Dynamics of herbicide transport and partitioning under event flow conditions in the lower Burdekin region, Australia

Marine Pollution Bulletin 65 (2012) 182–193

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Dynamics of herbicide transport and partitioning under event flow conditions in the lower Burdekin region, Australia Aaron M. Davis a,⇑, Stephen E. Lewis a, Zoë T. Bainbridge a, Lionel Glendenning a, Ryan D.R. Turner b, Jon E. Brodie a a b

Catchment to Reef Research Group, Australian Centre for Tropical Freshwater Research, James Cook University, Townsville, Queensland 4811, Australia Department of Environmental and Resource Management Block C, Level 1, 41 Dutton Park, Queensland 4102, Australia

a r t i c l e

i n f o

Keywords: Great Barrier Reef Pesticide delivery Pollutant loads Water–sediment partitioning Diuron Atrazine

a b s t r a c t This study examined the temporal variability in herbicide delivery to the Great Barrier Reef (GBR) lagoon (Australia) from one of the GBR catchment’s major sugarcane growing regions. Annual loads of measured herbicides were consistently in the order of 200+ kg. Atrazine, it’s degradate desethylatrazine, and diuron contributed approximately 90% of annual herbicide load, with early ‘first-flush’ events accounting for the majority of herbicide loads leaving the catchment. Assessment of herbicide water–sediment partitioning in flood runoff highlighted the majority of herbicides were transported in predominantly dissolved form, although a considerable fraction of diuron was transported in particulate-bound form (ca. 33%). Diuron was also the herbicide demonstrating the highest concentrations and frequency of detection in sediments collected from catchment waterways and adjacent estuarine–marine environments, an outcome aligning with previous research. Herbicide physico-chemical properties appear to play a crucial role in partitioning between water column and sediment habitat types in GBR receiving ecosystems. Crown Copyright Ó 2011 Published by Elsevier Ltd. All rights reserved.

1. Introduction 1.1. Great Barrier Reef pesticide water quality issues The degradation of certain coastal and inshore ecosystems of the Great Barrier Reef (GBR) has been attributed to the effects of land-based runoff from adjacent agricultural catchments (Brodie et al., 2007, 2012; DeVantier et al., 2006; Fabricius et al., 2005). Reducing the load of diffuse pollutants in freshwater entering the GBR lagoon constitutes one of two core objectives of the GBR Reef Water Quality Protection Plan (Anon., 2003; Reef Water Quality Protection Plan Secretariat, 2009). The effects of suspended sediment and nutrient discharge remain a long-standing theme for GBR water quality research (Devlin and Brodie, 2005; McCulloch et al., 2003), with recent studies also highlighting pesticide residues as an emergent issue for GBR ecosystem health. Pesticide residues have been documented across virtually the entire continuum of GBR associated environments including; catchment irrigation drainage systems and waterways (Davis et al., 2008, 2011; Hunter et al., 2001; Mitchell et al., 2005; Müller et al., 2000; Packett et al., 2009; Stork et al., 2008); estuaries (Duke et al., 2005); nearshore marine habitats (Haynes et al., 2000a); and coastal marine environ-

⇑ Corresponding author. Tel.: +61 7 4781 5989; fax: +61 7 4781 5589. E-mail address: [email protected] (A.M. Davis).

ments (Kennedy et al., 2012; Lewis et al., 2009, 2012; Shaw and Müller, 2005; Shaw et al., 2010). The potential deleterious impact of pesticide runoff, particularly herbicides, has been highlighted across an array of keystone GBR marine organisms including seagrass, corals and algae (Cantin et al., 2007; Haynes et al., 2000b; Jones et al., 2003; Magnusson et al., 2010). Recent estimates suggest at least 30,000 kg/yr of PSII herbicides (atrazine, ametryn, hexazinone, diuron, simazine and tebuthiuron) are exported to the GBR World Heritage Area (Kroon et al., 2012; Waterhouse et al., 2012). Considerable knowledge gaps still, however, exist regarding aspects of pesticide delivery to the GBR. These include transformations and partitioning during transport in streams of particular pesticides (i.e. dissolved versus particulate), and temporal variability in loads associated with the relatively high climatic (hydrologic) variability that characterise many GBR catchments (Brodie et al., 2012; Bainbridge et al., 2009b). Studies targeting herbicide residues across the GBR have documented considerable variability in the presence of a range of compounds across different habitat types. Several herbicide residues have been detected in benthic sediments sampled from irrigation drains and channels in GBR catchments, with diuron being the dominant herbicide in terms of frequency of detection and average concentrations (Müller et al., 2000; Stork et al., 2008). Diuron is similarly the dominant residue detected from inter- and sub-tidal sediments across the GBR (Duke et al., 2005; Haynes et al.,

0025-326X/$ - see front matter Crown Copyright Ó 2011 Published by Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2011.08.025

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2000a). A broad range of herbicides have been detected in catchment waterways during flood events (Bainbridge et al., 2009a; Davis et al., 2008; Lewis et al., 2009; Mitchell et al., 2005; Packett et al., 2009; Rohde et al., 2008). The herbicides commonly detected in the GBR lagoon waters are relatively limited, but include diuron, atrazine, hexazinone, ametryn, simazine and tebuthiuron (Davis et al., 2008; Kapernick et al., 2007; Kennedy et al., 2012; Lewis et al., 2009; Shaw and Müller, 2005; Shaw et al., 2010). Pesticides may be transported in the aqueous (dissolved) phase or particulate phase in runoff depending on their water solubility and sorption to soil (Rice et al., 2004). The importance of physico-chemical properties on the inherent transport potential and ultimate environmental fate of particular herbicides have been long appreciated (Hargreaves et al., 1999; Wauchope, 1978; Willis and McDowell, 1982). Due to conservative mixing behaviours observed in flood plumes, Lewis et al. (2009) suggested many of these key herbicides were at least initially transported in the dissolved (or colloidal) phase, and were not removed via sediment deposition, biological uptake or chemical degradation. For most pesticides detected in the GBR, the key chemical properties relating to mobility characteristics and potential environmental fate are well known (Table 1). There has, however, been surprisingly limited attention paid to the partitioning during wet season transit and delivery of the key pesticides relevant to the ecological integrity of GBR ecosystems.

