The Great Barrier Reef is the largest reef system in the world and extends for >2,300 km along the northern Queensland (Australian) continental shelf (Fig. 1). It consists of an archipelagic complex of >2,900 reefs and covers an area of approximately 225,000 km2. A majority of the reefs are situated on the mid- and outer-continental shelf and are located 40 to 150 km from the mainland. A significant number of reefs (approximately 750) also exist at “inshore” or “nearshore” sites within 40 km of the Queensland coast (Furnas and Brodie 1996).

Fig. 1
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The Great Barrier Reef and its catchments

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The Great Barrier Reef was listed on the World Heritage Register in 1981 in recognition of its outstanding universal value (Lukacs and others 1997). Protection of the ecologic systems of the Great Barrier Reef World Heritage Area from water-sourced pollutants is recognised as being one the critical issues for management of the World Heritage Area (Haynes and others 2001). Evidence derived from modelling and sampling of relatively undisturbed catchments of far-northern Cape York indicate that the export of sediments and nutrients from southern disturbed catchments to the marine environment has increased dramatically during the last 150 years (Furnas 2003; Table 1). There is also increasing evidence concerning the contamination of coastal ecosystems with a range of modern pesticide residues (Haynes and others 2000a; Mitchell and others 2005; Shaw and Műller 2005). Degradation of inshore reefs of the Great Barrier Reef has been associated with increased terrestrial runoff of pollutants in the region between Port Douglas and the Whitsundays (Udy and others 1999; van Woesik and others 1999; Fabricius and De’ath 2004; Fabricius and others 2005; Devantier and others 2006), and damage to both inshore and outer-shelf reefs of the central Great Barrier Reef from Crown of Thorns Starfish (Acanthaster planci) outbreaks has been attributed to increased terrestrial nutrient runoff (Brodie and others 2005).

Table 1 Estimates of change in pollutant inputs to the Great Barrier Reef (from Furnas 2003)
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Pollutant loads originating from land-based runoff are predicted to continue to increase, and there is an immediate need to improve current land-management regimes to minimise this runoff (Brodie and others 2001a; Brodie and Mitchell 2005). A key strategy to affect this change is the recently released Reef Water Quality Protection Plan (Anon 2003). This plan builds on existing government policies and industry and community initiatives and aims to assist in halting and reversing the decline in the quality of water entering the Great Barrier Reef. This will help in the overall protection of the ecological health of the Great Barrier Reef as well as the health of its adjacent catchments and waterways. The Reef Water Quality Protection Plan compliments previous policy that advocated water-quality targets for land-based pollutants discharged to the marine environment (Brodie and others 2001b).

This article summarises the impacts of key water-quality parameters on tropical marine ecosystems and provides a framework on which a long-term water-quality and ecosystem monitoring program has been developed. The monitoring program is described in detail in a companion article (Haynes and others 2007).

Conceptual Models and Monitoring Programs

Conceptual models are a key component in the development of an integrated monitoring program to assess trends in the ecologic status of a system. They assist in identifying the main environmental stressors or threats, appropriate indicators of ecosystem health, and relations or linkages between the stressors and the indicators. Additionally, conceptual models are an important tool for communicating how complex ecologic systems work, the ecologic and physico-chemical interactions between the various components, and which components and linkages might be targeted by specific management actions (Jorgensen 1988; Newton and others 1998; Gross 2003; Downs and others 2005).

In the Great Barrier Reef, conceptual models have been used to underpin process models that describe connections in the paddock-to-reef continuum, e.g., models describing catchments (the SedNet/ANNEX model system, McKergow and others 2005), in risk assessments (Greiner and others 2005), in models describing estuaries and coastal waters (Robson and others 2006) and in models describing reef waters (Wolanski and others 2003, 2004; Wolanski and De’ath 2005; Wooldridge and others 2006). Elsewhere, other models of land and coral reef interactions have been developed using so-called fuzzy logic (Meesters and others 1998; Ruitenbeek and others 1999), but these have only addressed coral reef impacts and fail to include the catchment drivers of environmental effects.

Development of the Reef Water Quality Protection Plan monitoring program was assisted by three conceptual models that encapsulate the risks posed to reef ecosystems from catchment (agricultural and urban) development:

  1. 1.

