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Changes in water quality following tidal inundation of coastal lowland acid sulfate soil landscapes
Estuarine, Coastal and Shelf Science 81 (2009) 257–266
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Changes in water quality following tidal inundation of coastal lowland acid sulfate soil landscapes Scott G. Johnston a, *, Richard T. Bush a, Leigh A. Sullivan a, Edward D. Burton a, Douglas Smith b, Michelle A. Martens b, Angus E. McElnea b, Col R. Ahern b, Bernard Powell b, Luisa P. Stephens b, Steve T. Wilbraham b, Simon van Heel b a b
Centre for Acid Sulfate Soil Research, Southern Cross Geoscience, Southern Cross University, Lismore, NSW 2480, Australia Department of Natural Resources and Water, Block C, 80 Meiers Road, Indooroopilly, Qld 4068, Australia
a r t i c l e i n f o
a b s t r a c t
Article history: Available online 18 November 2008
This study examines the remediation of surface water quality in a severely degraded coastal acid sulfate soil landscape. The remediation strategy consisted of partial restoration of marine tidal exchange within estuarine creeks and incremental tidal inundation of acidified soils, plus strategic liming of drainage waters. Time-series water quality and climatic data collected over 5 years were analysed to assess changes in water quality due to this remediation strategy. A time-weighted rainfall function (TWR) was generated from daily rainfall data to integrate the effects of antecedent rainfall on shallow groundwater levels in a way that was relevant to acid export dynamics. Significant increases in mean pH were evident over time at multiple monitoring sites. Regression analysis at multiple sites revealed a temporal progression of change in significant relationships between mean daily electrical conductivity (EC) vs. mean daily pH, and TWR vs. mean daily pH. These data demonstrate a substantial decrease over time in the magnitude of creek acidification per given quantity of antecedent rainfall. Data also show considerable increase in soil pH (2–3 units) in formerly acidified areas subject to tidal inundation. This coincides with a decrease in soil pe, indicating stronger reducing conditions. These observations suggest a fundamental shift has occurred in sediment geochemistry in favour of proton-consuming reductive processes. Combined, these data highlight the potential effectiveness of marine tidal inundation as a landscapescale acid sulfate soil remediation strategy. Ó 2008 Elsevier Ltd. All rights reserved.
Keywords: sea-water inundation tidal wetland estuary time-series analysis floodgate
1. Introduction Drainage and reclamation of low lying coastal sulfidic sediments for agricultural purposes is a widespread practice that has potential to cause substantial acidification and degradation of estuarine water quality (e.g. Dent, 1986; Minh et al., 1997; White et al., 1997). The international literature is replete with examples of such projects leading to acidified waterways, mobilisation of iron, aluminium and trace metals, along with other forms of environmental degradation (e.g. Sammut et al., 1996; Wilson et al., 1999; Sundstrom et al., 2002; Johnston et al., 2004a; Macdonald et al., 2004; Haraguchi, 2007). A variety of studies have examined the application of different remediation techniques to improve the water quality associated with drained coastal lowland acid sulfate soils (CLASS). Some examples of remediation techniques include partial tidal exchange
* Corresponding author. E-mail address: [email protected] (S.G. Johnston). 0272-7714/$ – see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.ecss.2008.11.002
within drains (Indraratna et al., 2002; Johnston et al., 2005a), the use of in-drain weirs or controlled drainage (Blunden and Indraratna, 2000; Johnston et al., 2004b; Åstro¨m et al., 2007), and techniques involving the use of alkaline reagents (Green et al., 2006). Tidal inundation of CLASS is a further remediation option available when their elevation is close to mean sea level. However, this strategy has rarely been applied on a large scale. Tidal inundation of acidified sediments may lead to re-establishment of proton-consuming and alkalinity-producing geochemical processes involving the decay of organic matter under sub-oxic/ anoxic conditions (e.g. Ponnamperuma, 1972; van Breemen, 1975; Portnoy and Giblin, 1997; Johnston et al., 2005b; Burton et al., 2007). In essence, establishing tidal inundation of CLASS may initiate similar diagenetic forces to those commonly found in intertidal sedimentary environments such as mangroves (i.e. high water tables, abundant sulfate and organic matter). Such conditions will stimulate upward migration of the redox boundary, favouring the reductive dissolution of Fe (III) minerals and the reduction of SO2 4 . In turn, this may lead to eventual formation of various solid phase
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Fe (II)-sulfide minerals (ie. pyrite, greigite and mackinawite) in formerly oxidised, acidified sulfuric horizons (Bush and Sullivan, 1997; Burton et al., 2006a, 2007). Any associated proton consumption and generation of alkalinity will be highly dependent upon iron and sulfur bio-mineralisation pathways. Whilst these pathways are variable and complex (Berner, 1984; van Breemen, 1993), the key reactions involving the generation of bicarbonate alkalinity can be described according to Eqs. (1) and (2), 2CH2 O þ SO2 4 /H2 S þ 2HCO3
(1)
CH2 O þ 4FeOOH þ 7Hþ /4Fe2þ þ HCO 3 þ 6H2 O
(2)
Although Fe (III) and SO2 reduction generates alkalinity, both 4 reactions produce potentially mobile aqueous species that may reoxidise and release acidity. Hence, the burial of these species via the formation of diagenetic minerals (e.g. FeS2) is a key to long-term amelioration of acidity. There are few examples of studies involving large scale reestablishment of tidal inundation in CLASS landscapes combined with an assessment of the effects of this technique upon the frequency and magnitude of acidification of adjacent surface waters. This study aims to a) analyse a medium term (w5 years) time-series water quality data set derived from waterways in a drained and reclaimed CLASS landscape that was subject to reintroduction of tidal exchange and tidal inundation and b) to identify any changes in water quality that are attributable to the above remediation strategy. 2. Materials and methods 2.1. Study site and remediation strategy The study site, East Trinity, is a Holocene sedimentary coastal plain of w800 ha located near Cairns in northern Australia (Fig. 1). The Cairns area has a tropical monsoonal climate with summerdominant rainfall. The site was diked and intensively drained for sugar cane production in the 1970s (Powell and Martens, 2005) and originally contained large areas of estuarine wetlands, including mangroves and saltmarsh. There are several estuarine creeks located at the site; Hills Creek, Magazine Creek, Firewood Creek and Georges Creek. The dike, or bund wall, effectively excluded tidal exchange from these creeks. One-way floodgates, located at the intersection of the bund wall with Hills Creek and Firewood Creek, allowed drainage waters to exit the site during the ebb tide. The exclusion of tidal exchange combined with intensive drainage led to lowering of water tables and subsequent oxidation and subsidence of sulfidic sediments in former mangrove areas (Hicks et al., 1999). This resulted in the formation of severe acid sulfate soils and seasonal export of large quantities of acidity from the site (Cook et al., 2000; Hicks et al., 2002; Russell and Helmke, 2002; Smith et al., 2003; Powell and Martens, 2005). Investigations have demonstrated that most of the actual acid sulfate soil material at the site is located below 0.5 m AHD (Australian Height Datum; 0 m AHD is wmean sea level) (Smith et al., 2003; Powell and Martens, 2005). The remediation strategy is detailed in Powell and Martens (2005) and consists mainly of controlled tidal exchange involving ‘‘.progressively and cautiously replacing the existing acidified freshwater environment with a managed tidal wetland system, by modifying and managing the infrastructure that initiated the problem some 30 years ago.’’ Key aspects of the strategy include; Incremental opening of floodgates to allow regular tidal inundation of the land surface inside the bund wall, up to a maximum elevation of approximately 0.5 m AHD.
Water quality monitoring and strategic treatment of acid drainage water with hydrated lime, [Ca(OH)2]. This was regarded as an important adjunct to the main strategy, particularly during the early stages of reintroducing tidal exchange, when the potential for mobilisation of stored acidity from the system was greatest (see Smith et al. (2003) and Powell and Martens (2005) for further details). A primary goal of the remediation strategy at East Trinity is to ‘‘have water of acceptable quality (pH > 6) exiting the site, in all seasonal conditions, under a self-managed tidal regime without lime augmentation’’ (Smith et al., 2004, pp. 3). Hills Creek has been subjected to daily tidal exchange since February 2003. The extent of regular tidal inundation within its sub-catchment was incrementally increased between 2003 and 2005, to a maximum level of approximately 0.5 m AHD. Firewood Creek was subject to partial and restricted tidal exchange commencing in mid-2002 and followed by more extensive tidal exchange from 2005 onward. Managed floodgate opening in Firewood Creek currently (i.e. 2007) allows tidal exchange to an elevation of approximately 0.25–0.3 m AHD. 2.2. Time-series water quality data The three water quality monitoring sites were chosen for detailed analysis and interrogation – Firewood Creek-Bund, Firewood Creek-Upper and Hills Creek-Middle (Fig. 1). These sites were chosen due to the availability of long duration data sets (>4 years) and the strategic and contrasting nature of their locations. The pH and EC (electrical conductivity) data were regarded as having greatest value in terms of their relevance to the goals of the remediation strategy. Data analysis focussed on these two key parameters as well as daily rainfall. Whilst pH can only be regarded as an indicator of the total acidity within waterways located in CLASS landscapes (due to the presence of acidic metal cations, e.g. Hicks et al., 1999; Cook et al., 2000), it is nonetheless a very useful proxy due to the primary control it exerts over trace metal mobility, solubility, speciation and bio-toxicity. The pH, EC and temperature were logged at 10–15 min intervals using either TPS 90FL or Hydrolab Quanta multi-parameter sondes in combination with Campbell Scientific CR10X loggers. Sensor probes were situated in the upper 1 m of the water column and were maintained and serviced regularly with intervals between calibration ranging from 1 to 2 weeks. The following general processing and collation tasks were applied to the raw time-series water quality data files for each site. Screening to identify and remove any spurious or outlier values. Screening to identify and remove data from periods that were subject to logger calibration drift. Screening to identify and remove parameter values outside the practical environmental range and/or the instrument range.