1.2. Study overview This study investigates the temporal dynamics of wet season pesticide runoff from one of the GBR’s major sugarcane growing areas, the lower Burdekin region, using data collected from a five year wet season-intensive monitoring program. The study focuses specifically on documenting the variability of the loads of herbicides (i.e. diuron, hexazinone, ametryn and atrazine and degradation products) identified as posing greatest risks to GBR marine ecosystems (Lewis et al., 2009, 2012) from one of the Great Barrier Reef catchment’s major agricultural zones. The potential to detect future reductions in loads due to improved management practices are discussed. Water quality sampling from one wet season (2009– 2010) also investigates the transport partitioning of key pesticides, specifically those transported in dissolved versus particulate bound forms. Results from this study component are related to an associated assessment of pesticide presence in benthic sedi-

ments from freshwater, estuarine and marine environments across the study area. 1.3. Herbicides and the Australian sugar industry A broad range of pesticides have been detected from GBR waters, with herbicide detections predominating, and in many cases are attributable to specific land uses in upstream catchments (Mitchell et al., 2005; Davis et al., 2008; Lewis et al., 2009; Packett et al., 2009; Rohde et al., 2008). Sugarcane cultivation is second only to pastoral cattle grazing as the dominant agricultural land use in the GBR catchment area, with the industry concentrated almost exclusively along the coastal zone. Australian sugar production currently relies heavily on a wide variety of herbicidal applications to maintain productivity (Hamilton and Haydon, 1996; Johnson and Ebert, 2000). Elevated herbicide concentrations have been particularly associated with sugarcane cultivation in the adjacent catchment area (Bainbridge et al., 2009a; Davis et al., 2008; Lewis et al., 2009). A number of the photosystem II inhibiting (PSII) herbicides that are mainstays of Australian sugarcane industry (atrazine, ametryn, diuron, hexazinone) have been identified as a particular concern for GBR ecosystems (Brodie et al., 2012; Lewis et al., 2009). 1.4. The lower Burdekin catchment The Burdekin River watershed accounts for a sizable proportion (33%) of the entire GBR catchment area. Pastoral cattle production represents the dominant land use in the Burdekin catchment (80% of catchment area), with the lower Burdekin delta also containing one of Australia’s largest and most intensively developed agricultural floodplains (Dight, 2009a). The lower Burdekin region includes Australia’s largest single sugar growing environment, accounting for approximately one-third of Australia’s total sugar production. The Burdekin sugarcane district (North Queensland) possesses several management features that set it apart from other GBR sugarcane districts. This includes a wet–dry monsoonal climate, high productivity, high crop inputs, and a virtual total reliance on furrow irrigation and burnt cane harvesting (Thorburn et al., 2007). Pesticides used in the lower Burdekin (Table 1) broadly parallel those used across the Australian sugarcane industry (see Hamilton and Haydon, 1996; Johnson and Ebert, 2000). Definitive data for yearly application rates of pesticides are not readily available for the lower Burdekin district. Sales figures

Table 1 Summary table outlining the commonly detected pesticides in GBR marine and freshwaters, and environmentally relevant chemical properties. Chemical properties compiled using the ‘Footprint’ Pesticide Properties Database http://sitem.herts.ac.uk/aeru/footprint/en/index.htm. Chemical active ingredient (A.I.) Herbicides Ureas/phenylureas Diuron Tebuthiuron Triazines/triazinones Atrazine Ametryn Hexazinone Insecticides Organochlorines DDT Lindane Endosulphan Dieldrin Organophosphates Chlorpyrifos

Mode of action

Solubility in water at 20 °C (mg L 1)

Half-life in soil (DT50: days)

Organic carbon sorption constant (KOC ml g 1)

Photosystem II inhibition Photosystem II inhibition

35.6 2500

75 400

1067 80

Photosystem II inhibition Photosystem II inhibition Photosystem II inhibition

35 200 33,000

29 37 90

100 316 54

Sodium channel modulator Nervous system stimulant Nervous system stimulant Nervous system stimulant

0.06 8.5 0.32 0.14

6200 121 50 1400

151,000 1100 11,500 1200

Acetylcholinesterase inhibitor

1.05

50

8151

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obtained from a single major Burdekin agro-chemical supplier for 2005 can provide some general overview of local application rates of various products during certain years when water quality monitoring has occurred. Paraquat (28,000 kg), glyphosate (22,000 kg), atrazine (20,000 kg including formulations with ametryn), 2,4-D (18,000 kg), diuron (6500 kg), MCPA 500 (3500 kg), asulam (2500 kg) and ametryn (1800 kg) were the most purchased herbicides from this particular supplier in 2005 (see also Davis et al., 2008). The Burdekin River delta and floodplain are drained by the Burdekin River, Haughton River, Barratta Creek and a number of additional natural and artificial drainage channels (Fig. 1). A notable hydrologic characteristic of the region is that due to local topography north of the main Burdekin River channel, the majority of the drainage from canelands between the Burdekin and Haughton Rivers flows into the Barratta Creek system (Davis et al., 2008, 2011). A similar situation occurs to the south of the main Burdekin River channel where most drainage occurs through small coastal creek systems other than the main river channel. In addition to the area’s high economic values, the Burdekin floodplain also supports environments with significant ecological values at both national and international levels. The Burdekin River delta encompasses a number of overlapping wetland complexes listed on either Australia’s National Directory of Important Wetlands or included on RAMSAR’s list of wetlands of international significance (see Fig. 1; ANCA, 1996). The site includes areas of the Great Barrier

Reef World Heritage Area and is contiguous with the Great Barrier Reef Marine Park area (ANCA, 1996).

2. Methods 2.1. Sub-catchment and catchment water quality sampling Water quality sampling for this study focused upon two major sub-catchments of the lower Burdekin floodplain; Barratta Creek and the Haughton River (Fig. 1), for the period 2005–2010. These are the only two gauged sub-catchments of the lower Burdekin floodplain (allowing for contaminant load calculation), and are also relatively accessible during flood events. Land use in the Barratta Creek sub-catchment (1167 km2) is predominantly grazing on native pastures, although approximately 31% of land area is dedicated to irrigated sugar production. The larger Haughton River subcatchment (2324 km2) is similarly dominated by pastoral grazing in the upper parts of the catchment, with approximately 5% of land area dedicated to irrigated sugar and 1% for horticulture, with these intensive agricultural industries confined to the lower coastal plain (Dight, 2009b). One sub-catchment site (Upper Barratta Creek) and three end-of-catchment sites (Haughton River, East and West Barratta Creek) were selected for water quality monitoring. High flow events typically occurred in the wet-season (November–April) when flow was driven by high and intensive

Fig. 1. Map of lower Burdekin land uses, catchment and estuarine–marine sample locations.

