There are several factors to consider when determining the opportunities for and constraints on the safe use, treatment and disposal of agricultural drainage water. Information and data desired at the site of drainage water production include: rate of drainage water production per unit area, chemical concentration of constituents of concern, and the rates of mass emission. Drainage water management requires additional information and data on drainage water quality and its suitability for the intended water uses as well as an understanding of environmental and health concerns. Upstream drainage water management affects the needs and water quality requirements of downstream water users. As reliable references are already available on estimating drainage water production volumes (Smedema and Rycroft, 1988; Skaggs and Van Schilfgaarde, 1999), this chapter concentrates on the main factors affecting drainage water quality and water quality needs and concerns for other water users and incorporates previous work done by FAO (1997c).
Table 7. Expected quality characteristics of irrigation return flow as related to applied irrigation waters
|
Quality parameters |
Operational spills |
Irrigation tailwater |
Subsurface drainage |
|
General quality |
0 |
+ |
++ |
|
Salinity |
0 |
0, + |
++ |
|
Nitrogen |
0 |
0, +, ++ |
++, + |
|
Oxygen demanding organics |
0 |
+, 0 |
0, -, - |
|
Sediments |
+, - |
++ |
- |
|
Pesticide residues |
0 |
++ |
0, -, + |
|
Phosphorus |
0, + |
++ |
0, -, + |
0 not expected to be much different than the supply water.
+, - slight increase/pickup or decrease/deposition expected.
++ expected to be significantly higher due to concentrating effects, application of agricultural chemicals, erosional losses, pickup of natural geochemical sources, etc
- expected to be significantly lower due to filtration, fixation, microbiological degradation, etc.
Source: Tanji et al., 1977.
Table 7 provides a summary of changes in water quality expected in irrigation return flow relative to irrigation water applied (Tanji et al., 1977). The expected differences in quality in the return flow are described relative to the supply water because actual concentrations in supply waters vary. The operational spill waters (bypass water) from distribution conveyances are not expected to differ much from the quality of the supply water except for some pickup or deposition of sediments. In contrast, surface runoff or irrigation tailwater tends to pick up considerable amounts of sediments and associated nutrients, phosphorus in particular, as well as water-applied agricultural chemicals such as pesticides and nitrogen fertilizers (especially anhydrous ammonia). Tailwater is typically similar to the applied waters in salinity and oxygen demanding organics, termed biochemical oxygen demand (BOD). Subsurface drainage is enriched in soluble components such as dissolved mineral salts and nitrates, very low in sediments, whereas other quality parameters are similar to the irrigation water. These changes in water quality of irrigation return flow depend on a number of factors including irrigation application methods, soil properties and conditions, application of agricultural chemicals, hydrogeology, drainage system, climate, and farmers' water management.
In many regions of the world, municipalities and industries discharge wastewater into open drains initially intended for the conveyance of only agricultural drainage and storm water. In developing countries especially, municipal and industrial wastewater is often insufficiently treated before disposal into such open drains. The result is a risk that agricultural drainage water quality might be seriously contaminated with microbes, pathogens, toxic organics and trace elements including heavy metals.
A knowledge of the composition of the drainage effluent and the ability to predict changes in the composition as a result of changes in crop, irrigation or drainage water management practices are important in the planning and management of drainage water.
The geology of the region plays an important role in drainage water quality. Through weathering processes, the types of rocks (both primary and sedimentary) in the upper and lower strata define the types and quantities of soluble constituents found in the irrigated area. The oceans have submerged many parts of the continents during a period in their geological history. The uplift of these submerged geological formations and receding seas have left marine evaporites and sedimentary rocks behind, high in sea salts including sodium, chloride, magnesium, sulphate and boron. These geological formations exist in varying thicknesses, depths and extents on the continents. Through hydrological processes, solutes can enter the upper stratum by irrigation or floodwater, upward groundwater flow in seepage zones, with rising groundwater levels, or capillary rise. Once the solutes are in the upper strata, they influence the quality of agricultural drainage water through farmers' irrigation and drainage water management. The following example shows how the geology and hydrology of an area influence the quality of agricultural drainage water. It also illustrates the relationship between geomorphology, waterlogging and salinization.