    A broad-scale system conceptual model (Fig. 2) shows the relations between contaminant generation in the Great Barrier Reef catchment, transport of these contaminants through the catchment by way of local rivers and their estuaries, and their effects on water quality and ecosystem health in the Great Barrier Reef lagoon. In particular, the model illustrates the conceptual links between dry-land grazing and intensive cropping activity (pressures and threats) and the production and dispersion of key water-quality contaminants (stressors). Also shown are the ecosystem attributes (indicators) that can be used to monitor ecosystem health at the catchment, riverine, and marine scales. This conceptual model also includes reference to longer-term threats to Great Barrier Reef water quality and ecosystem health, such as global climate change, that will potentially confound direct model relationships. The model also recognise that system-specific attributes, such as monsoonal wind direction, rainfall intensity and flood plume residence times, will act as system filters to modify the effects of the stressors (Cloern 2001; Devlin and others 2003). In particular, concentrations of suspended sediments, particulate and dissolved nutrients, and pesticide residues in rivers flowing into the Great Barrier Reef region reach peak values during flood events, which are typically associated with monsoonal rainfall and cyclonic activity (Devlin and others 2003), and the movement of river-discharged pollutants in the marine environment is modified by wind direction and intensity and wave action.

  2. 2.

    A process-based conceptual model (Fig. 3) illustrates the movement and partitioning of contaminants between the water column, marine sediments, and marine biota. The physical and biogeochemic properties of contaminants determine the ways in which they disperse in the environment (Bewers and others 1992). This model has been used to ensure that the appropriate ecosystem compartments were sampled and that information was collected on parameters that could influence the biogeochemical properties of the contaminants.

  3. 3.

    A risk-based conceptual model identifies the regions of the Great Barrier Reef lagoon that are at high risk from contaminants transported from the catchment. In particular, adjacent land use and distance and direction of a reef or other ecosystem from the mouth of the major rivers both have significant influence on the potential risk posed by land runoff (Devlin and others 2003; Greiner and others 2005). This latter model has been most useful in optimising sample collection sites and is fully discussed in the second article in this series (Haynes and others 2007).

Fig. 2
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Great Barrier Reef pressure, Vector, response water-quality conceptual model

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Fig. 3
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Great Barrier Reef lagoon process based conceptual model (adapted from Bewers and others 1992)

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Catchment Activity (Pressures and Threats)

The coastal region adjoining the Great Barrier Reef World Heritage Area is divided into 40 wet and dry tropical catchments draining directly into the Great Barrier Reef lagoon (Gilbert and Brodie 2001). Most catchments are small (<10,000 km2); however, the Burdekin and Fitzroy River catchments (133,000 and 143,000 km2, respectively) are among the largest in Australia (Fig. 1). Human activity in these catchments is the primary determinant of altered water quality that is ultimately transmitted to the Great Barrier Reef World Heritage Area.

Although the region remains relatively sparsely populated, extensive land clearing has occurred during the last 200 years since European settlement, and approximately 80% of the land area of catchments adjacent to the Great Barrier Reef World Heritage Area now supports some form of agricultural production (Gilbert and Brodie 2001). Grazing of beef cattle is the largest single land use in Great Barrier Reef catchments, and this has resulted in extensive clearance of forests for conversion to pasture (Gilbert and Brodie 2001). Intensive cropping, mainly of sugarcane but with considerable areas of horticulture and cotton and grain crops, is a significant agricultural industry in the catchments between Bundaberg and Port Douglas. Agriculture and urban expansion have resulted in the loss of substantial areas of coastal wetlands in north Queensland, particularly during the last 50 years (Johnston and others 1998; Finlayson and Lukacs 2003). This has significant ramifications for coastal habitat and local water quality, with the loss of nutrient removal and transformation mechanisms, sediment and toxicant retention, flood flow alteration, and disruption of groundwater discharge and recharge (Lukacs 1998).

Catchment-Based Contaminants (Stressors)

The three major classes of water-quality contaminants increased by human activities in Great Barrier Reef catchments are sediments, nutrients, and pesticides. Land clearing and removal of woodlands for the establishment of grazing pasture has greatly increased soil erosion and sediment loss to local watercourses. Sediment loss is exacerbated by chronic overgrazing during drought periods (McIvor and others 1995). Sugarcane cultivation was also a major source of eroded material under conventional cultivation practices. However, recent management improvements in the sugarcane industry, including green cane harvesting and trash blanketing and minimum tillage, have decreased soil losses by approximately 80% (Rayment 2003). Trash and stubble retention in other cropping systems (cotton, bananas) has also decreased erosion rates (Faithful and Finlayson 2005).