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A.M. Davis et al. / Marine Pollution Bulletin 65 (2012) 182–193 Table 2 Overview of lower Burdekin floodplain catchment monitoring sites and sampling frequencies across wet seasons. Site

Upper Barratta Creek (B1) East Barratta Creek (B2) West Barratta Creek (B3) Haughton River (H1)

Water year 2005–2006

2006–2007

2007–2008

2008–2009

2009–2010

Y N Y Y

Y Y Y Y

Y Y Y Y

Y Y Y Y

Y Y Y N

Total event flow sample numbers 59 41 42 36

rainfall. Samples were also collected in low flow conditions in Barratta Creek in a separate study focusing on irrigation tailwater, with the results presented in Davis et al. (2011). Particular effort was paid to the collection of frequent samples during the rising stages of flood events, the period when wetseason contaminant concentrations tend to be highest (Davis et al., 2008). Manual grab samples were collected at all sites through the use of a sampling pole. Samples were typically collected from mid-stream, or if this was not possible, from the streambank. A total of 178 high flow event samples were collected over the monitoring period (Table 2). Additional duplicate samples were also collected at selected sites for analytical precision estimates. All pesticide samples were collected into 1 L amber glass bottles, supplied by the Queensland Health, Forensic and Scientific Services (QHFSS) laboratory. The amber bottles were pre-cleaned with acetone and ethanol and blow-dried with nitrogen fitted with a carbon filter.

spectrometry (LCMS) and gas chromatography mass spectrometry (GCMS) at the National Association of Testing Authorities accredited QHFSS laboratory. Organochlorine, organophosphorus and synthetic pyrethroid pesticides, urea and triazine herbicides and polychlorinated biphenyls were extracted from the sample with dichloromethane. The dichloromethane extract was concentrated prior to instrumentation quantification by LCMS and GCMS (QHFSS method number 16315). Samples collected for analytical precision estimates produced values for pesticides of key interest (ametryn, atrazine, diuron) typically within ±10%. For full methodology see Lewis et al. (2009). Benthic sediment samples were prepared for analysis by extraction with dichloromethane and agitation on a flat bed shaker or use of separatory funnel extraction techniques. The organic extracts were pre-concentrated prior to analysis by use of a Kuderna–Danish evaporator (as described in US EPA methods 1699, 619 and 632).

2.2. Herbicide partitioning between dissolved and particulate forms in floodwaters

2.5. Load calculations

During 2009–2010 wet season event sampling, 2 L samples were collected on nine sampling occasions across the Upper, East and West Barratta catchment monitoring sites and transported to the laboratory. The 2 L sample from each site was then subsampled, with a 1 L sample stored in an amber glass bottle and refrigerated (4 °C) as per standard procedure. The remaining litre of water was filtered through sterile filter modules (Sartorius MiniSart 0.45 lm cellulose acetate) into a 1 L amber glass bottle before refrigeration in preparation for analysis of the dissolved-only fraction. 2.3. Catchment waterway and marine benthic sediment pesticide sampling In July 2009 sediment samples were collected across four sites in the freshwater reaches of the Barratta Creek catchment. The four sites included the three catchment monitoring sites (Upper, East and West Barratta Creek) and an intermediate site on Barratta Creek at Allen Road (Fig. 1). The East Barratta Creek site was resampled in January 2010. Sediment samples were also collected in January 2010 from nine estuarine and nearshore marine sites across Bowling Green Bay, the initial marine receiving environment for Haughton River and Barratta Creek outflows (Fig. 1). All benthic sediment samples, whether sourced from marine or catchment waterway environments, were collected using a stainless steel van-Veen sampler (depth of sampled sediment was 120 mm) and transferred to 1 L, solvent-washed glass containers. All estuarine– marine sites were within the bounds of the Bowling Green Bay RAMSAR wetland site and GBR World Heritage Area.

Loads of herbicides (in kilograms) discharged from the Barratta Creek and Haughton River catchment sampling sites into downstream receiving environments during each wet season were calculated using the Queensland Department of Environment and Resource Management’s (DERM) Brolga loads program (NR&W, 2007). Flow data were obtained from DERM gauging stations on the Haughton River at Powerline (GS119003a) and (upper) Barratta Creek at Northcote (GS119101a), and SunWater gauging stations on East (GS119102a) and West (GS119103a) Barratta Creeks. Continuous time series flow data from the different gauging stations, and water quality concentration data, were entered into the Brolga database, and loads were calculated using the linear interpolation technique. Several reviews of techniques for load calculation suggest that in situations like this study, with gauged catchments, representative event data and continuous sample collection (e.g. daily, near daily), linear interpolation is one of the more suitable load estimation approaches (Fox et al., 2005; Letcher et al., 1999). Herbicide loads at all sites were calculated using hourly cumecs (m3 s 1) as the discharge unit. Due to gaps in the flow data at East and West Barratta Creeks over some of the monitored wet seasons, flow data had to be periodically estimated at these two sites. Relatively consistent relationships between flow at these sites, as well as the Barratta Creek at Northcote gauge, were used to derive estimated flow data for these periods (see Supplement 1). It should also be noted that loads were only calculated for the monitored (sampled) period of streamflow at each site. Due to constraints on contaminant sample numbers or site inaccessibility in some years, the entire period of wet-season flow was not monitored for herbicide concentrations. 3. Results

2.4. Analytical methodology 3.1. Wet season flood event herbicide dynamics All pesticide samples (water and sediment) were couriered to the QHFSS laboratory in Brisbane, Queensland for analysis. The water samples were analysed by liquid chromatography mass

The temporal dynamics of wet season herbicide concentrations followed a similar pattern across all monitored lower Burdekin

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Fig. 2. Upper Barratta Creek discharge (GS119101a: 01/10/2008 to 01/06/2010) and concentrations (lg L 1) of the herbicides atrazine and diuron detected during wet season flood events. Shading indicates sequential discharge hydrograph sections for which separate loads were calculated.

Table 3 Annual peak wet season herbicide concentrations (lg L

1

) in flow events recorded at lower Burdekin catchment monitoring sites.