Figure 12. Cross-section of the San Joaquin Valley
Figure 12 shows a schematic cross-section of the San Joaquin Valley with the San Joaquin River as the principal drainage course for this river basin. The eastern side of the valley was formed from the alluvium of the Sierra Nevada, which consists mainly of granitic rocks. The soils derived from Sierran alluvium tend to be coarse textured and non-saline. The eastern groundwaters are characterized as low-salt calcium-bicarbonate-type water with total dissolved solids (TDS) typically in the 200-500 mg/litre range. In contrast, the soils on the western side were formed from alluvium of the Coast Range made up of uplifted marine sedimentary rocks. The soils on the western side tend to be finer textured and saline. The groundwaters on the western side are characterized as moderately saline sodium-sulphate-type waters with TDS typically in the 1 000-10 000 mg/litre range. The unconfined aquifer in both sides of the valley is gradually being filled up with decades of irrigation deep percolation. The soils in the valley and lowest part of the alluvial fans in the western side are waterlogged and salt affected. A nearly water-impermeable clay layer known as the Corcoran clay, about 200 m deep, serves as the boundary between the unconfined and confined aquifer. The groundwaters in the confined aquifer contain from 500 to 1 000 mg/litre TDS. During the geologic past, plate tectonics caused the horizontal-lying Corcoran clay in the shallow sea to tilt upwards forming the Coast Range.
Figure 13. Freebody diagram of water flows in the San Joaquin Valley
Figure 13 is a freebody diagram of the waterlogged irrigated lands on the western side showing the water flow pathways in the surface and subsurface. The applied irrigation water is about 450 mg/litre calcium-bicarbonate-type imported water from the Sacramento River basin to the north. Much of the surface runoff is captured and reused on site. Much of the collected saline subsurface drainage water (4 000-10 000 mg/litre TDS, sodium-sulphate type) is discharged into the San Joaquin River. Especially high concentrations of trace elements such as boron and selenium originating from the marine sedimentary rocks, found in the subsurface drainage waters, have given rise to environmental and health concerns. Discharges from these areas are now constrained by waste discharge requirements.
A second example comes from the Aral Sea Basin, where in total 137 million tonnes of salt are annually discharged, of which 81 million tonnes (59 percent) originate from the irrigation water and 56 million tonnes (41 percent) from the mobilization of salts from the subsoil (World Bank, 1996). In the mid-stream areas, mobilization of subsoil salts is the most substantial. The annual discharge from the Karshi oblast (Uzbekistan) is 10.8 million tonnes of salts, which corresponds to about 34 tonnes per hectare. About 4.3 million tonnes (about 40 percent) originate from irrigation water and the remainder are mobilized salts from the subsoil through irrigation and drainage (World Bank, 1998). Further, in Australia, large amounts of salts are added to the soil profile by atmospheric deposition of salts from upwind wind erosion of salt pans.
Figure 14 depicts the water movement over the soil surface and through the soil profile. Soils serve not only as a medium for plant growth but also store water and nutrients and serve as the porous transport media. The soil's eroding capacity and chemical weathering leads to the generation of water-borne suspended particles and solutes, ranging from nutrients to all kinds of contaminants (Van Dam et al., 1997). Therefore, to understand the role of soils on drainage water quality, it is necessary to understand water movement over and through the soil and the associated suspended and dissolved substances it carries.
Figure 14. Water flow over and through the soil
Of the water added to the soil, either in the form of rainfall or irrigation, part is lost through runoff and direct evaporation at the soil surface. Runoff water collects in natural and constructed surface drains from where it finds its way to the final disposal site (a river, evaporation pond or outfall drain to the ocean or saline lake). The other part infiltrates into the soil. This water fills up the soil pores and restores the soil moisture content up to field capacity under free drainage. The stored water is now available for plant root extraction to satisfy the water requirement of the crop. Any water in excess of field capacity percolates below the rootzone to greater depth in the vadose zone. The deep percolation water may eventually serve as recharge to the groundwater or saturated zone. In irrigated areas with shallow groundwater tables, the recharge is immediate and causes the water table to rise. Where subsurface drainage is installed in waterlogged soils, the drainage system removes deep percolation and groundwater. Where the soil moisture content in the rootzone drops as a result of evapotranspiration and if there is no recharge from irrigation or rainfall, capillary rise into the rootzone might occur, depending on the water table depth, soil texture and structure, and seepage.