2.3. Water quality trend assessment and statistical analysis The mean daily pH and EC were calculated for each of the three sites for the duration of the assessment period (2002–2007). Mean daily values were regarded as more useful for analysis of longer term trends due to the smoothing of extreme temporal variation associated with tidal modulation and decreasing auto-correlation via aggregation. Acidification episodes in coastal floodplain waterways are typically strongly related to shallow groundwater levels and thus antecedent rainfall (e.g. Wilson et al., 1999; Cook and Rassam, 2002;
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Fig. 1. East Trinity study site location map showing dike/bund wall and water quality monitoring stations. Soil monitoring sites are displayed in inset ‘a’ and ‘b’.
Johnston et al., 2004a). The two main climatic drivers of the floodplain water balance which govern floodplain water tables and thus acid export are precipitation (P) and evapotranspiration (ET) (White et al., 1997). To assess the effectiveness of the remediation strategy it was necessary to interrogate the data sets for changes in relationships between rainfall, shallow groundwater levels and drainage water quality. However, there was a lack of time-series shallow groundwater level data for the project site during the assessment period. Thus, there was a need for a quantitative measure of rainfall, beyond that of simple daily data, which adequately captured the influence of antecedent rainfall in a manner meaningful to acid export. For this purpose a timeweighted rainfall (TWR) function was developed using Eq. (3).
TWRðxÞ ¼
30 X
½ðpx1 þ /px4 Þ þ ðpx5 þ /px8 Þ=2
x¼1
þ ðpx9 þ /px12 Þ=3 þ ðpx13 þ /px16 Þ=4 þ ðpx17 þ /px30 Þ=5
ð3Þ
where at time (t) ¼ x, px1 is the daily rainfall at t ¼ x 1 day, and px2 is the daily rainfall at t ¼ x 2 days etc., generating
a weighted rainfall total over a prior 30 day period. The denominators in Eq. (3) follow a simple arithmetic progression. This pattern of weighting is based on the recognition that shallow ground water levels in coastal floodplain environments rise rapidly in response to rainfall and decay slowly over time (typically in the order of months) in response to losses from drainage and evapotranspiration, thus producing a typical sawtooth pattern (e.g. Johnston et al., 2004b). Hence, rainfall events most closely preceding a given date of water quality monitoring are likely to have a greater influence on groundwater levels and therefore drainage water quality than more temporally distant rainfall events. However, temporally distant rainfall events will still have a diminished influence on antecedent shallow groundwater levels, as some of that rainfall remains in storage within the soil profile. The TWR broadly mimics such a pattern. While the precise choice of time intervals and associated weighting is somewhat arbitrary, the TWR provides a far more integrated approach to accounting for antecedent rainfall effects than either daily precipitation values alone or a running mean of daily precipitation values calculated over a set time interval. All statistical analysis was conducted using SPSS 15 software. Time-series water quality data is typically auto-correlated and
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Fig. 2. Temporal changes in mean daily pH and EC in relation to daily rainfall at Firewood-Bund and Firewood-Upper water quality monitoring stations from January 2002 to August 2007. Dashed lines distinguish the temporal periods used in comparative statistical analysis.
often non-normally distributed which can inhibit the reliability of ordinary parametric statistical analysis. Linear regression analysis was used to determine whether there was evidence of temporal changes in the nature of relationships (regression coefficients) between key parameters. Auto-correlation can increase the likelihood of type 1 error when assessing relationship significance using ordinary least squares techniques. Therefore a first order autoregressive process using an exact maximum likelihood model was used to assess the significance of those relationships (Ostrom, 1990). Residuals were then analysed to ensure there had been adequate removal of auto-correlation errors and ensure nonviolation of model assumptions regarding normality and variance. These analyses were applied to mean daily pH in relation to both mean daily EC and TWR at each site for discrete time intervals. Rather than create time intervals based on calendar years, a seasonal division was used where possible, thus allowing a more meaningful comparison of successive wet seasons. For the
Firewood Creek data sets, this interval was from 1st September to August 31st the following year. The Hills Creek-middle data set was divided into four time intervals based on periods of data availability. The mean pH for these time intervals were compared using an unequal variance version of the t-test after rank transformation to account for non-normal distribution (Conover, 1998). 2.4. Rainfall data Rainfall data were derived from a weather station (automatic tipping bucket rain gauge) located at the Firewood-Bund wall and a manual rain gauge located at the House site (Fig. 1). Data from the automatic weather station was used where possible. However, after 2005 the automatic weather station failed. From 2005 onwards all rainfall data is derived from the manual rain gauge. Due to concern that this may bias the TWR, rainfall data from the two sites were compared for a w4 month period during 2004. While the House
Fig. 3. Changes in mean daily pH at a) Firewood Creek-Bund and b) at Firewood Creek-Upper water quality monitoring stations during select time periods from 2002 to 2006-07. The 90th, 75th, 25th and 10th percentiles are shown along with the median and outliers (open circles).
S.G. Johnston et al. / Estuarine, Coastal and Shelf Science 81 (2009) 257–266 Table 1 Comparison of the difference between the mean of daily pH values for select time periods at Firewood Creek-Bund and Firewood Creek-Upper water quality monitoring stations. Period
Differencea
261
monitoring stations located in Hills Creek sub-catchment between 2001 and 2007. Measurements were made in-situ using platinum tipped electrodes (details are described by Hicks et al., 1999) approximately every 2–6 months and data are corrected to a standard hydrogen electrode.
Firewood Creek-Bund/Firewood Creek-Upper
2002–03 2003–04 2004–05 2005–06 2006–07
2002
2002–03
2003–04
2004–05
2005–06
**/n.s. n.s./** **/** **/** **/**
– **/** **/** **/** **/**
– – **/** **/** **/**
– – – **/* **/n.s.
– – – – **/n.s.
a Unequal variance t-test on rank transformed data, * P < 0.05, ** P < 0.01, n.s., not significant.
site tends to have slightly higher registrations, the two sites exhibit a strong linear relationship (r2 ¼ 0.86, y ¼ 1.247x þ 10.9, where x ¼ Bund wall). However, in order to reduce the likelihood of this change in rainfall data sources introducing a temporal bias to the TWR, the TWR calculated on the basis of Bund wall weather station data was transformed according to the above relationship prior to further analysis.
2.5. Soil pH and redox potential Soil pH data is based on field measurements derived from repeat sampling at set locations in both Hills Creek and Firewood Creek sub-catchments. Soils were collected via gouge augers or push corers and analysis conducted immediately by inserting a freshly calibrated probe either directly into moist soil or a 1:1 fresh soil/ water solution. Redox potential was measured at two fixed
3. Results 3.1. Firewood Creek Firewood Creek catchment is relatively small (w2 km2) and is confined to the coastal floodplain. A large proportion of the Firewood Creek catchment contained highly acidic soils (pH > 3.5) (Smith et al., 2003). Data was compiled from two water quality monitoring sites in this catchment area, Firewood-Bund and Firewood-Upper (Fig. 1). The seasonal dynamics in mean daily pH for both sites relative to daily rainfall are shown in Fig. 2. During the dry season there were long periods of relatively high (w6.5–8) and comparatively stable pH coupled with high EC. This contrasts with alternating periods of low (<4.5) and highly dynamic pH and low EC associated with wet seasons. The mean daily pH values at both monitoring sites display a clear increase with time whilst the variability of mean daily pH values decreases (Fig. 3). These changes in aggregated mean daily pH values are significant for most periods at P ¼ 0.01 (Table 1). Firewood-Upper generally had lower mean values than the Firewood-Bund site, which accords with its greater distance from the estuary, the source of marine buffering. There are significant positive linear relationships between EC and pH at both sites, with distinct temporal groupings evident in the data (Fig. 4a and b). The slope of the relationships between EC and pH change progressively over time (Table 2). Low EC events continue to occur, but they become less strongly associated with
Fig. 4. Correlation between mean daily EC and mean daily pH at a) Firewood Creek-Bund and b) at Firewood Creek-Upper water quality monitoring stations during select time periods. Correlation between time-weighted rainfall (TWR) and mean daily pH at c) Firewood Creek-Bund and d) at Firewood Creek-Upper water quality monitoring stations during select time periods. Linear regression, dashed lines are 99% confidence intervals (Note: 1997–98 data is based on spot monitoring by Hicks et al., 1999).