Site

Water year

Ametryn

Atrazine

Desethyl atrazine

Desisopropyl atrazine

Diuron

Hexazinone

Tebuthiuron

Upper Barratta Creek (B1)

2005/06 2006/07 2007/08 2008/09 2009/10

0.14 0.04 0.1 0.06 0.07

6.5 2.3 5 22 16

1.2 0.33 0.64 1.5 0.95

0.29 0.1 0.15 0.24 0.29

3.8 1 2.2 8.5 6.5

0.04 0.04 0.04 0.03 0.12

n.d. n.d. n.d. 0.02 0.03

West Barratta Creek (B3)

2005/06 2006/07 2007/08 2008/09 2009/10

0.14 0.06 0.05 0.02 0.02

4.4 2.1 4.5 6.9 4.2

0.66 0.5 0.55 0.89 0.46

0.14 0.12 0.16 0.26 0.15

3.4 1.4 1.7 3.2 2.9

0.03 0.09 0.02 0.01 0.08

n.d. n.d. n.d. n.d. 0.02

East Barratta Creek (B2)

2006/07 2007/08 2008/09 2009/10

0.05 0.05 0.02 0.06

2.6 2.1 5.1 13

0.51 0.39 0.86 0.69

0.11 0.11 0.28 0.27

1.1 1.1 3.5 3.5

0.03 0.03 0.06 0.06

n.d. n.d. 0.01 0.03

Haughton River (H1)

2005/06 2006/07 2007/08 2008/09

n.d. n.d. 0.12 0.01

0.17 0.04 0.1 0.12

0.04 n.d. 0.03 0.04

1.5 0.15 0.67 0.79

0.02 n.d. n.d. n.d.

n.d. n.d. 0.01 0.3

n.d. = below analytical detection limits (0.01 lg L

2.2 0.09 0.75 1

1

).

sites. We present the temporal dynamics of herbicide concentrations over the duration of wet season flood events occurring at Upper Barratta Creek during the 2008–2009 and 2009–2010 wet seasons as an example (Fig. 2). Highest detected concentrations invariably occurred during the initial stages of early wet season flow events, and particularly in the cases of atrazine and diuron, were often considerably higher than relevant ANZECC and ARMCANZ (2000) water quality guidelines for ecosystem protection. For example, several early wet season detections of atrazine were greater than 15 lg L 1 (Table 3), frequently exceeding the ANZECC and ARMCANZ (2000) 99% ecosystem protection trigger value of 0.7 lg L 1, and also the 95% trigger value of 13 lg L 1. Similarly, several early wet season concentrations of diuron were greater than 6 lg L 1(Table 3), exceeding the ANZECC and ARMCANZ (2000) ‘low reliability’ trigger value of 0.2 lg L 1. These higher initial concentrations were followed by rapid dilution and/or exhaustion during the subsequent peak and recession of major flood events. Lower concentrations (<0.1 lg L 1) of diuron and atrazine tended to persist throughout the duration of the wet season, with

ametryn, hexazinone and tebuthiuron often undetectable after these initial ‘first flushes’. The peak herbicide concentrations measured during event flows (Table 3) were highly consistent over the five year monitoring program. While there were some potential exceptions (i.e. atrazine in Upper and East Barratta Creeks in 2009–2010), these may be expected given the sampling resolution of the study (typically 1–2 samples per day during event flow) and the rapid dilution of herbicides in the waterways over time as the wet season proceeds. 3.2. Wet season herbicide loads Atrazine was the herbicide that consistently had the highest discharged loads throughout the study period, in both the Barratta Creek system and the Haughton River (Table 4). Atrazine contributed almost half (49.2%) of the measured herbicide load averaged across all years and well over 50% when the atrazine degradate products are also considered (see below). Diuron had the second highest load and accounted for approximately one-third (33.2%)

Table 4 Calculated yearly total event loads for herbicides (kg yr

Upper Barratta Creek (B1)

2005/06 2006/07 2007/08 2008/09 2009/10

54,500 144,000 76,500 410,000 244,000

Average

Haughton River (H1)

c

Sampled discharge (ML)

Total discharge (ML)

1

).

Ametryn load kg (EMC)

Atrazine load kg (EMC)

Desethyl atrazine load kg (EMC)

Desisopropyl atrazine load kg (EMC)

Diuron load kg (EMC)

Hexazinone load kg (EMC)

Tebuthiuron load kg (EMC)

98,000 175,000 325,000 410,000 244,000

4.3 (0.08) 7.5 (0.05) 1.9 (0.02) 0.93 (<0.01) 0.42 (<0.01)

57 (1.05) 100 (0.69) 77 (1.01) 114 (0.28) 55 (0.23)

12 22 20 26 14

3.1 (0.06) 6 (0.04) 4.9 (0.06) 4.3 (0.01) 3.6 (0.01)

37 44 53 63 43

(0.68) (0.31) (0.69) (0.15) (0.18)

0.32 (0.01) 1.4 (0.01) 2.5 (0.03) 1.7 (<0.01) 0.49 (<0.01)

0.03 (<0.01) 0 (<0.01) 0.04 (<0.01) 0.12 (<0.01) 0.44 (<0.01)

185,800

250,400

3.01 (0.02)

80.6 (0.43)

18.8 (0.10)

4.38 (0.02)

48 (0.26)

1.28 (<0.01)

0.13 (<0.01)

2005/06b 2006/07b 2007/08 2008/09 2009/10a

91,000 140,000 62,000 300,000 160,000

105,000 175,000 230,000 305,000 245,000

4.4 (0.05) 4.6 (0.03) 1.3 (0.02) 0.09 (<0.01) 0.54 (<0.01)

64 (0.7) 109 (0.78) 70 (1.13) 54 (0.18) 73 (0.46)

13 29 16 19 19

2.8 (0.03) 6.6 (0.05) 4.5 (0.07) 4 (0.01) 6.6 (0.04)

40 87 44 32 49

1.4 (0.02) 2.1 (0.02) 1 (0.02) 1.8 (0.01) 1.1 (0.01)

0 (<0.01) 0 (<0.01) 0.01 (0.0) 0.01 (0.0) 0.15 (0.0)

Average

150,600

212,000

2.186 (0.02)

74.0 (0.49)

19.2 (0.13)

4.9 (0.03)

50.4 (0.34)

1.48 (0.01)

0.03 (<0.01)

2006/07 2007/08 2008/09c 2009/10a

145,000 53,000 300,000 160,000

180,000 400,000 305,000 245,000

7.7 (0.05) 1.4 (0.03) 0.23 (<0.01) 0.45 (<0.01)

115 (0.79) 42 (0.79) 51 (0.17) 93 (0.58)

31 13 19 19

7.6 3.7 3.5 6.9

66 28 29 46

0.98 (0.01) 0.92 (0.02) 1.5 (0.01) 0.79 (<0.01)

0 (<0.01) 0.03 (<0.01) 0.02 (<0.01) 0.32 (<0.01)

Average

164,500

282,500

2.45 (0.02)

75.25 (0.46)

20.5 (0.13)

5.43 (0.03)

42.3 (0.26)

1.05 (<0.01)

0.09 (<0.01)

2005/06 2006/07 2007/08 2008/09

260,000 475,000 270,000 1100,000

260,000 475,000 715,000 1100,000

0 (<0.01) 0 (<0.01) 0.95 (<0.01) 0.04 (<0.01)

72 24 25 11

5.5 (0.02) 11 (0.02) 3.8 (0.01) 1.8 (<0.01)

1.1 (<0.01) 0.02 (<0.01) 0.68 (<0.01) 0.43 (<0.01)