Runoff in irrigated agriculture is mainly related to the intensity of irrigation and rainfall events in comparison to the infiltration capacity of the soil. Where the infiltration rate is smaller than the irrigation or rainfall intensity, water will accumulate on the soil surface and run off under a minimum surface slope. Soil degradation, in terms of compaction and crust formation as well as cultivation on steep slopes, promotes surface runoff. Through the physical forces of the running water, soil particles become suspended in the water and are transported to open drains, ditches, streams, rivers and lakes. Deposition of suspended sediments may occur downstream when current velocities decrease. Suspended soil particles are harmful to aquatic life as they diminish light transmission, but also because chemical contaminants may be associated with suspended sediments (NRCS, 1997). Degradation of drainage water quality as affected by runoff from agricultural land is especially important in hilly areas and in areas with excess rainfall.
In the more arid areas and in flat plains, water flowing through the soil profile and associated solute fluxes affected by mineral solubility and adsorption processes are of more importance for the final quality of the drainage effluent than surface runoff. In some places, weathering of soil particles might play a major role in the quality of drainage water (such as dissolution of gypsum). However, in general the soil's ability to adsorb and release through ion exchange and transform chemical elements through microbially-mediated redox reactions plays a more important role. In this context, soil characteristics such as water holding capacity, hydraulic conductivity, clay and organic matter content, soil minerals and soil microbes are important characteristics. Soil degradation in the form of erosion, compaction and loss of biological activity reduces the water and solute holding capacity of the soils. This increases the mobility of solutes through the soil and increases the risk that pollutants such as salts, nutrients and pesticides will be lost both to groundwater and through interception by subsurface drainage to surface water (NRCS, 1997).
As the major transport of solutes through the soil is by the movement of water, climate plays a major role in determining drainage water quality. In humid tropics and temperate regions, the dominant movement of water through the soil is vertically downwards. Solutes, which are brought onto the soil by farmers or are naturally present in the upper soil layers, are leached into deeper soil layers and groundwater. Conversely, in arid climates where evaporation largely exceeds precipitation the dominant water movement through the soil is vertically upwards except during rainfall or irrigation events. Therefore, the chemical composition of deeper soil layers influences the quality of the shallow groundwater and the composition of the soil moisture in the rootzone. Climate and temperature also play a role in the rate of weathering and chemical processes.
Cropping patterns play an important role in the quality of drainage water in a number of respects. First, crops extract water from the rootzone resulting in an evapoconcentration of salts and other solutes in the soil solution. Where the solubility product of minerals is exceeded through evapoconcentration, minerals precipitate out. This changes the composition of the soil solution and thus influences the chemical quality of subsurface drainage waters. Second, crop residues add organic matter to the soil profile. Organic matter in the soil increases the adsorptive capacity for metals and other solutes. Furthermore, organic matter enhances the soil structure, which increases the water holding capacity of the soil. The organic matter also serves as a carbon source for soil microbes involved in transformations such as denitrification, sulphate reduction and methane production in submerged soils. Third, plants extract nutrients through their rooting system and some plants have the capacity to accumulate large amounts of certain salts and toxic elements. Fourth, as not all crops have the same salt tolerance the type of crop largely determines the maximum salt concentration in the rootzone and the amount of water needed to maintain a favourable salt balance in the rootzone. Last, incorporation of nitrogen fixing crops such as legumes can help to reduce nitrogen leaching. Legumes in symbiosis with nitrogen fixing bacteria are both users and producers of nitrogen. They can substitute chemical nitrogen fertilizer in the crop rotation. Deep-rooted perennial crops such as alfalfa can also help to prevent nitrogen leaching by absorbing large amounts of nitrogen (Blumenthal, et al., 1999).
Application of fertilizers, pesticides, soil and water amendments, and animal manures may influence the quality of drainage water to a great extent. The amounts and timing of application in relation to the growing stage of the crops, timing of irrigation, drainage practices and applied soil conservation measures largely define the influence of fertilizer, amendment and pesticide application on drainage water quality. Furthermore, the characteristics of the fertilizers themselves play a major role, also on the possible contamination of drainage water. Most nitrogen fertilizers are highly soluble and mobile in the soil, and nitrates readily enter drainage water through leaching processes. A portion of the nitrogen fertilizers made up of ammonium or anhydrous ammonia is initially adsorbed to the soil exchange complex but ammonium ions oxidize readily to nitrate. This is also true of urea containing nitrogen fertilizers that eventually oxidize to nitrates. Conversely, phosphorus fertilizers are less mobile in the soil because they have very low solubility, and phosphates are adsorbed on positive sites in soil organic matter and clay minerals (Westcot, 1997). The main route for phosphorous to drainage water is through runoff as sediment-bound inorganic and organic phosphate. Runoff waters may contain residues of anhydrous ammonia injected into irrigation water. Ammonia is highly toxic to fish. Runoff waters may also contain sediment-bound organic nitrogen. Excessive levels of nitrogen and phosphate in discharge waters may result in the eutrophication of water bodies.