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Table 2 Comparison of linear regression equations and P for mean daily EC (x) and mean daily pH (y) for select time periods at Firewood Creek-Bund and Firewood CreekUpper water quality monitoring stations. Period
2002 2002–03 2003–04 2004–05 2005–06 2006–07 a
Firewood Creek-Bund
Firewood Creek-Upper a
y ¼ ax þ b
P
y ¼ ax þ b
Pa
y ¼ 0.081x þ 2.84 y ¼ 0.067x þ 3.79 y ¼ 0.031x þ 5.35 y ¼ 0.044x þ 5.56 y ¼ 0.029x þ 6.01 y ¼ 0.029x þ 5.98
** ** ** ** ** **
– y ¼ 0.079x þ 3.10 y ¼ 0.090x þ 2.63 y ¼ 0.036x þ 5.67 y ¼ 0.030x þ 5.91 y ¼ 0.015x þ 6.56
** ** ** ** **
Using an autoregressive exact maximum likelihood model, * P < 0.05, ** P < 0.01.
Table 3 Comparison of linear regression equations and P for for TWR (x) and mean daily pH (y) for select time periods at Firewood Creek-Bund and Firewood Creek-Upper water quality monitoring stations. Period
2002 2002–03 2003–04 2004–05 2005–06 2006–07 a
Firewood Creek-Bund
Firewood Creek-Upper a
y ¼ ax þ b
P
y ¼ 0.016x þ 7.58 y ¼ 0.008x þ 7.62 y ¼ 0.003x þ 7.06 y ¼ 0.004x þ 7.95 y ¼ 0.002x þ 7.56 y ¼ 0.002x þ 7.68
** ** ** * ** **
y ¼ ax þ b
Pa
y ¼ 0.020x þ 7.53 y ¼ 0.011x þ 7.60 y ¼ 0.008x þ 7.57 y ¼ 0.003x þ 7.65 y ¼ 0.002x þ 7.47 y ¼ 0.002x þ 7.49
** ** ** ** ** **
Using an autoregressive exact maximum likelihood model, * P < 0.05, **P < 0.01.
decreases in pH. This trend shift is particularly evident at the Firewood-Upper site (Fig. 4b). Spot monitoring water quality data collected from Firewood Creek in 1997–1998 prior to the implementation of tidal exchange (derived from Hicks et al., 1999), is consistent with the EC/pH signature observed during acid export events in 2002–2003, during the early stages of remediation (Fig. 4b).
There are also significant negative linear relationships between TWR and pH at both sites (i.e. increasing antecedent rainfall ¼ decreasing pH) (Fig. 4c and d). A clear progressive change over time is evident in the slope of relationships between TWR and mean daily pH at both locations (Table 3), whereby the decrease in pH becomes less in relation to a given amount of antecedent rainfall. 3.2. Hills Creek Hills Creek catchment is larger than Firewood Creek’s and includes a substantial upland area in the adjacent foothills (Fig. 1; Powell and Martens, 2005). Hence, Hills Creek periodically receives large quantities of upland-derived fresh water with low acidity. Thus the water quality behaviour of this system is very dynamic and contrasts with that of Firewood Creek. Data was compiled from one water quality monitoring site in this catchment area – Hills Middle (Fig. 1). Water quality data for this site was analysed from February 2002 onwards, however there were large gaps in the available data set. The seasonal dynamics of mean and minimum daily pH relative to daily rainfall are shown in Fig. 5. This reveals a highly dynamic and complex system with longer periods of lower and variable EC associated with the combined interaction of upper catchment inflows and tidal exchange. While seasonal trends are evident in the EC time-series data, they are less obvious in the pH time-series data. However, aggregated mean daily pH values show a clear improvement over time between all periods (Fig. 6) which is significant at P ¼ 0.01 (Table 4). When mean daily pH is plotted against mean daily EC there is a wide scatter in the data (Fig. 7a). This is directly related to the additional influence of non-acidic, fresh water upper catchment inflows into this system. However, there are still significant positive relationships (Table 5). More importantly, a distinct temporal progression is evident in the data, whereby the slope of the linear regression between EC and pH remains similar from 2003 onward, but the y-intercept steadily increases (Table 5). Thus, while low EC
Fig. 5. Temporal changes in mean and minimum daily pH and mean daily EC in relation to daily rainfall at Hills Creek-Middle water quality monitoring station from February 2002 to August 2007. Dashed lines distinguish the temporal periods used in comparative statistical analysis.
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3.3. Soil pH and redox potential
Fig. 6. Changes in mean daily pH at Hills Creek-Middle water quality monitoring stations during select time periods from 2002 to present. The 90th, 75th, 25th and 10th percentiles are shown along with the median and outliers (open circles). Table 4 Comparison of the difference between the mean of daily pH values for select time periods at Hills Creek-Middle water quality monitoring station. Period
2003–04 2004–05 2006–07
Select soil profiles located in both the Hills Creek and Firewood Creek sub-catchments (Fig. 1, inset ‘a’ and ‘b’) were analysed for changes in field pH over time. These data (Fig. 8) demonstrate that at most of the locations studied where tidal inundation has occurred, there has been a substantial increase in soil pH (order of w2.5–3 pH units), most notably in the interval 0.2–0.7 m below the ground surface (e.g. sites 143, 148, 87, 41, 42 and 43). However, a thin layer of surface acidic soil remains at some locations (e.g. sites 143, 148). Sites above the tidal inundation zone (above the target inundation maximum of 0.5 m AHD), show minimal change in pH (e.g. sites 141, 142 and 146, Fig. 8). This is consistent with the higher elevation sites being less prone to tidal inundation and thus having decreased potential for either the development of reducing conditions and proton-consuming geochemical processes or the addition of marine derived alkalinity. Soil redox data were examined from two stations situated in the Hills Creek sub-catchment with a measurement record spanning 2001–2007. Both of these sites display substantial lowering of mean annual pe, with the decreases also following an approximate temporal progression (Fig. 9). Decreases appear to be most pronounced in the upper 0.6 m of the soil profile, though considerable seasonal variability is still evident. These trends accord with those reported by Hicks et al. (2002) and Smith et al. (2004) and are consistent with expected behaviour following tidal inundation.
Differencea 2002
2003–04
2004–05
** ** **
– ** **
– – **
a Unequal variance t-test on rank transformed data, *P < 0.05, **P < 0.01, n.s., not significant.
events continue to occur, the magnitude of associated acidification appears to have decreased between 2002 and 2007. There are also significant negative linear relationships between TWR and pH (i.e. increasing antecedent rainfall results in decreasing pH, Fig. 7b). Again, there is a wide scatter in the data directly related to the additional influence of non-acidic upper catchment inflows into this system. However, a clear progressive change over time is evident in the slope of the regression lines (Fig. 7b), with the magnitude of pH decline in relation to a given quantity of antecedent rainfall decreasing steadily from 2002 to 2007 (Table 5).
4. Discussion Analysis of the water quality data demonstrates there has been a substantial improvement in drainage water quality at the study site, evident by a combination of the following: A temporal progression of statistically significant increases in the mean pH of drainage water at multiple sites. Changes in significant relationships between antecedent rainfall and drainage water pH at multiple sites (i.e. less acidity per given amount of rainfall). These changes indicate a progressive improvement in water quality over time. Changes in significant relationships between drainage water pH and EC at multiple sites (i.e. fresh events increasingly less acidic). These changes indicate a progressive improvement in water quality with time. These data, combined with observations of decreasing soil redox potential and increases in the pH of inundated soils over time,
Fig. 7. Correlation between a) mean daily EC and mean daily pH and b) time-weighted rainfall (TWR) and mean daily pH at Hills Creek-Middle water quality monitoring station during select time periods. Linear regression, dashed lines are 99% confidence intervals.
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Table 5 Comparison of linear regression equations and P for mean daily EC (x) and mean daily pH (y) and TWR (x) and mean daily pH (y), for select time periods at Hills Creek-Middle water quality monitoring station. Period
2002 2003 2004–05 2006–07 a
Hills Creek – Middle, x ¼ EC
Hills Creek – Middle, x ¼ TWR
y ¼ ax þ b
Pa
y ¼ ax þ b
Pa
y ¼ 0.054x þ 5.17 y ¼ 0.017x þ 5.60 y ¼ 0.014x þ 6.01 y ¼ 0.014x þ 6.47
** ** ** *
y ¼ 0.007x þ 6.10 y ¼ 0.004x þ 6.30 y ¼ 0.002x þ 6.47 y ¼ 0.002x þ 6.85
** ** ** **
Using an autoregressive exact maximum likelihood model, * P < 0.05, **P < 0.01.
strongly indicate that the remediation strategy of lime assisted tidal exchange is working successfully to date by: a) allowing regular in-stream buffering, neutralisation and dilution with marine derived tidal water and, b) initiating fundamental shifts in soil geochemistry via reflooding of previously drained and acidified soils. The time-series data demonstrate that the remediation strategy is well advanced towards meeting the primary management goal of having ‘‘.water of acceptable quality (pH > 6) exiting the site, in all
Fig. 8. Changes in soil field pH over time at sites in the sub-catchments of Hills Creek (87, 141, 142, 143, 146 and 148) and Firewood Creek (41, 42 and 43). Error bars are standard deviation of replicate cores. Depth at Hills Creek sites is based on an estimate of surface elevation derived from a digital elevation model (Smith et al., 2003).