64 (0.25) 39 (0.08) 16 (0.06) 8.5 (0.01)

0.14 (<0.01) 0 (<0.01) 0 (<0.01) 0 (<0.01)

0.29 (<0.01) 0 (<0.01) 1.7 (<0.01) 0.77 (<0.01)

Average

526,250

637,500

0.25 (<0.001)

33.0 (0.06)

5.53 (0.01)

0.56 (<0.01)

31.9 (0.06)

0.04 (<0.01)

0.69 (<0.01)

(0.28) (0.05) (0.09) (0.01)

(0.22) (0.15) (0.21) (0.06) (0.06)

(0.14) (0.21) (0.26) (0.06) (0.12)

(0.21) (0.25) (0.06) (0.12)

(0.05) (0.07) (0.01) (0.04)

(0.44) (0.62) (0.71) (0.11) (0.31)

(0.46) (0.53) (0.1) (0.29)

A.M. Davis et al. / Marine Pollution Bulletin 65 (2012) 182–193

Water year

East Barratta Creek (B2)

b

) for the sampled discharges on the lower Burdekin floodplain. Values in parentheses are event mean concentrations (EMC; lg L

Site

West Barratta Creek (B3)

a

1

Upper Barratta Creek flow data used to derive site discharge. East Barratta Creek flow data used to derive site discharge. West Barratta flow data used to derive site discharge.

187

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tored wet season discharge was conducted for the Upper Barratta Creek monitoring site. This site had the most comprehensive pollutant concentration dataset and high quality discharge data (GS119101a), and allowed an assessment of whether the magnitude of wet season flood flows was a reliable predictor of herbicide load. Linear regression of annual herbicide loads versus wet season discharge at Upper Barratta Creek highlighted no significant relationship between wet season herbicide loads and wet season discharge (R2 = 0.328, p > 0.05). This suggests the total magnitude of flood discharge over the course of a wet season is not a reliable predictor of herbicide losses in the lower Burdekin region.

of the average total herbicide load. Load calculation for separate wet season discharge components (i.e. initial first flushes, first major discharge events and subsequent discharge) at upper Barratta Creek highlighted the majority of pesticide loads in the 2008– 2009 and 2009–2010 wet seasons were lost in early flow events (Fig. 2). For example, during the 2008–2009 wet season 89 kg of atrazine (78% of measured annual load) and 48 kg of diuron (76% of annual load) were lost in the first 49,200 ML of runoff, which accounted for only 12% of total wet season discharge. During the 2009–2010 wet season 48 kg of atrazine (87% of total wet season load) and 39 kg of diuron (91% of total wet season load) were lost in the first 99,000 ML of wet season flow, which accounted for 41% of total wet season discharge. Ametryn, hexazinone and tebuthiuron all made relatively minor contributions to the measured herbicide loads across all sites, with none of these three herbicides contributing more than 5% of total load at any site in any year (Table 4). Herbicide degradates, specifically desisopropylatrazine and desethylatrazine, made larger proportionate contributions to annual herbicide loads than the combined annual loads of ametryn, hexazinone and tebuthiuron in all years. The combination of desisopropylatrazine and desethylatrazine accounted for approximately 15% of the total wet season loads averaged across all sites and years (Table 4). It should be noted that in addition to being a metabolite of atrazine, desisopropylatrazine is also a degradation product of other herbicides such as simazine and cyanazine (Graymore et al., 2001), and as such its presence may not be totally attributed to atrazine use alone. While simazine is occasionally recorded in lower Burdekin waterways at concentrations just above analytical detection limits (0.01 lg L 1; Davis et al., 2008, 2011), the dominance of atrazine use in the local sugar industry and detections in these waterways suggest that most desisopropylatrazine could be attributed to atrazine breakdown. Other minor herbicides measured occasionally in the lower Burdekin and contributing <10 kg include 2,4-D, metolachlor, MCPA and propazine. The IUPAC chemical nomenclature for all pesticides discussed in this study are available in Supplement 2. A simple linear regression of annual loads of diuron and atrazine (the two most commonly detected herbicides) versus moni-

Table 5 Comparison of filtered versus unfiltered herbicide concentrations (lg L Date Upper Barratta Creek (B1) 26/01/2010 28/01/2010 28/01/2010 29/01/2010

East Barratta Creek (B2) 26/01/2010 28/01/2010 29/01/2010

West Barratta Creek (B3) 26/01/2010 28/01/2010

3.3. Wet season herbicide partitioning between dissolved and particulate-bound forms Comparison of filtered versus unfiltered concentrations for samples collected during the 2009–2010 wet season highlighted that the herbicides of greatest concern to GBR ecosystems (diuron, atrazine, ametryn, hexazinone, tebuthiuron) were all transported predominantly in dissolved rather than particulate-bound forms (Table 5). Atrazine and metabolite desisopropyl atrazine demonstrated the highest transport in dissolved form, with on average ca. <10% bound to particles. Diuron had the highest proportion associated with the particulate form, with an average of one-third being particulate-bound. 3.4. Lower Burdekin benthic sediment pesticide concentrations A range of herbicides, specifically diuron, atrazine, ametryn, hexazinone, tebuthiuron and metolachlor were all detected in lower Burdekin catchment waterway benthic sediments (Table 6). Diuron was the herbicide detected at greatest concentration in sediments, with a peak detection concentration of 28 lg kg 1 recorded at East Barratta Creek in January 2010. Indeed, the highest sediment herbicide concentrations were measured at the lower Barratta Creek sites (B2 and B4) before progressively decreasing in the estuarine and inshore marine sites. No organochlorine (DDT, aldrin, dieldrin etc.) or organophosphate (chlorpyrifos) residues were detected in any catchment waterway sediments.

1

) in 2009–2010 wet season runoff samples.

Ametryn

Atrazine

Desethyl atrazine

Desisopropyl atrazine

Diuron

Hexazinone

Tebuthiuron

Unfiltered Filtered Unfiltered Filtered Unfiltered Filtered Unfiltered Filtered

0.02 0.01 0.01 n.d. n.d. n.d. n.d. n.d.

6.3 6.2 2.6 1.4 0.98 0.88 0.26 0.23

0.46 0.34 0.57 0.27 0.22 0.23 0.06 0.05

0.2 0.17 0.23 0.11 0.1 0.08 0.03 0.02

2.3 1.8 1.5 0.48 0.55 0.42 0.38 0.28

0.01 0.01 0.02 n.d. n.d. n.d. n.d. n.d.

0.02 0.01 n.d. n.d. n.d. n.d. n.d. n.d.

Unfiltered Filtered Unfiltered Filtered Unfiltered Filtered

0.02 0.02 0.02 0.02 n.d. n.d.

