone parameters are based on measurements in the Haughton River (see Section 3.5), and are also assumed to be constant. A constant river width of w = 100 m has been used for upstream areas, increasing to 200 m at 17 km, and 500 m between 12 km and the mouth. (Although 26

27 the river depth probably increases towards the mouth, the model is relatively insensitive to river depth, as discussed below, and so we have used a constant value.) A constant gas exchange velocity of k = 1 m/day has also been used for the entire river length. Studies on low gradient rivers suggest values for the gas exchange velocity of 0.5 < k < 2 m/day (Hibbs et al., 1998; Chapra and Wilcock, 2000; Raymond and Cole, 2001). The evaporation rate has been modelled as E = 7 mm/day, although the simulations are insensitive to this parameter. Model parameters are given in Table 3.2. Figure 3.8. Simplified river width (w), river depth (d) and hyporheic zone depth (h) in the Burdekin River. River depth and hyporheic zone thickness are assumed to be constant along the river length. River width is modeled as 100 m in the upstream reaches, increasing to 500 m at the mouth. Pumping also takes place from the Burdekin River, and must be included in the model. Six pumping stations located between Clare Weir and the river mouth distribute water into a network of creeks and irrigation channels. Total water extraction from these pumping stations in April 2004 and May 2011 was approximately 7.9 m 3 /s and 4.1 m 3 /s, respectively. Estimates of the pumping rate of each of the six major pumping stations at the time of sampling have been entered into the model as sink nodes. In addition, 96 private irrigators pump water directly from the river, and are together licensed to extract an additional m 3 /year. 27