Sugarcane cultivation requires substantial use of inorganic fertiliser, particularly nitrogen. It is estimated that only 35% of the fertiliser applied to sugarcane (typically 150–200 kg/ha/y nitrogen fertiliser) is used by the crop in the year of application (Reghenzani and others 1996); the remainder is lost to the environment (to the atmosphere by way of volatilisation and denitrification and to groundwaters and runoff) or stored in the soil, including trash storage (Freeny and others 1994; Bohl and others 2000; Rasiah and Armour 2001). The proportion lost to each compartment depends on climate, weather, soil type, cultivation practices, fertiliser-application practices, and hydrology (McShane and others 1993; Reghenzani and others 1996).

The use of pesticides (herbicides, insecticides, and fungicides) in Great Barrier Reef catchments has increased progressively in areas under crop cultivation (Hamilton and Haydon 1996). There has also been a shift from the use of organochlorine-based compounds such as DDT, dieldrin, and heptachlor as they were progressively restricted and banned for use between 1973 and 1994. Modern use includes triazine, organophosphate, and urea-based pesticides. Organochlorines and modern agricultural pesticides are widely distributed in Queensland catchment soils, irrigation drains, and river sediments and in the nearshore marine environment adjacent to human activity (Haynes and others 2000a 2006; Müller and others 2000; McMahon and others 2005; Mitchell and others 2005; Shaw and Müller 2005).

Contaminant Transport (Vectors and Filters)

Land-based contaminants generated in the catchment from various land uses are transported to waterways by way of surface runoff and subsurface water flows. Most of this transport occurs during high-intensity rainfall events, when the contaminants are released from the landscape in major flow events (Mitchell and others 1997, 2005). The loads of sediment, nutrients, and pesticides discharged from any one Great Barrier Reef river are proportional to rainfall, to the volume of freshwater discharged during the wet season, and to the extent and type of agricultural land in the catchment (Brodie and Mitchell 2005).

Estimates of runoff from Great Barrier Reef catchments show that most of the sediment (approximately 85%) and nutrients (approximately 65% for N and 78% for P) originate from the southern “dry” catchments that are dominated by grazing, such as the Burdekin and Fitzroy Rivers, where the larger volumes of discharge from these rivers are characterised by greater loads of sediments and nutrients per unit volume of discharge. However, wet tropical catchments (e.g., Tully and Johnstone) have the highest sediment and nutrient loss rates per unit of catchment area, reflecting their higher local rainfall and erosion rates (Brodie and Mitchell 2005).

Waters discharged in flooding rivers form plumes that extend out into the nearshore marine environment of the Great Barrier Reef (Brodie and Furnas 1996; Devlin and others 2001). Mapping and coring of sediments in the Great Barrier Reef lagoon shows that most of the eroded sediment transported by these river systems is deposited within 10 km of the coast (Orpin and others 1999; Lambeck and Woolfe 2000; Neil and others 2002; Orpin and others 2004; Pfitzner and others 2004). Under normal conditions (south-easterly trade winds), these flood plumes flow northward from the river mouth for distances of up to 200 km but are usually constrained to within 20 km of the coast. Dissolved materials, including dissolved nutrients, are transported hundreds of kilometers in flood plumes both offshore and more commonly alongshore (Devlin and others 2001; Devlin and Brodie 2005; Rohde and others 2006). Major floodwaters and their associated dissolved nutrients originating from large rivers, such as the Burdekin River, may extend northward for up to 450 km (Wolanski and van Senden 1983; Devlin and Brodie 2005). Nutrient concentrations in these river plumes can be up to 100 times higher than seawater concentrations in nonflood periods (Devlin and others 2001; Devlin and Brodie 2005; Rohde and others 2006). Inshore ecosystems may be exposed to plume waters and their entrained contaminants for periods of days to weeks during the wet season (Devlin and Brodie 2005).

Secondary transport of contaminants, especially particulate matter, occurs in the period after flood plumes dissipate as wind-generated turbulence and currents resuspend inshore sediments and transport them northward along the coast (Orpin and others 1999). Ultimately, most of the fine sediment in inshore waters is trapped in northward-facing bays (e.g., Broad Sound, Bowling Green Bay, Princess Charlotte Bay) (Orpin and others 2004).