1.6 1.5 1.2 1.2 0.43 0.4

0.23 0.17 0.4 0.37 0.13 0.12

0.08 0.09 0.16 0.18 0.05 0.05

0.38 0.28 0.67 0.42 0.27 0.16

n.d. n.d. 0.02 0.02 n.d. n.d.

0.02 0.02 n.d. n.d. 0.04 0.04

Unfiltered Filtered Unfiltered Filtered

n.d. n.d. 0.01 0.02

0.73 0.74 1.2 1.1

0.23 0.2 0.31 0.3

0.08 0.1 0.12 0.14

0.57 0.43 0.75 0.51

0.01 0.01 0.03 0.03

0.02 0.02 n.d. n.d.

20

10.5

16.5

6

33.3

20

12.5

Average proportion (%) pesticide particulate bound n.d. = below analytical detection limits (0.01 lg L

1

).

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A.M. Davis et al. / Marine Pollution Bulletin 65 (2012) 182–193 Table 6 Concentrations of pesticides in lower Burdekin riverine sediments and Cape Bowling Green marine–estuarine sediments through the period 2009–2010. Site

Site coordinates

Diuron (lg kg 1)

Atrazine (lg kg 1)

Ametryn (lg kg 1)

Hexazinone (lg kg 1)

Tebuthiuron (lg kg 1)

Metolachlor (lg kg 1)

Catchment sites B1 26/06/2009 B4 26/06/2009 B2 26/06/2009 B3 26/06/2009 B2 15/01/2010

19.707S, 19.619S, 19.568S, 19.570S, 19.568S,

147.147E 147.203E 147.207E 147.222E 147.207E

n.d. 9.4 1.3 n.d. 28

n.d. 3.1 n.d. n.d. 2.4

n.d. 4 n.d. n.d. 1.56

n.d. n.d. n.d. n.d. n.d.

n.d. n.d. n.d. n.d. 0.44

n.d. n.d. n.d. n.d. 0.34

Estuarine–marine sites H2 H4 H12 B10 B12 HB4 CBG2 CBG3 B5

19.386S, 19.368S, 19.284S, 19.340S, 19.324S, 19.392S, 19.382S, 19.339S, 19.448S,

147.141E 147.147E 147.153E 147.258E 147.259E 147.198E 147.380E 147.379E 147.237E

n.d. 0.31 n.d. n.d. 0.15 0.16 0.17 0.12 1.62

n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.

n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.

n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.

n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.

n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.

n.d. = below analytical detection limits (0.01 lg kg

1

).

Diuron was the only herbicide residue detected in Bowling Green Bay estuarine or inshore marine benthic sediments, being found at two-thirds of sample sites (6 out of 9; Table 6). Detectable sediment diuron concentrations ranged between 0.12 lg kg 1 (site CBG3) to 1.62 lg kg 1 at site B5 located in the Barratta Creek estuary (Fig. 1). No organochlorine (DDT, aldrin, dieldrin etc.) or organophosphate (chlorpyrifos) pesticide residues were detected in any marine or estuarine sediments.

4. Discussion 4.1. Dynamics of wet season herbicide transport This study documented the consistent wet season delivery of herbicide loads to downstream freshwater and marine environments from a catchment with significant sugarcane land use. Combined annual loads of measured herbicides exported from these lower Burdekin catchments were consistently in the order of 200+ kg. These loads are comparable to those documented in other GBR catchments containing significant areas of sugarcane production (Lewis et al., 2009). The total loads of herbicides leaving the entire lower Burdekin floodplain on an annual basis are more difficult to accurately quantify. The Barratta Creek complex and Haughton River represent just one component of diffuse drainage characterising surface water runoff from the lower Burdekin floodplain. Several other ungauged catchments which drain a sizeable proportion of the Burdekin’s agriculturally developed floodplain (i.e. Plantation Creek, Sheep Station Creek, Iyah Creek) all demonstrate similar wet season pesticide dynamics to Barratta Creek and the Haughton River (Davis et al., 2008, 2011). The discharge of herbicides to marine environments from intensive agriculture is not a phenomenon unique to the Great Barrier Reef catchment area, and has been repeatedly documented on a global scale (Clark et al., 1999; Power et al., 1999; Southwick et al., 2002). Discussing the variability in herbicide loads evident in this study compared to other areas of the GBR is difficult given the similar temporally limited datasets available in most other GBR catchments (see Lewis et al., 2009). The inter-annual variability of herbicide loads evident in the lower Burdekin is considerably less than that observed in other larger GBR catchments such as the Fitzroy River (Packett et al., 2009), where annual end-of-catchment loads can vary by several orders of magnitude. This is not unexpected given the much smaller catchment area of the lower Burd-

ekin floodplain, where significant rainfall events are likely to occur over the entire agriculturally developed catchment area each wet season, and the much shorter transit times. The lack of a strong correlation between magnitude of wet season flood events and herbicide load losses documented in this study needs to be treated with some caution given the limited temporal extent of the available dataset. However, the similarity in herbicide loads across water years and peak concentrations (Tables 3 and 4) regardless of wet-season magnitude may reflect the nature and amount of herbicide application, and the pronounced ‘first-flush’ characteristics of herbicide loss (Fig. 2). Relatively consistent amounts of herbicides are likely to be applied across the district on an annual basis. Recent data on paddock-scale losses of herbicides from lower Burdekin sugarcane farms suggest proportionate losses of herbicide applied to paddocks typically vary between 1% and 6% (Davis et al., 2011). With most product being lost in the initial stages following the commencement of the wet season, regardless of the extent of subsequent rainfall events, annual loads may be expected to be relatively stable through time. It should be noted that the timing of wet season rainfall over the monitoring period of this study was similar on a yearly basis, with the first major wet season flood events typically occurring in January to February. An earlier wet season onset (October to November) could therefore introduce more variability into yearly loads, but would require considerably longer data records to quantify. Similarly, other events outside the control of sugarcane farmers such as wet weather or mill breakdowns during the harvest period can also extend the post-harvest herbicide application period into higher risk wet season climatic conditions, and can potentially add more variability into yearly loads. The consistent and pronounced ‘first-flush’ behaviours, where the highest herbicides concentrations, as well the majority of total herbicide load, were transported from catchments in the early wet season were also a notable feature of temporal herbicide dynamics. This predictable rapid decay in herbicide concentrations evident at a catchment scale through the course of a wet season is probably due to a combination of several factors including: application timing; relatively rapid herbicide exhaustion; increasing levels of dilution; and catchment dissolution as the wet season proceeds (see Cook et al., in press; Smith et al., in press). Herbicides such as diuron and atrazine are largely applied in the lower Burdekin through the period April to December (Davis et al., 2008). Product applied to topsoil in close proximity to the wet season (i.e. October to November) would be most prone to off-site movement, and

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represent a key point for management intervention to improve water quality (see Davis et al., 2011). These results also suggest the early stages of the wet season likely pose considerable risk to marine environments in terms of exposure to the highest herbicide concentrations. The magnitude, and hence spatial extent of marine flood plumes, of these early wet season events (October to November) may be limited in relation to the typically later major flood events occurring in January to April. The much higher concentrations as well as timing of these earlier flood events, which coincide with key reproductive periods for many keystone marine biota may thus pose greater threats to inshore marine ecosystems.