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Fig. 9. Changes in soil redox status from 2001 to 2006–07 at two locations, Hills Peat and site 69, both situated in Hills Creek sub-catchment a), b) is Hills Peat and c), d) is site 69. Depth is below ground surface, data points are means and error bars are standard deviation.
seasonal conditions, under a self-managed tidal regime without lime augmentation’’ (Smith et al., 2004, pp. 3). The encouragement of reducing conditions in the soil by reintroduction of tidal inundation is highly likely to favour the reformation of various Fe(II)-sulfide minerals, including pyrite and monosulfides. Both are known to reform in CLASS landscapes due to seasonal shifts in hydrology or the formation of localised, reducing sub-environments (e.g. Bush and Sullivan, 1997). If the current remediation strategy was halted and regular tidal inundation once again excluded from the landscape, then any reformed sulfides may undergo oxidation, thus producing acidity and degrading surface water quality (Burton et al., 2006b). Therefore the current remediation strategy can be considered a one-way process that cannot be easily reversed without serious environmental consequences. This also highlights the need for a detailed assessment of the magnitude, forms and spatial distribution of any Fe(II)-sulfide minerals that have reformed in the landscape as a consequence of the remediation strategy. Further detailed investigation of sediment and porewater geochemistry is also warranted to assess the relative importance of reductive geochemical processes and examine their relationship with the observed changes in water quality. The estimated total quantity of acidity generated at the study site following diking and drainage has been reported to be 1.5 109 mol Hþ, with an average annual acidity loss from the soil profile in the order of 7.0 105 mol Hþ ha1 y1 (Hicks et al., 1999). The total amount of hydrated lime applied to waterways at the
study site since the start of the remediation strategy is w150 tonnes, which equates to a theoretical neutralisation capacity of 4.05 106 mol Hþ. Whilst the influence of hydrated lime additions on neutralising acidity and raising stream water pH can be seen at the level of individual acidification events (Barry et al., 2003), its overall significance to the remediation of the site is less clear. A first order estimate of the acid neutralisation capacity of tidal exchange waters can be calculated by combining estimates of tidal prism volumes (based on a digital elevation model and hydrological modelling of the site, see Smith et al., 2003) and an assumed marine alkalinity of 2.2 mmol L1 (based on seawater). This analysis reveals that the acid neutralisation capacity of tidal exchange waters in Hills Creek and Firewood Creek are in the order of 2–4 108 mol Hþ y1. This is several orders of magnitude greater than the capacity of the added hydrated lime and strongly suggests that tidal exchange is likely to have had a far more significant overall effect. These data have international significance for the development and application of low-cost technologies to remediate coastal ASS. Whilst previous studies have demonstrated improvements in water quality using other remediation techniques, some of these techniques can require costly infrastructure (Indraratna et al., 2005) or regular monitoring and active intervention (Green et al., 2006), or are only partially effective (Johnston et al., 2005a). This study clearly demonstrates that controlled marine tidal inundation can be a powerful and effective landscape-scale technique for reducing the acidity in waterways of severely acidified CLASS landscapes.
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5. Conclusions The data demonstrates that the re-introduction of tidal exchange and regular tidal inundation of acid sulfate soils is improving water quality within the treated areas. Less acidity is being exported via waterways, soil acidity has decreased and the soil geochemistry appears to be changing in favour of reductive, acid-consuming geochemical processes. While it is likely that iron and sulfate reduction are now important geochemical processes in the tidally inundated soils, further investigation of sediment and porewater geochemistry is required to assess the relative importance of these processes and examine their relationship with the changes in water quality. In particular, further research on the tidally inundated soils is needed to, a) determine the magnitude, rates and pathways of internal alkalinity generation, b) examine the extent and forms of solid phase iron (II)-sulfide reformation, c) assess the extent of iron cycling associated with the reductive dissolution of Fe III minerals/ oxidation of aqueous ferrous iron and d) assess the significance of these geochemical processes for mobilisation and sequestration of trace metals and the long-term rehabilitation of the site. Acknowledgements Funding for this project was provided by the Cooperative Research Centre for Contamination Assessment and Remediation of the Environment (project no. 50773), the Queensland Department of Natural Resources and Water and the Australian Government Natural Heritage Trust Coastal Acid Sulfate Soils Program. The authors thank several anonymous reviewers for their constructive comments on the manuscript. References ¨ sterholm, P., Ba¨rlund, I., Tattari, S., 2007. Hydrochemical effects of Åstro¨m, M., O surface liming, controlled drainage and lime-filter drainage on boreal acid sulfate soils. Water, Air and Soil Pollution 179, 107–116. Barry, E.V., Ahern, C.R., Martens, M.A., Hopgood, G.L., Smith, C.D., 2003. Surface water monitoring and treatment. 2003. In: Smith, C.D., Martens, M.A., Ahern, C.R., Eldershaw, V.J., Powell, B., Barry, E.V., Hopgood, G.L., Watling, K.M. (Eds.), Demonstration of Management and Rehabilitation of Acid Sulfate Soils at East Trinity: Technical Report. Department of Natural Resources and Mines, Indooroopilly, Queensland, Australia, pp. 176–214. Berner, R.A., 1984. Sedimentary pyrite formation: an update. Geochimica et Cosmochimica Acta 48, 605–615. Blunden, B.G., Indraratna, B., 2000. Evaluation of surface and groundwater management strategies for drained sulfidic soil using numerical simulation models. Australian Journal of Soil Research 38, 569–590. van Breemen, N., 1975. Acidification and deacidification of coastal plain soils as a result of periodic flooding. Journal of the Soil Science Society of America 39, 1153–1157. van Breemen, N., 1993. Environmental aspects of acid sulphate soils. In: Dent, D., van Mensvoort, M.E.F. (Eds.), Selected Papers of the Ho Chi Minh City Symposium on Acid Sulfate Soils, Vietnam. International Institute for Land Reclamation and Improvement, Wageningen, Netherlands, pp. 391–402. Burton, E.D., Bush, R.T., Sullivan, L.A., 2006a. Sedimentary iron geochemistry in acidic waterways associated with coastal lowland acid sulfate soils. Geochimica et Cosmochimica Acta 70, 5455–5468. Burton, E.D., Bush, R.T., Sullivan, L.A., 2006b. Acid-volatile sulfide oxidation in coastal floodplain drains: iron-sulfur cycling and effects on water quality. Environmental Science & Technology 40, 1217–1222. Burton, E.D., Bush, R.T., Sullivan, L.A., Mitchell, D.R.G., 2007. Reductive transformation of iron and sulfur in schwertmannite-rich accumulations associated with acidified coastal lowlands. Geochimica et Cosmochimica Acta 71, 4456–4473. Bush, R.T., Sullivan, L.A., 1997. Morphology and behaviour of greigite from a Holocene sediment in Eastern Australia. Australian Journal of Soil Research 35, 853–861.
Cook, F.J., Hicks, W., Gardner, E.A., Carlin, G.D., Froggatt, D.W., 2000. Export of acidity in drainage water from acid sulfate soils. Marine Pollution Bulletin 41, 319–326. Cook, F.J., Rassam, D.W., 2002. An analytical model for predicting water table dynamics during drainage and evaporation. Journal of Hydrology 263, 105–113. Conover, W.J., 1998. Practical Nonparametric Statistics, third ed. Wiley, New York, 584 pp. Dent, D., 1986. Acid Sulphate Soils: a Baseline for Research and Development. Publication No. 39. International Institute for Land Reclamation and Improvement, Wageningen, Netherlands, 204 pp. Green, R., Waite, T.D., Melville, M.D., 2006. Treatment of acid sulfate soil drainage by direct application of alkaline reagents. Water, Air and Soil Pollution 178, 59–68. Haraguchi, A., 2007. Effect of sulfuric acid discharge on river water chemistry in peat swamp forests in central Kalimantan, Indonesia. Limnology 8, 175–182. Hicks, W.S., Bowman, G.M., Fitzpatrick, R.W., 1999. East Trinity Acid Sulfate Soils Part 1: Environmental Hazards. CSIRO Land and Water Technical Report 14/99, Australia, 79 pp. Hicks, W., Bowman, G., Fitzpatrick, R., 2002. The geochemistry of Australian tropical acid sulfate soils and their environmental hazard. In: 17th World Congress of Soil Science, August 2002, Book of Abstracts: V Bangkok, Thailand. International Union of Soil Science, pp. 14–21. Indraratna, B., Glamore, W., Tularam, G.A., 2002. The effects of tidal buffering on acid sulphate soil environments in coastal areas of NSW. Geotechnical and Geological Engineering 20, 181–199. Indraratna, B., Golab, A., Glamore, W., Blunden, B., 2005. Acid sulphate soil remediation techniques on the Shoalhaven River Floodplain, Australia. Quarterly Journal of Engineering Geology and Hydrogeology 38, 129–142. Johnston, S.G., Slavich, P.G., Hirst, P., 2004a. The acid flux dynamics of two artificial drains in acid sulfate soil backswamps on the Clarence River floodplain, Australia. Australian Journal of Soil Research 42, 623–637. Johnston, S.G., Slavich, P.G., Hirst, P., 2004b. The effects of a weir on reducing acid flux from a drained coastal acid sulfate soil backswamp. Agricultural Water Management 69, 43–67. Johnston, S.G., Slavich, P.G., Hirst, P., 2005a. The effects of controlled tidal exchange on improving drainage water quality in acid sulfate soil backswamps. Agricultural Water Management 73, 87–111. Johnston, S.G., Slavich, P.G., Hirst, P., 2005b. Changes in surface water quality after inundation of acid sulfate soils of different vegetation cover. Australian Journal of Soil Research 43, 1–12. Macdonald, B.C.T., Smith, J., Keene, A.F., Tunks, M., Kinsela, A., White, I., 2004. Impacts of runoff from sulfuric soils on sediment chemistry in an estuarine lake. Science of the Total Environment 329, 115–130. Minh, L.Q., Tuong, T.P., van Mensvoort, M.E.F., Bouma, J., 1997. Contamination of surface water as affected by land use in acid sulfate soils in the Mekong River Delta, Vietnam. Agriculture Ecosystems and Environment 61, 19–27. Ostrom, C.W., 1990. Time Series Analysis: Regression Techniques. Sage, Beverly Hills, 96 pp. Ponnamperuma, F.N., 1972. The chemistry of submerged soils. Advances in Agronomy 24, 29–96. Portnoy, J.W., Giblin, A.E., 1997. Biogeochemical effects of sea water restoration to diked salt marshes. Ecological Applications 7, 1054–1063. Powell, B., Martens, M., 2005. A review of acid sulfate soil impacts, actions and policies that impact on water quality in the Great Barrier Reef catchments, including a case study on remediation at East Trinity. Marine Pollution Bulletin 51, 149–164. Russell, D.J., Helmke, S.A., 2002. Impacts of acid leachate on water quality and fisheries resources of a coastal creek in northern Australia. Marine and Freshwater Research 53, 19–33. Sammut, J., White, I., Melville, M.D., 1996. Acidification of an estuarine tributary in Eastern Australia due to drainage of acid sulfate soils. Marine and Freshwater Research 47, 669–684. Smith, C.D., Martens, M.A., Ahern, C.R., Eldershaw, V.J., Powell, B., Barry, E.V., Hopgood, G.L., Watling, K.M., 2003. Demonstration of Management and Rehabilitation of Acid Sulfate Soils at East Trinity: Technical Report. Department of Natural Resources and Mines, Indooroopilly, Queensland, Australia, 335 pp. Smith, C.D., Martens, M.A., McElnea, A.E., Powell, B., 2004. East Trinity Acid Sulfate Soil Remediation Action Plan (ASSRAP) Report 2003–2004. Department of Natural Resources and Mines, Indooroopilly, Queensland, Australia, 49 pp. ¨ sterholm, P., 2002. Comparison of the metal content in Sundstrom, R., Åstro¨m, M., O acid sulfate soil runoff and industrial effluents in Finland. Environmental Science & Technology 36, 4269–4272. White, I., Melville, M.D., Wilson, B.P., Sammut, J., 1997. Reducing acidic discharges from coastal wetlands in eastern Australia. Wetlands Ecology and Management 5, 55–72. Wilson, B.P., White, I., Melville, M.D., 1999. Floodplain hydrology, acid discharge and change in water quality associated with a drained acid sulfate soil. Marine and Freshwater Research 50, 149–157.