Impacts of Contaminants in the Marine Environment (Indicators)

Three of the major ecosystem types that make up the Great Barrier Reef system (mangroves, seagrass, and corals) are affected differently by water-column contaminants. Mangroves are believed to be affected by herbicides, whereas corals and seagrasses are more affected by increases in suspended sediment and nutrient concentrations as well as by herbicide exposure (Bell and Duke 2005; Duke and others 2005; Fabricius 2005; Jones 2005; Negri and others 2005; Waycott and others 2005).

Nutrient Impact Indicators

The waters of the Great Barrier Reef are characterised by high ambient light intensities and water temperatures. As a consequence, available nutrients are rapidly taken up by phytoplankton and converted to organic matter, particularly in interreef regions (Furnas and others 2005). These recycled and transformed nutrients largely determine the (nutrient) water-quality status of these waters and any impacts on benthic organisms. Phytoplankton biomass measured as chlorophyll a concentrations are two to three times higher in inshore waters of the central and southern Great Barrier Reef compared with areas in the northern Great Barrier Reef (Brodie and others 2007). This is believed to reflect their enhanced nutrient status, which is attributed to terrestrial nutrient discharge from rivers in the central and southern Great Barrier Reef associated with agricultural activities.

Fabricius (2005) and Schaffelke and others (2005) have recently summarised the ways in which increased nutrient concentrations may result in a range of impacts on coral and seagrass communities. Macro-and micro-algal blooms are produced by higher nutrient concentrations, and macro-algae may overgrow coral structures, outcompeting polyps for space and shading coral colonies to critical levels (Hunter and Evans 1995; Stimson and others 2001). Excessive nutrients may also depress coralline algal growth and inhibit reef consolidation (Björk and others 1995). Chronic exposure to increased concentrations of dissolved inorganic nutrients, such as nitrate, can interfere with the relationship between corals and their zoooxanthellae and result in decreased calcification rates (Stambler and others 1994; Marubini and Atkinson 1999; Ferrier-Pages and others 2001). This may weaken the coral skeleton and make coral colonies more susceptible to damage from storm action (Stambler and others 1991; Ferrier-Pages and others 2000). Boring organisms may bioerode coral reef structures at higher rates under increased nutrient conditions, further decreasing overall reef consolidation (Kiene and Hutchings 1994).

Additionally, there is evidence suggesting that coral reproduction and recruitment are highly sensitive to increased nutrient concentrations (Wittenberg and Hunte 1992; Hunte and Wittenberg 1992; Fabricius 2005), with increased concentrations of dissolved organic nutrients inhibiting coral reproductive processes, such as egg size, fertilisation rates, and embryo formation (Ward and Harrison 2000; Koop and others 2001; Cox and Ward 2002). Increased concentrations of particulate organic matter can inhibit all aspects of coral recruitment, including egg fertilization, larval development, larval survival and larval settlement and metamorphosis (Gilmour 1999).

Increased nutrient concentrations can also be deleterious to seagrasses by lowering ambient light levels by way of proliferation of local light-absorbing algae (including water-column phytoplankton, benthic macro algae, or algal epiphytes) and decreasing the photosynthetic capability of seagrass (Walker and others 1999). Increased nutrient concentrations can also cause deleterious disruptions to nitrogen and phosphorus metabolism in seagrass (Touchette and Burkholder 2000), although Australian seagrasses are generally regarded as being nitrogen - limited.

Increased concentrations of nutrients sourced from agricultural activities have been linked to an expansion in the range of seagrasses (e.g., Syringodium isoetifolium) around Green Island since the 1970s (Udy and others 1999). Also, data collected along the north Queensland coast suggest that the tissue nutrient status of the seagrass Halophila ovalis has increased during a 20-year period in concert with increasing fertiliser use by the local agricultural industry and that this seagrass species may be a good bioindicator of local nutrient conditions (Mellors and others 2005).

Sediment Impact Indicators

Evidence exists that high, chronic inputs of terrestrial sediment and organic matter lead to a range of coral and seagrass impacts through shading, smothering, burial, disruption of recruitment, or deleterious community shifts (Dodge and others 1974; Longstaff and Dennison 1999). In contrast, mangroves often respond to increased sediment concentrations through expansion of their range, particularly in areas where river flows and flushing are decreased (Schaffelke and others 2005).