4.2. Detection of end-of-catchment trends in herbicide water quality Long-term end-of-catchment monitoring to assess the water quality benefits of improved agricultural practices on pesticide discharge to marine environments are a fundamental component of GBR Reef Plan (Carroll et al., 2012; Reef Water Quality Protection Plan Secretariat, 2009). Detecting genuine improvements in water quality in the face of the inherent hydrologic variability of GBR catchments is recognised as a challenge for the Reef Plan monitoring and reporting process (Bainbridge et al., 2009b; Brodie et al., 2012; Darnell et al., 2012). Other GBR water quality contaminants, such as suspended sediment or nutrients, have a natural loading and long-term lag times to management responses (Brodie et al., 2008, 2012). In contrast, the anthropogenic nature of pesticides coupled with their relatively short half-lives (<1 year) may be expected to make detection of management practice change on water quality signatures more likely over much shorter time frames (e.g. Bainbridge et al., 2009b). With this desired outcome of the GBR Reef Plan process it is informative to compare loads lost from lower Burdekin catchments with those loads of herbicide active ingredient (a.i.) lost at a finer paddock scale. Approximately 31% (362 km2) of the Barratta Creek catchment (1167 km2) is used for irrigated sugar production (Dight, 2009b). The average annual load of the dominant herbicide atrazine from East and West Barratta Creek combined (149 kg y 1; Table 4) equates to 4.0 g of a.i. ha 1 y 1 being lost from the catchment over the duration of a wet season. This represents an admittedly coarse ‘upscaling’ that takes no specific account of yearly application rates of these products across the catchment. This catchment-scale herbicide loss per hectare does, however, equate favourably with actual wet season paddock scale losses documented recently from sugarcane farms in the Barratta Creek catchment, particularly when herbicides were applied in close proximity to significant wet season rainfall (Davis et al., 2011). Davis et al. (2011) documented rainfall driven early wet season paddock-scale losses of atrazine ranging between 0.3 and 2.0 g a.i. ha 1 y 1 on monitored farms. These measured paddock scales losses amounted to <1% of herbicide applied, although most paddock-scale herbicide loss on furrow irrigated Burdekin canefarms tends to occur in association with dry season irrigation events (Davis et al., 2011). Given the consistency of this relationship as well as that evident in end-of-catchment peak herbicide concentrations and loads over the five year monitoring period, it would be expected that management change to reduce herbicide runoff should be detectable in the lower Burdekin region should future monitoring continue. Statistical frameworks that incorporate load uncertainty would, however, need to be employed to reliability detect and quantify water quality change (see Wang et al., 2011). The results of this study also clearly highlight the critical need for catchment monitoring programs to allocate a substantial proportion of annual sample numbers (e.g. 30–50%) to the first few early catchment runoff events of the wet season to accurately quantify loads.

4.3. The importance of pesticide degradates in risk assessments The detection of significant loads of several herbicide degradates documented during this study raises some questions regarding the appropriate treatment of chemical degradation products in current GBR risk assessments. Pesticide degradates, a pervasive presence in water quality monitoring program results at a global scale (Clark et al., 1999; Groschen et al., 2000; Kalkhoff et al., 2003; Stork et al., 2008), are an emergent issue for risk assessments due to the often higher toxicity or environmental persistence of these degradates (Graymore et al., 2001; Giacomazzi and Cochet, 2004; Tixier et al., 2000). For example, the two breakdown products of atrazine assessed during this study, desethylatrazine and desisopropylatrazine, are both phytotoxic (Graymore et al., 2001). Desethylatrazine, which contributed the third highest load of the herbicides assessed during this study, and has been detected in lower Burdekin flood plume waters (Davis et al., 2008), is considered almost as toxic as it’s parent compound (Graymore et al., 2001). Similarly, several of the biodegradation degradates of diuron (particularly 3,4-dichloroaniline), one of the priority herbicides from a GBR perspective, have been found to be both several times more toxic to an array of standard toxicological biota as well as more persistent than the parent compound (Giacomazzi and Cochet, 2004; Tixier et al., 2000). Stork et al. (2008) also found significant levels of diuron degradation products in sediments of sugarcane growing areas of the Burnett River catchment of south-east Queensland. These diuron degradates were not assessed in this study, but are almost certainly also present in the lower Burdekin environment, and are yet to be incorporated into any risk assessment process. In fact, most preceding GBR catchment monitoring programs, as well as recent GBR water quality guidelines (GBRMPA, 2009), have dealt only with parent compounds, with minimal appraisal of the occurrence and impacts of intermediate pesticide degradates. 4.4. Herbicide transport and environmental partitioning Results of sediment analyses also have relevance to preceding research conducted in the region. The study of Haynes et al. (2000a) detected no pesticide residues in Bowling Green Bay, although the study sites were further offshore and analytical detection limits were considerably higher than those obtained in this study (0.5 lg kg 1 versus 0.1 lg kg 1). Interestingly none of the organochlorine or organophosphate pesticides that have been previously detected in marine sediments (Haynes et al., 2000a), and particularly irrigation drains in the lower Burdekin (Müller et al., 2000) were detected at any location during this study including the creek sites. This result suggests that these relatively longlived compounds have either degraded to below detectable levels, or have not been transported to lower Burdekin waterway and inshore environments to any great extent. The sediment diuron concentrations documented in this particular study align closely with similar studies from other north Queensland sugarcane growing catchments (Duke et al., 2005; Stork et al., 2008; Wake, in this issue). The discrepancy between the relative end-of-catchment loads of particular herbicides and their presence (or lack thereof) in Bowling Green Bay sediments was an interesting outcome. Atrazine was consistently the dominant herbicide discharged to marine environments during this study, diuron, however, was the only herbicide detected in Bowling Green Bay estuarine and marine sediments. Diuron was also the dominant herbicide in terms of detection frequency and measured concentrations in waterway sediments during this study, both outcomes aligning with previous research on pesticide presence in GBR sediments (Haynes et al., 2000a; Müller et al., 2000). In fact, Haynes et al. (2000a)