Contents lists available at ScienceDirect
Estuarine, Coastal and Shelf Science journal homepage: www.elsevier.com/locate/ecss
Changes in water quality following tidal inundation of coastal lowland acid sulfate soil landscapes Scott G. Johnston a, *, Richard T. Bush a, Leigh A. Sullivan a, Edward D. Burton a, Douglas Smith b, Michelle A. Martens b, Angus E. McElnea b, Col R. Ahern b, Bernard Powell b, Luisa P. Stephens b, Steve T. Wilbraham b, Simon van Heel b a b
Centre for Acid Sulfate Soil Research, Southern Cross Geoscience, Southern Cross University, Lismore, NSW 2480, Australia Department of Natural Resources and Water, Block C, 80 Meiers Road, Indooroopilly, Qld 4068, Australia
a r t i c l e i n f o
a b s t r a c t
Article history: Available online 18 November 2008
This study examines the remediation of surface water quality in a severely degraded coastal acid sulfate soil landscape. The remediation strategy consisted of partial restoration of marine tidal exchange within estuarine creeks and incremental tidal inundation of acidified soils, plus strategic liming of drainage waters. Time-series water quality and climatic data collected over 5 years were analysed to assess changes in water quality due to this remediation strategy. A time-weighted rainfall function (TWR) was generated from daily rainfall data to integrate the effects of antecedent rainfall on shallow groundwater levels in a way that was relevant to acid export dynamics. Significant increases in mean pH were evident over time at multiple monitoring sites. Regression analysis at multiple sites revealed a temporal progression of change in significant relationships between mean daily electrical conductivity (EC) vs. mean daily pH, and TWR vs. mean daily pH. These data demonstrate a substantial decrease over time in the magnitude of creek acidification per given quantity of antecedent rainfall. Data also show considerable increase in soil pH (2–3 units) in formerly acidified areas subject to tidal inundation. This coincides with a decrease in soil pe, indicating stronger reducing conditions. These observations suggest a fundamental shift has occurred in sediment geochemistry in favour of proton-consuming reductive processes. Combined, these data highlight the potential effectiveness of marine tidal inundation as a landscapescale acid sulfate soil remediation strategy. Ó 2008 Elsevier Ltd. All rights reserved.
Keywords: sea-water inundation tidal wetland estuary time-series analysis floodgate
1. Introduction Drainage and reclamation of low lying coastal sulfidic sediments for agricultural purposes is a widespread practice that has potential to cause substantial acidification and degradation of estuarine water quality (e.g. Dent, 1986; Minh et al., 1997; White et al., 1997). The international literature is replete with examples of such projects leading to acidified waterways, mobilisation of iron, aluminium and trace metals, along with other forms of environmental degradation (e.g. Sammut et al., 1996; Wilson et al., 1999; Sundstrom et al., 2002; Johnston et al., 2004a; Macdonald et al., 2004; Haraguchi, 2007). A variety of studies have examined the application of different remediation techniques to improve the water quality associated with drained coastal lowland acid sulfate soils (CLASS). Some examples of remediation techniques include partial tidal exchange
* Corresponding author. E-mail address: [email protected] (S.G. Johnston). 0272-7714/$ – see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.ecss.2008.11.002
within drains (Indraratna et al., 2002; Johnston et al., 2005a), the use of in-drain weirs or controlled drainage (Blunden and Indraratna, 2000; Johnston et al., 2004b; Åstro¨m et al., 2007), and techniques involving the use of alkaline reagents (Green et al., 2006). Tidal inundation of CLASS is a further remediation option available when their elevation is close to mean sea level. However, this strategy has rarely been applied on a large scale. Tidal inundation of acidified sediments may lead to re-establishment of proton-consuming and alkalinity-producing geochemical processes involving the decay of organic matter under sub-oxic/ anoxic conditions (e.g. Ponnamperuma, 1972; van Breemen, 1975; Portnoy and Giblin, 1997; Johnston et al., 2005b; Burton et al., 2007). In essence, establishing tidal inundation of CLASS may initiate similar diagenetic forces to those commonly found in intertidal sedimentary environments such as mangroves (i.e. high water tables, abundant sulfate and organic matter). Such conditions will stimulate upward migration of the redox boundary, favouring the reductive dissolution of Fe (III) minerals and the reduction of SO2 4 . In turn, this may lead to eventual formation of various solid phase
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Fe (II)-sulfide minerals (ie. pyrite, greigite and mackinawite) in formerly oxidised, acidified sulfuric horizons (Bush and Sullivan, 1997; Burton et al., 2006a, 2007). Any associated proton consumption and generation of alkalinity will be highly dependent upon iron and sulfur bio-mineralisation pathways. Whilst these pathways are variable and complex (Berner, 1984; van Breemen, 1993), the key reactions involving the generation of bicarbonate alkalinity can be described according to Eqs. (1) and (2), 2CH2 O þ SO2 4 /H2 S þ 2HCO3
(1)
CH2 O þ 4FeOOH þ 7Hþ /4Fe2þ þ HCO 3 þ 6H2 O
(2)
Although Fe (III) and SO2 reduction generates alkalinity, both 4 reactions produce potentially mobile aqueous species that may reoxidise and release acidity. Hence, the burial of these species via the formation of diagenetic minerals (e.g. FeS2) is a key to long-term amelioration of acidity. There are few examples of studies involving large scale reestablishment of tidal inundation in CLASS landscapes combined with an assessment of the effects of this technique upon the frequency and magnitude of acidification of adjacent surface waters. This study aims to a) analyse a medium term (w5 years) time-series water quality data set derived from waterways in a drained and reclaimed CLASS landscape that was subject to reintroduction of tidal exchange and tidal inundation and b) to identify any changes in water quality that are attributable to the above remediation strategy. 2. Materials and methods 2.1. Study site and remediation strategy The study site, East Trinity, is a Holocene sedimentary coastal plain of w800 ha located near Cairns in northern Australia (Fig. 1). The Cairns area has a tropical monsoonal climate with summerdominant rainfall. The site was diked and intensively drained for sugar cane production in the 1970s (Powell and Martens, 2005) and originally contained large areas of estuarine wetlands, including mangroves and saltmarsh. There are several estuarine creeks located at the site; Hills Creek, Magazine Creek, Firewood Creek and Georges Creek. The dike, or bund wall, effectively excluded tidal exchange from these creeks. One-way floodgates, located at the intersection of the bund wall with Hills Creek and Firewood Creek, allowed drainage waters to exit the site during the ebb tide. The exclusion of tidal exchange combined with intensive drainage led to lowering of water tables and subsequent oxidation and subsidence of sulfidic sediments in former mangrove areas (Hicks et al., 1999). This resulted in the formation of severe acid sulfate soils and seasonal export of large quantities of acidity from the site (Cook et al., 2000; Hicks et al., 2002; Russell and Helmke, 2002; Smith et al., 2003; Powell and Martens, 2005). Investigations have demonstrated that most of the actual acid sulfate soil material at the site is located below 0.5 m AHD (Australian Height Datum; 0 m AHD is wmean sea level) (Smith et al., 2003; Powell and Martens, 2005). The remediation strategy is detailed in Powell and Martens (2005) and consists mainly of controlled tidal exchange involving ‘‘.progressively and cautiously replacing the existing acidified freshwater environment with a managed tidal wetland system, by modifying and managing the infrastructure that initiated the problem some 30 years ago.’’ Key aspects of the strategy include; Incremental opening of floodgates to allow regular tidal inundation of the land surface inside the bund wall, up to a maximum elevation of approximately 0.5 m AHD.