Regardless of whether such sediment loads are natural or the result of human activity, excessive sediment loads can impact corals through smothering when particles settle out (Riegl 1995; Philipp and Fabricius 2003), and by decreasing light availability, coral photosynthesis, and growth (from increased water turbidity) (Rogers 1990; Kleypas 1996; Anthony 1999). Sediments with a higher content of organic matter are more damaging than “clean” sediments in a smothering situation (Weber and others 2006). Sediment impacts on corals can include changes to coral population structure and colony size, altered growth forms, and decreased growth and survival (Rogers 1990; Anthony 2000; Anthony and Fabricius 2000). In particular, low concentrations of sediments and dissolved mucopolysaccharides released by bacteria and other microorganisms can coat corals (Fabricius and Wolanski 2000; Fabricius and others 2003). This creates a metabolic energy drain when the coral removes the aggregate that may decrease reproductive capacity and the organism’s capacity to grow (Stafford-Smith 1993; Riegl and Branch 1995; Telesnicki and Goldberg 1995). Early life–stage corals are at most risk from accumulated sediment through prevention of larval settlement (Hodgson 1990; Gilmour 1999) or burial of the juvenile recruit (Babcock and Davies 1991; Babcock and Mundy 1996; Fabricius and others 2003). Sedimentation is also suspected to adversely impact abundance of crustose coraline algae and influence the development of algal turfs. Both of these effects will compromise coral recruitment (Birrell and others 2005; Harrington and others 2005).

Increases in suspended sediment concentrations and water-column turbidity cause a decrease in water-column light penetration that can adversely affect seagrass photosynthesis rates (Longstaff and Dennison 1999). Seagrass losses of over 1,000 km2 were observed in Hervey Bay, southern Queensland in 1992 and again in 1999 during cyclone and storm events (Preen and others 1995; Campbell and McKenzie 2004). Seagrass loss was attributed to river turbidity decreasing the amount of light reaching the plants as well as to physical removal of seagrass by cyclone and storm action (Preen and others 1995; Longstaff and others 1999).

Pesticide Impact Indicators

The herbicides diuron, simazine and, atrazine are commonly found in flood plumes, coastal waters, and sediments in the Great Barrier Reef region (Haynes and others 2000a; McMahon and others 2005; Rohde and others 2005; Shaw and Muller 2005). More than half of the commonly used herbicides such as diuron exert a toxic effect by restricting electron transfer within the photosynthetic chloroplast of the target plant or alga, leading to a decrease in photosynthetic efficiency (Jones 2005). Because most adult corals rely on symbiotic dinoflagellates to provide additional energy requirements for colony functioning, this may result in a loss of fitness in the host coral polyp (Jones and others 2003). Under extreme conditions, high concentrations of the herbicide, or long periods of exposure to it will result in expulsion of the symbiont (bleaching) from the adult coral (Jones and Kerswell 2003; Jones and others 2003).

Laboratory trials have indicated that these types of herbicides are unlikely to affect fertilisation or metamorphosis in corals at the concentrations likely to be present in inshore Great Barrier Reef waters (Negri and others 2005). However, they may depress photosynthetic activity in juvenile and adult coral symbionts (Negri and others 2005). Laboratory experiments have also shown that low concentrations (3 µg l−1) of diuron will inhibit photosynthesis in crustose coralline algae (Harrington and others 2005). Photosynthetic inhibition is increased when coralline algae are exposed to both diuron and fine sediment (Harrington and others 2005). This may compromise coral recruitment because crustose coralline algae are a critical settlement inducer for many coral species (Heyward and Negri 1999).

A limited number of laboratory trials have investigated the toxicity of diuron and other herbicides to tropical seagrass species (Haynes and others 2000b; Ralph 2000). Effective quantum yield as a measure of photosynthetic activity in Halophila ovalis and Zostera capricorni was found to be significantly depressed by diuron concentrations between 0.1 and 100 µg L−1 after 5 days of laboratory herbicide exposure, whereas effective quantum yield in Cymodocia serrulata was only significantly decreased in plants exposed to higher diuron concentrations (10 and 100 µg L−1). H. ovalis was similarly affected by the herbicides atrazine and simazine. These results indicate that exposure to herbicide concentrations present in inshore Queensland sediments present a potential risk to seagrass functioning, particularly during flood conditions (Haynes and others 2000a; McMahon and others 2005). Laboratory trials have also indicated the relative sensitivity of some mangrove species to diuron exposure compared with other commonly used herbicides (Bell and Duke 2005; Schaffelke and others 2005).