A.M. Davis et al. / Marine Pollution Bulletin 65 (2012) 182–193

documented diuron sediment concentrations up to 10 times higher than those in the overlying waters. While compounds such as atrazine are likely to breakdown more rapidly in alkaline marine conditions (see Smith et al., in press), this selective habitat partitioning may also be due in large part to the chemical characteristics of diuron relative to other herbicides. Diuron’s behaviour aligns with both theoretical sorption-partitioning predictions (see Wauchope et al., 2002) as well as field data (see Silburn et al., 2011) on pesticide movement from paddocks. Diuron has one of the lower solubilities, longest half-lives and highest propensities for particulate sorption (KOC) of all the herbicides commonly detected in GBR waterways and marine-estuarine environments (Table 1). The particle sorption capacity of diuron is comparable to several of the organochlorine insecticides that dominated pesticide sediment monitoring in previous GBR studies (Haynes et al., 2000a). The results of this study addressing transport partitioning suggested approximately one-third of diuron was transported during wetseason flood events in particulate-bound form. While the major transport component of diuron was in dissolved form, the greater relative capacity for particulate binding demonstrated by diuron compared to the other herbicides is likely a major reason for it’s prevalence in sediments across the GBR. The role that benthic systems play in sequestering contaminants has been well documented, as well as the capacity for sediment to act as both a sink and a source of contaminants (Baker, 1980; Warren et al., 2003). The risk posed to aquatic environments by sediment concentrations documented in this study in both freshwater and estuarine–marine environments is difficult to predict. Routes of organism exposure may include uptake across body walls from interstitial pore water and overlying water, respiration, and ingestion of contaminated sediment particles (Jantunen et al., 2008). Research on the toxicity of these compounds to aquatic organisms has focused primarily on aqueous (water column) concentrations (Haynes et al., 2000b; Magnusson et al., 2010), with a relative lack of studies addressing sedimentassociated toxicity. The limited available sediment ecotoxicology data focusing on triazine herbicide effects on mangroves (Bell and Duke, 2005; Wake, in this issue) suggest the concentrations detected in this lower Burdekin study were well below known effects levels. Sediment characteristics (particle size, organic fraction) have been demonstrated to have considerable effect on bioavailability of many pesticides and poses considerable methodological challenges to ecotoxicological testing (Mäenpää et al., 2003; Rice et al., 2004). An interesting development in sediment-based toxicology of potential relevance to GBR ecosystems is in regard to bioaccumulation. Most of the perceived risk of pesticide bioaccumulation in biota has been associated with ‘older’ organochlorine pesticides (Haynes et al., 2005; Mortimer, 2000; Negri et al., 2009). The capacity for bioaccumulation of triazine herbicides in benthic fauna (Jantunen et al., 2008) has received comparatively limited attention (although see Haynes et al., 2000a; Russell et al., 1996). Most of the remaining herbicides of interest to GBR ecosystems (ametryn, atrazine, hexazinone, simazine, tebuthiuron) are either highly soluble, have low soil-binding capacity or short half-lives, and often a combination of all three (see Table 1). This study, not surprisingly, highlights delivery of these particular herbicides in predominantly dissolved form. It should be noted that the process of pesticide partitioning calculation used in this study, where cases of no residues being detected in filtered samples (n.d.) were treated as zero dissolved pesticide, may also have exaggerated the average proportions in sediment bound form. A number of these samples were close to the limits of analytical detection (0.01 lg L 1), and a substantial proportion of herbicide could still have been present in dissolved form, but below detectable levels.

191

5. Conclusions Relatively consistent loadings of herbicides appear to be discharged to marine environments on a year-to-year basis across the Burdekin floodplain. Atrazine and diuron accounted for the majority of the measured herbicide load in each year, although herbicide degradates also emerged as a significant contributor to total loads. Pronounced ‘first-flush’ behaviours, where the majority of herbicide load was lost from the catchment in early wet season flood events, regardless of subsequent wet season magnitude, were a notable feature of temporal load dynamics. The PSII herbicides of greatest concern to GBR receiving environments all appear to be transported predominantly in the dissolved phase. Diuron, however, was particularly notable for its relative propensity for partitioning into particulate-bound forms compared to the other PSII herbicides. Diuron is characterised by a combination of physicochemical properties somewhat anomalous to other PSII herbicides. These features are likely responsible for diuron being by far the dominant herbicide detected in benthic sediment surveys across the GBR and its associated catchment area. Improved knowledge of the role of sediment to sequester or influence degradation of pesticides in GBR catchments and ecosystems would enhance understanding of the bioavailability and likely toxicity of these compounds to aquatic biota. Relatively consistent annual variability in catchment herbicide loads, as well as considerable correspondence between end-of-catchment and paddock scale losses, suggest changes in on-farm management practices may be relatively quickly reflected in GBR Reef Plan monitoring program results. Acknowledgements The Australian and Queensland Governments and NQ Dry Tropics are acknowledged for funding of sub-catchment and catchment monitoring, under programs including The Coastal Catchments Initiative, Water Quality Improvement Plans and Reef Plan and support from the Reef Rescue Initiative. Data for the upper Barratta Creek site for 2009/10 were supplied from the Queensland State Government’s Great Barrier Reef Loads Monitoring program. This research was also supported by the Marine and Tropical Sciences Research Facility, implemented in North Queensland by the Reef and Rainforest Research Centre Ltd. Two anonymous reviewers are thanked for comments which greatly improved an earlier version of the manuscript. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.marpolbul.2011.08.025. References Anon., 2003. Reef Water Quality Protection Plan: For Catchments Adjacent to the Great Barrier Reef World Heritage Area. Queensland Government, Department of Premier and Cabinet, Brisbane. ANZECC and ARMCANZ, 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality. Australian and New Zealand Environmental and Conservation Council, Agriculture and Resource Management Council of Australia and New Zealand, Canberra. ANCA, 1996. A Directory of Important Wetlands in Australia, second ed. Australian Nature Conservation Agency, Canberra. Baker, R.A., 1980. Contaminants and Sediments. Ann Arbor Science, Ann Arbor, Michigan. Bainbridge, Z.T., Brodie, J.E., Faithful, J.W., Sydes, D.A., Lewis, S.E., 2009a. Identifying the land-based sources of suspended sediments, nutrients and pesticides discharged to the Great Barrier Reef from the Tully Basin, Queensland, Australia. Mar. Freshwat. Res. 60, 1081–1090. Bainbridge, Z., Brodie, J., Lewis, S., Waterhouse, J., Wilkinson, S., 2009b. Utilising catchment modelling as a tool for monitoring Reef Rescue outcomes in the Great Barrier Reef catchment area. In: Anderssen, R.S., Braddock, R.D., Newham,

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