Water quality monitoring and strategic treatment of acid drainage water with hydrated lime, [Ca(OH)2]. This was regarded as an important adjunct to the main strategy, particularly during the early stages of reintroducing tidal exchange, when the potential for mobilisation of stored acidity from the system was greatest (see Smith et al. (2003) and Powell and Martens (2005) for further details). A primary goal of the remediation strategy at East Trinity is to ‘‘have water of acceptable quality (pH > 6) exiting the site, in all seasonal conditions, under a self-managed tidal regime without lime augmentation’’ (Smith et al., 2004, pp. 3). Hills Creek has been subjected to daily tidal exchange since February 2003. The extent of regular tidal inundation within its sub-catchment was incrementally increased between 2003 and 2005, to a maximum level of approximately 0.5 m AHD. Firewood Creek was subject to partial and restricted tidal exchange commencing in mid-2002 and followed by more extensive tidal exchange from 2005 onward. Managed floodgate opening in Firewood Creek currently (i.e. 2007) allows tidal exchange to an elevation of approximately 0.25–0.3 m AHD. 2.2. Time-series water quality data The three water quality monitoring sites were chosen for detailed analysis and interrogation – Firewood Creek-Bund, Firewood Creek-Upper and Hills Creek-Middle (Fig. 1). These sites were chosen due to the availability of long duration data sets (>4 years) and the strategic and contrasting nature of their locations. The pH and EC (electrical conductivity) data were regarded as having greatest value in terms of their relevance to the goals of the remediation strategy. Data analysis focussed on these two key parameters as well as daily rainfall. Whilst pH can only be regarded as an indicator of the total acidity within waterways located in CLASS landscapes (due to the presence of acidic metal cations, e.g. Hicks et al., 1999; Cook et al., 2000), it is nonetheless a very useful proxy due to the primary control it exerts over trace metal mobility, solubility, speciation and bio-toxicity. The pH, EC and temperature were logged at 10–15 min intervals using either TPS 90FL or Hydrolab Quanta multi-parameter sondes in combination with Campbell Scientific CR10X loggers. Sensor probes were situated in the upper 1 m of the water column and were maintained and serviced regularly with intervals between calibration ranging from 1 to 2 weeks. The following general processing and collation tasks were applied to the raw time-series water quality data files for each site. Screening to identify and remove any spurious or outlier values. Screening to identify and remove data from periods that were subject to logger calibration drift. Screening to identify and remove parameter values outside the practical environmental range and/or the instrument range.
2.3. Water quality trend assessment and statistical analysis The mean daily pH and EC were calculated for each of the three sites for the duration of the assessment period (2002–2007). Mean daily values were regarded as more useful for analysis of longer term trends due to the smoothing of extreme temporal variation associated with tidal modulation and decreasing auto-correlation via aggregation. Acidification episodes in coastal floodplain waterways are typically strongly related to shallow groundwater levels and thus antecedent rainfall (e.g. Wilson et al., 1999; Cook and Rassam, 2002;
S.G. Johnston et al. / Estuarine, Coastal and Shelf Science 81 (2009) 257–266
259
Fig. 1. East Trinity study site location map showing dike/bund wall and water quality monitoring stations. Soil monitoring sites are displayed in inset ‘a’ and ‘b’.
Johnston et al., 2004a). The two main climatic drivers of the floodplain water balance which govern floodplain water tables and thus acid export are precipitation (P) and evapotranspiration (ET) (White et al., 1997). To assess the effectiveness of the remediation strategy it was necessary to interrogate the data sets for changes in relationships between rainfall, shallow groundwater levels and drainage water quality. However, there was a lack of time-series shallow groundwater level data for the project site during the assessment period. Thus, there was a need for a quantitative measure of rainfall, beyond that of simple daily data, which adequately captured the influence of antecedent rainfall in a manner meaningful to acid export. For this purpose a timeweighted rainfall (TWR) function was developed using Eq. (3).
TWRðxÞ ¼
30 X
½ðpx1 þ /px4 Þ þ ðpx5 þ /px8 Þ=2
x¼1
þ ðpx9 þ /px12 Þ=3 þ ðpx13 þ /px16 Þ=4 þ ðpx17 þ /px30 Þ=5
ð3Þ
where at time (t) ¼ x, px1 is the daily rainfall at t ¼ x 1 day, and px2 is the daily rainfall at t ¼ x 2 days etc., generating
a weighted rainfall total over a prior 30 day period. The denominators in Eq. (3) follow a simple arithmetic progression. This pattern of weighting is based on the recognition that shallow ground water levels in coastal floodplain environments rise rapidly in response to rainfall and decay slowly over time (typically in the order of months) in response to losses from drainage and evapotranspiration, thus producing a typical sawtooth pattern (e.g. Johnston et al., 2004b). Hence, rainfall events most closely preceding a given date of water quality monitoring are likely to have a greater influence on groundwater levels and therefore drainage water quality than more temporally distant rainfall events. However, temporally distant rainfall events will still have a diminished influence on antecedent shallow groundwater levels, as some of that rainfall remains in storage within the soil profile. The TWR broadly mimics such a pattern. While the precise choice of time intervals and associated weighting is somewhat arbitrary, the TWR provides a far more integrated approach to accounting for antecedent rainfall effects than either daily precipitation values alone or a running mean of daily precipitation values calculated over a set time interval. All statistical analysis was conducted using SPSS 15 software. Time-series water quality data is typically auto-correlated and
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Fig. 2. Temporal changes in mean daily pH and EC in relation to daily rainfall at Firewood-Bund and Firewood-Upper water quality monitoring stations from January 2002 to August 2007. Dashed lines distinguish the temporal periods used in comparative statistical analysis.
often non-normally distributed which can inhibit the reliability of ordinary parametric statistical analysis. Linear regression analysis was used to determine whether there was evidence of temporal changes in the nature of relationships (regression coefficients) between key parameters. Auto-correlation can increase the likelihood of type 1 error when assessing relationship significance using ordinary least squares techniques. Therefore a first order autoregressive process using an exact maximum likelihood model was used to assess the significance of those relationships (Ostrom, 1990). Residuals were then analysed to ensure there had been adequate removal of auto-correlation errors and ensure nonviolation of model assumptions regarding normality and variance. These analyses were applied to mean daily pH in relation to both mean daily EC and TWR at each site for discrete time intervals. Rather than create time intervals based on calendar years, a seasonal division was used where possible, thus allowing a more meaningful comparison of successive wet seasons. For the
Firewood Creek data sets, this interval was from 1st September to August 31st the following year. The Hills Creek-middle data set was divided into four time intervals based on periods of data availability. The mean pH for these time intervals were compared using an unequal variance version of the t-test after rank transformation to account for non-normal distribution (Conover, 1998). 2.4. Rainfall data Rainfall data were derived from a weather station (automatic tipping bucket rain gauge) located at the Firewood-Bund wall and a manual rain gauge located at the House site (Fig. 1). Data from the automatic weather station was used where possible. However, after 2005 the automatic weather station failed. From 2005 onwards all rainfall data is derived from the manual rain gauge. Due to concern that this may bias the TWR, rainfall data from the two sites were compared for a w4 month period during 2004. While the House
Fig. 3. Changes in mean daily pH at a) Firewood Creek-Bund and b) at Firewood Creek-Upper water quality monitoring stations during select time periods from 2002 to 2006-07. The 90th, 75th, 25th and 10th percentiles are shown along with the median and outliers (open circles).
S.G. Johnston et al. / Estuarine, Coastal and Shelf Science 81 (2009) 257–266 Table 1 Comparison of the difference between the mean of daily pH values for select time periods at Firewood Creek-Bund and Firewood Creek-Upper water quality monitoring stations. Period
Differencea
261
monitoring stations located in Hills Creek sub-catchment between 2001 and 2007. Measurements were made in-situ using platinum tipped electrodes (details are described by Hicks et al., 1999) approximately every 2–6 months and data are corrected to a standard hydrogen electrode.
Firewood Creek-Bund/Firewood Creek-Upper
2002–03 2003–04 2004–05 2005–06 2006–07
2002
2002–03
2003–04
2004–05
2005–06
**/n.s. n.s./** **/** **/** **/**
– **/** **/** **/** **/**
– – **/** **/** **/**
– – – **/* **/n.s.