Confounding Stressors and Impacts

The quantity and subsequent impacts of land sourced sediment, nutrient, and pesticide discharges to tropical marine ecosystems may also be fundamentally influenced and/or confounded by a range of indirect water-quality threats. These indirect threats include global climate change, river floods, cyclonic weather systems, and increased capacity for reef algal growth caused by loss of algal grazers, particularly grazing fish (Hughes 1994). These confounding impacts may all be interlinked (Pandolfi and others 2003; Bellwood and others 2004). The expected primary consequence of global climate change is increased seawater temperatures and changes in ocean chemistry (Hughes and others 2003). Increased seawater temperatures increase the frequency and intensity of coral bleaching as well as the incidence of coral disease (Hoegh-Guldberg 1999; Hughes and others 2003). The recognised biological effects of bleaching are decreased coral growth and calcification, decreased reproductive output, and increased mortality (Goreau and MacFarlane 1990). Increased carbon dioxide concentrations in seawater enhance the dissolution of calcium carbonate. This may decrease calcification rates in coral species and lead to changes in coral community structure, reproduction, and overall functioning in coral reef environments (Kleypas and others 1999). The frequency and intensity of tropical cyclones and the damage they induce is also expected to increase as a consequence of global warming with the long-term elevation of seawater temperatures (Hughes and others 2003). The resultant increase in frequency and intensity of river flooding and freshwater inundation of inshore marine waters will result in increased coral mortality, particularly when this coincides with periods of increased seawater temperatures and coral bleaching (Berkelmans and Oliver 1999). The impacts of these stressors must be quantified in a monitoring program to accurately partition the influence of land-sourced pollutants on the reef systems being monitored.

The Reef Water Quality Protection Plan Monitoring Program

The implementation of a long-term water-quality and ecosystem monitoring program based on the conceptual models previously described is central to assessment of the success of land management strategies initiated under the Reef Water Quality Protection Plan (Anon 2003) to improve reef water quality. This Great Barrier Reef marine-monitoring program is comprised of four complimentary subprograms. River-mouth water quality monitoring is carried out to assesses long-term change in the concentrations and loads of the major land-sourced pollutants (including sediments, nutrients, and pesticides) discharged to the marine environment that have the potential to adversely affect coral reef, mangrove, and seagrass ecosystems. Water-quality monitoring is also carried out in the inshore waters (i.e., within 10 to 15 kms of the coast) of the Great Barrier Reef to assess changes with time in concentrations of similar water-quality indicators. Monitoring of the major marine ecosystem types most at risk from land-based pollutants (e.g., intertidal seagrass beds and inshore coral reefs) is carried out to ensure that any change in their status is identified. Coral and seagrass monitoring sites are associated with the inshore marine water-quality monitoring program to enable correlation with concurrently collected water-quality information. In the fourth subprogram, mud crab (Scylla serrata) pollutant concentrations are assessed to monitor the accumulation of specific pesticide and polyaromatic hydrocarbon concentrations in inshore marine biota. Information on weather and sea conditions—including rainfall, seawater temperature and salinity, and river discharge volumes—is also collected to help with interpretation of biological data. In addition, two further subprograms are carried out in the catchments of the Great Barrier Reef. Change in the export of pollutants correlated with land-management change at the property scale are measured through intensive, small-scale monitoring programs, whereas at the subcatchment scale, monitoring of pollutants in event flows allows identification of primary source areas of pollutants and their correlation with different land-management practices. The Reef Water Quality Protection Plan monitoring program is described in detail in a companion article (Haynes and others 2007).

Conclusion

The conceptual models presented here are the first attempt to produce a comprehensive model of catchment to reef pollution and ecological response. This has enabled the construction of numeric models based on these articulated relationships. Quantification of the concepts is now being attempted by way of a number of approaches, including the use of Bayesian Belief Networks (e.g., Thomas and others 2005), fine-scale process models (Robson and others 2006), and empirical process approaches (e.g., Wooldridge and other 2006).