– – – – **/n.s.
a Unequal variance t-test on rank transformed data, * P < 0.05, ** P < 0.01, n.s., not significant.
site tends to have slightly higher registrations, the two sites exhibit a strong linear relationship (r2 ¼ 0.86, y ¼ 1.247x þ 10.9, where x ¼ Bund wall). However, in order to reduce the likelihood of this change in rainfall data sources introducing a temporal bias to the TWR, the TWR calculated on the basis of Bund wall weather station data was transformed according to the above relationship prior to further analysis.
2.5. Soil pH and redox potential Soil pH data is based on field measurements derived from repeat sampling at set locations in both Hills Creek and Firewood Creek sub-catchments. Soils were collected via gouge augers or push corers and analysis conducted immediately by inserting a freshly calibrated probe either directly into moist soil or a 1:1 fresh soil/ water solution. Redox potential was measured at two fixed
3. Results 3.1. Firewood Creek Firewood Creek catchment is relatively small (w2 km2) and is confined to the coastal floodplain. A large proportion of the Firewood Creek catchment contained highly acidic soils (pH > 3.5) (Smith et al., 2003). Data was compiled from two water quality monitoring sites in this catchment area, Firewood-Bund and Firewood-Upper (Fig. 1). The seasonal dynamics in mean daily pH for both sites relative to daily rainfall are shown in Fig. 2. During the dry season there were long periods of relatively high (w6.5–8) and comparatively stable pH coupled with high EC. This contrasts with alternating periods of low (<4.5) and highly dynamic pH and low EC associated with wet seasons. The mean daily pH values at both monitoring sites display a clear increase with time whilst the variability of mean daily pH values decreases (Fig. 3). These changes in aggregated mean daily pH values are significant for most periods at P ¼ 0.01 (Table 1). Firewood-Upper generally had lower mean values than the Firewood-Bund site, which accords with its greater distance from the estuary, the source of marine buffering. There are significant positive linear relationships between EC and pH at both sites, with distinct temporal groupings evident in the data (Fig. 4a and b). The slope of the relationships between EC and pH change progressively over time (Table 2). Low EC events continue to occur, but they become less strongly associated with
Fig. 4. Correlation between mean daily EC and mean daily pH at a) Firewood Creek-Bund and b) at Firewood Creek-Upper water quality monitoring stations during select time periods. Correlation between time-weighted rainfall (TWR) and mean daily pH at c) Firewood Creek-Bund and d) at Firewood Creek-Upper water quality monitoring stations during select time periods. Linear regression, dashed lines are 99% confidence intervals (Note: 1997–98 data is based on spot monitoring by Hicks et al., 1999).
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Table 2 Comparison of linear regression equations and P for mean daily EC (x) and mean daily pH (y) for select time periods at Firewood Creek-Bund and Firewood CreekUpper water quality monitoring stations. Period
2002 2002–03 2003–04 2004–05 2005–06 2006–07 a
Firewood Creek-Bund
Firewood Creek-Upper a
y ¼ ax þ b
P
y ¼ ax þ b
Pa
y ¼ 0.081x þ 2.84 y ¼ 0.067x þ 3.79 y ¼ 0.031x þ 5.35 y ¼ 0.044x þ 5.56 y ¼ 0.029x þ 6.01 y ¼ 0.029x þ 5.98
** ** ** ** ** **
– y ¼ 0.079x þ 3.10 y ¼ 0.090x þ 2.63 y ¼ 0.036x þ 5.67 y ¼ 0.030x þ 5.91 y ¼ 0.015x þ 6.56
** ** ** ** **
Using an autoregressive exact maximum likelihood model, * P < 0.05, ** P < 0.01.
Table 3 Comparison of linear regression equations and P for for TWR (x) and mean daily pH (y) for select time periods at Firewood Creek-Bund and Firewood Creek-Upper water quality monitoring stations. Period
2002 2002–03 2003–04 2004–05 2005–06 2006–07 a
Firewood Creek-Bund
Firewood Creek-Upper a
y ¼ ax þ b
P
y ¼ 0.016x þ 7.58 y ¼ 0.008x þ 7.62 y ¼ 0.003x þ 7.06 y ¼ 0.004x þ 7.95 y ¼ 0.002x þ 7.56 y ¼ 0.002x þ 7.68
** ** ** * ** **
y ¼ ax þ b
Pa
y ¼ 0.020x þ 7.53 y ¼ 0.011x þ 7.60 y ¼ 0.008x þ 7.57 y ¼ 0.003x þ 7.65 y ¼ 0.002x þ 7.47 y ¼ 0.002x þ 7.49
** ** ** ** ** **
Using an autoregressive exact maximum likelihood model, * P < 0.05, **P < 0.01.
decreases in pH. This trend shift is particularly evident at the Firewood-Upper site (Fig. 4b). Spot monitoring water quality data collected from Firewood Creek in 1997–1998 prior to the implementation of tidal exchange (derived from Hicks et al., 1999), is consistent with the EC/pH signature observed during acid export events in 2002–2003, during the early stages of remediation (Fig. 4b).
There are also significant negative linear relationships between TWR and pH at both sites (i.e. increasing antecedent rainfall ¼ decreasing pH) (Fig. 4c and d). A clear progressive change over time is evident in the slope of relationships between TWR and mean daily pH at both locations (Table 3), whereby the decrease in pH becomes less in relation to a given amount of antecedent rainfall. 3.2. Hills Creek Hills Creek catchment is larger than Firewood Creek’s and includes a substantial upland area in the adjacent foothills (Fig. 1; Powell and Martens, 2005). Hence, Hills Creek periodically receives large quantities of upland-derived fresh water with low acidity. Thus the water quality behaviour of this system is very dynamic and contrasts with that of Firewood Creek. Data was compiled from one water quality monitoring site in this catchment area – Hills Middle (Fig. 1). Water quality data for this site was analysed from February 2002 onwards, however there were large gaps in the available data set. The seasonal dynamics of mean and minimum daily pH relative to daily rainfall are shown in Fig. 5. This reveals a highly dynamic and complex system with longer periods of lower and variable EC associated with the combined interaction of upper catchment inflows and tidal exchange. While seasonal trends are evident in the EC time-series data, they are less obvious in the pH time-series data. However, aggregated mean daily pH values show a clear improvement over time between all periods (Fig. 6) which is significant at P ¼ 0.01 (Table 4). When mean daily pH is plotted against mean daily EC there is a wide scatter in the data (Fig. 7a). This is directly related to the additional influence of non-acidic, fresh water upper catchment inflows into this system. However, there are still significant positive relationships (Table 5). More importantly, a distinct temporal progression is evident in the data, whereby the slope of the linear regression between EC and pH remains similar from 2003 onward, but the y-intercept steadily increases (Table 5). Thus, while low EC
Fig. 5. Temporal changes in mean and minimum daily pH and mean daily EC in relation to daily rainfall at Hills Creek-Middle water quality monitoring station from February 2002 to August 2007. Dashed lines distinguish the temporal periods used in comparative statistical analysis.
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3.3. Soil pH and redox potential
Fig. 6. Changes in mean daily pH at Hills Creek-Middle water quality monitoring stations during select time periods from 2002 to present. The 90th, 75th, 25th and 10th percentiles are shown along with the median and outliers (open circles). Table 4 Comparison of the difference between the mean of daily pH values for select time periods at Hills Creek-Middle water quality monitoring station. Period
2003–04 2004–05 2006–07
Select soil profiles located in both the Hills Creek and Firewood Creek sub-catchments (Fig. 1, inset ‘a’ and ‘b’) were analysed for changes in field pH over time. These data (Fig. 8) demonstrate that at most of the locations studied where tidal inundation has occurred, there has been a substantial increase in soil pH (order of w2.5–3 pH units), most notably in the interval 0.2–0.7 m below the ground surface (e.g. sites 143, 148, 87, 41, 42 and 43). However, a thin layer of surface acidic soil remains at some locations (e.g. sites 143, 148). Sites above the tidal inundation zone (above the target inundation maximum of 0.5 m AHD), show minimal change in pH (e.g. sites 141, 142 and 146, Fig. 8). This is consistent with the higher elevation sites being less prone to tidal inundation and thus having decreased potential for either the development of reducing conditions and proton-consuming geochemical processes or the addition of marine derived alkalinity. Soil redox data were examined from two stations situated in the Hills Creek sub-catchment with a measurement record spanning 2001–2007. Both of these sites display substantial lowering of mean annual pe, with the decreases also following an approximate temporal progression (Fig. 9). Decreases appear to be most pronounced in the upper 0.6 m of the soil profile, though considerable seasonal variability is still evident. These trends accord with those reported by Hicks et al. (2002) and Smith et al. (2004) and are consistent with expected behaviour following tidal inundation.
Differencea 2002
2003–04
2004–05
** ** **
– ** **
– – **
a Unequal variance t-test on rank transformed data, *P < 0.05, **P < 0.01, n.s., not significant.
events continue to occur, the magnitude of associated acidification appears to have decreased between 2002 and 2007. There are also significant negative linear relationships between TWR and pH (i.e. increasing antecedent rainfall results in decreasing pH, Fig. 7b). Again, there is a wide scatter in the data directly related to the additional influence of non-acidic upper catchment inflows into this system. However, a clear progressive change over time is evident in the slope of the regression lines (Fig. 7b), with the magnitude of pH decline in relation to a given quantity of antecedent rainfall decreasing steadily from 2002 to 2007 (Table 5).
4. Discussion Analysis of the water quality data demonstrates there has been a substantial improvement in drainage water quality at the study site, evident by a combination of the following: A temporal progression of statistically significant increases in the mean pH of drainage water at multiple sites. Changes in significant relationships between antecedent rainfall and drainage water pH at multiple sites (i.e. less acidity per given amount of rainfall). These changes indicate a progressive improvement in water quality over time. Changes in significant relationships between drainage water pH and EC at multiple sites (i.e. fresh events increasingly less acidic). These changes indicate a progressive improvement in water quality with time. These data, combined with observations of decreasing soil redox potential and increases in the pH of inundated soils over time,
Fig. 7. Correlation between a) mean daily EC and mean daily pH and b) time-weighted rainfall (TWR) and mean daily pH at Hills Creek-Middle water quality monitoring station during select time periods. Linear regression, dashed lines are 99% confidence intervals.
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Table 5 Comparison of linear regression equations and P for mean daily EC (x) and mean daily pH (y) and TWR (x) and mean daily pH (y), for select time periods at Hills Creek-Middle water quality monitoring station. Period
2002 2003 2004–05 2006–07 a
Hills Creek – Middle, x ¼ EC
Hills Creek – Middle, x ¼ TWR
y ¼ ax þ b
Pa
y ¼ ax þ b
Pa
y ¼ 0.054x þ 5.17 y ¼ 0.017x þ 5.60 y ¼ 0.014x þ 6.01 y ¼ 0.014x þ 6.47
** ** ** *
y ¼ 0.007x þ 6.10 y ¼ 0.004x þ 6.30 y ¼ 0.002x þ 6.47 y ¼ 0.002x þ 6.85
** ** ** **
Using an autoregressive exact maximum likelihood model, * P < 0.05, **P < 0.01.
strongly indicate that the remediation strategy of lime assisted tidal exchange is working successfully to date by: a) allowing regular in-stream buffering, neutralisation and dilution with marine derived tidal water and, b) initiating fundamental shifts in soil geochemistry via reflooding of previously drained and acidified soils. The time-series data demonstrate that the remediation strategy is well advanced towards meeting the primary management goal of having ‘‘.water of acceptable quality (pH > 6) exiting the site, in all
Fig. 8. Changes in soil field pH over time at sites in the sub-catchments of Hills Creek (87, 141, 142, 143, 146 and 148) and Firewood Creek (41, 42 and 43). Error bars are standard deviation of replicate cores. Depth at Hills Creek sites is based on an estimate of surface elevation derived from a digital elevation model (Smith et al., 2003).
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Fig. 9. Changes in soil redox status from 2001 to 2006–07 at two locations, Hills Peat and site 69, both situated in Hills Creek sub-catchment a), b) is Hills Peat and c), d) is site 69. Depth is below ground surface, data points are means and error bars are standard deviation.
seasonal conditions, under a self-managed tidal regime without lime augmentation’’ (Smith et al., 2004, pp. 3). The encouragement of reducing conditions in the soil by reintroduction of tidal inundation is highly likely to favour the reformation of various Fe(II)-sulfide minerals, including pyrite and monosulfides. Both are known to reform in CLASS landscapes due to seasonal shifts in hydrology or the formation of localised, reducing sub-environments (e.g. Bush and Sullivan, 1997). If the current remediation strategy was halted and regular tidal inundation once again excluded from the landscape, then any reformed sulfides may undergo oxidation, thus producing acidity and degrading surface water quality (Burton et al., 2006b). Therefore the current remediation strategy can be considered a one-way process that cannot be easily reversed without serious environmental consequences. This also highlights the need for a detailed assessment of the magnitude, forms and spatial distribution of any Fe(II)-sulfide minerals that have reformed in the landscape as a consequence of the remediation strategy. Further detailed investigation of sediment and porewater geochemistry is also warranted to assess the relative importance of reductive geochemical processes and examine their relationship with the observed changes in water quality. The estimated total quantity of acidity generated at the study site following diking and drainage has been reported to be 1.5 109 mol Hþ, with an average annual acidity loss from the soil profile in the order of 7.0 105 mol Hþ ha1 y1 (Hicks et al., 1999). The total amount of hydrated lime applied to waterways at the
study site since the start of the remediation strategy is w150 tonnes, which equates to a theoretical neutralisation capacity of 4.05 106 mol Hþ. Whilst the influence of hydrated lime additions on neutralising acidity and raising stream water pH can be seen at the level of individual acidification events (Barry et al., 2003), its overall significance to the remediation of the site is less clear. A first order estimate of the acid neutralisation capacity of tidal exchange waters can be calculated by combining estimates of tidal prism volumes (based on a digital elevation model and hydrological modelling of the site, see Smith et al., 2003) and an assumed marine alkalinity of 2.2 mmol L1 (based on seawater). This analysis reveals that the acid neutralisation capacity of tidal exchange waters in Hills Creek and Firewood Creek are in the order of 2–4 108 mol Hþ y1. This is several orders of magnitude greater than the capacity of the added hydrated lime and strongly suggests that tidal exchange is likely to have had a far more significant overall effect. These data have international significance for the development and application of low-cost technologies to remediate coastal ASS. Whilst previous studies have demonstrated improvements in water quality using other remediation techniques, some of these techniques can require costly infrastructure (Indraratna et al., 2005) or regular monitoring and active intervention (Green et al., 2006), or are only partially effective (Johnston et al., 2005a). This study clearly demonstrates that controlled marine tidal inundation can be a powerful and effective landscape-scale technique for reducing the acidity in waterways of severely acidified CLASS landscapes.
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5. Conclusions The data demonstrates that the re-introduction of tidal exchange and regular tidal inundation of acid sulfate soils is improving water quality within the treated areas. Less acidity is being exported via waterways, soil acidity has decreased and the soil geochemistry appears to be changing in favour of reductive, acid-consuming geochemical processes. While it is likely that iron and sulfate reduction are now important geochemical processes in the tidally inundated soils, further investigation of sediment and porewater geochemistry is required to assess the relative importance of these processes and examine their relationship with the changes in water quality. In particular, further research on the tidally inundated soils is needed to, a) determine the magnitude, rates and pathways of internal alkalinity generation, b) examine the extent and forms of solid phase iron (II)-sulfide reformation, c) assess the extent of iron cycling associated with the reductive dissolution of Fe III minerals/ oxidation of aqueous ferrous iron and d) assess the significance of these geochemical processes for mobilisation and sequestration of trace metals and the long-term rehabilitation of the site. Acknowledgements Funding for this project was provided by the Cooperative Research Centre for Contamination Assessment and Remediation of the Environment (project no. 50773), the Queensland Department of Natural Resources and Water and the Australian Government Natural Heritage Trust Coastal Acid Sulfate Soils Program. The authors thank several anonymous reviewers for their constructive comments on the manuscript. References ¨ sterholm, P., Ba¨rlund, I., Tattari, S., 2007. Hydrochemical effects of Åstro¨m, M., O surface liming, controlled drainage and lime-filter drainage on boreal acid sulfate soils. Water, Air and Soil Pollution 179, 107–116. Barry, E.V., Ahern, C.R., Martens, M.A., Hopgood, G.L., Smith, C.D., 2003. Surface water monitoring and treatment. 2003. In: Smith, C.D., Martens, M.A., Ahern, C.R., Eldershaw, V.J., Powell, B., Barry, E.V., Hopgood, G.L., Watling, K.M. (Eds.), Demonstration of Management and Rehabilitation of Acid Sulfate Soils at East Trinity: Technical Report. Department of Natural Resources and Mines, Indooroopilly, Queensland, Australia, pp. 176–214. Berner, R.A., 1984. Sedimentary pyrite formation: an update. Geochimica et Cosmochimica Acta 48, 605–615. Blunden, B.G., Indraratna, B., 2000. Evaluation of surface and groundwater management strategies for drained sulfidic soil using numerical simulation models. Australian Journal of Soil Research 38, 569–590. van Breemen, N., 1975. Acidification and deacidification of coastal plain soils as a result of periodic flooding. Journal of the Soil Science Society of America 39, 1153–1157. van Breemen, N., 1993. Environmental aspects of acid sulphate soils. In: Dent, D., van Mensvoort, M.E.F. (Eds.), Selected Papers of the Ho Chi Minh City Symposium on Acid Sulfate Soils, Vietnam. International Institute for Land Reclamation and Improvement, Wageningen, Netherlands, pp. 391–402. Burton, E.D., Bush, R.T., Sullivan, L.A., 2006a. Sedimentary iron geochemistry in acidic waterways associated with coastal lowland acid sulfate soils. Geochimica et Cosmochimica Acta 70, 5455–5468. Burton, E.D., Bush, R.T., Sullivan, L.A., 2006b. Acid-volatile sulfide oxidation in coastal floodplain drains: iron-sulfur cycling and effects on water quality. Environmental Science & Technology 40, 1217–1222. Burton, E.D., Bush, R.T., Sullivan, L.A., Mitchell, D.R.G., 2007. Reductive transformation of iron and sulfur in schwertmannite-rich accumulations associated with acidified coastal lowlands. Geochimica et Cosmochimica Acta 71, 4456–4473. Bush, R.T., Sullivan, L.A., 1997. Morphology and behaviour of greigite from a Holocene sediment in Eastern Australia. Australian Journal of Soil Research 35, 853–861.
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