Water and soil amendments such as gypsum may contribute significantly to salinity. Although gypsum is sparingly soluble, calcium contributed from gypsum may exchange with adsorbed sodium, and sodium sulphate is a highly soluble mineral. Acidic amendments react with soil calcium carbonates but do not contribute to salinity because the solubility of calcium carbonate is very low. Where animal manures are mineralized, nitrates and salts are produced. Some animal manure (e.g. poultry manure) contains appreciable amounts of salts.
Irrigation and drainage management are the main factors influencing the flow of water over and through the soil in arid zone croplands. As solute transport takes place mainly through soil water fluxes, irrigation and drainage management determine to a great extent the solute fluxes through the soil. In irrigation management, timing in relation to crop water requirements and to fertilizer and pesticide application is key to controlling the amount of soluble elements that will leach below the rootzone, from where they can be intercepted by subsurface drains. The timing of irrigation also affects capillary rise into the rootzone, which might cause an accumulation of salts in the rootzone but at the same time reduces the need for irrigation water. Of equal importance to the timing is the amount of irrigation water applied. Excess water, including infiltrated rainfall, leaches to deeper soil layers.
The choice of drainage technology and certain design choices influence the quality drainage effluent considerably. Subsurface drainage enhances the flow of water through the soil. Studies from the United States of America have shown that, in comparison to soils drained by means of surface drains, subsurface drainage reduces the amount of runoff and subsequently reduces the phosphorous contamination of surface water. At the same time, subsurface drainage enhances nitrogen leaching. Upon interception of leaching water, the nitrogen concentration of the final drainage effluent increases. In the same way, subsurface drainage intercepts other soluble elements present in the soil. This offers the possibility to control the concentration of harmful salts and toxic trace elements in the rootzone for optimal crop production but reduces the possibility for disposal and reuse of drainage water.
For water table and salinity control in irrigated lands, two types of drainage technologies are generally employed: subsurface tile drainage, and tubewell drainage. Subsurface tile drains are generally installed at a depth of less than 2 m below the soil surface. However, exceptionally, drains are installed at depths of up to 3 m. Tubewell drains vary between 6 and 10 m but might reach a depth of 100 m, e.g. deep tubewell drains in Pakistan. These differences in drain depth influence the quality of the effluent. The quality of drainage water from subsurface tile drains is influenced by the quality of irrigation water, the applied farm inputs and the quality of the shallow groundwater. In contrast, the quality of drainage effluent from tubewell drains is related mainly to the quality of the groundwater and to a lesser extent to the quality of irrigation water. Based on computer simulations in Pakistan, a comparison of tubewell and tile drainage showed that for Kairpur the salinity of drainage effluent was 11 750 and 2 100 ppm for tubewell and tile drainage, respectively, whereas for Panjnad Abbiana it was 22 918 and 610 ppm, respectively (Chandio and Chandio, 1995).
Table 8. Salt applied in irrigation water and removed by drains
|
Treatment |
Salt removed |
Salt removed/Salt |
|
Deep drains |
607 |
11 |
|
Shallow drains |
32 |
0.7 |
* Mean salt applied = 50 kg/ha
Source: Christen and Skehan, 2001.
The depth of tile drain installation also influences the amount and quality of the generated drainage water. In the Aral Sea Basin, deep drains increase the flow that has to be evacuated and its salt content through mobilization of fossil groundwater through deep flow paths. For example, drainage water in the Karshi oblast collector is extremely saline (up to 8 g/litre). Expected mineralization levels of drainage water resulting from irrigation only are much less (case study Aral Sea Basin, Part II). Experiments in Australia with shallow and deep drains in a vineyard on a clay soil led to the same conclusion. Less drainage effluent was generated with shallow drains (0.7 m deep) than with deep drains (1.8 m deep). Moreover, the shallow drainage water had a lower salinity than the deep drainage water; together with reduced drainage volumes this resulted in a reduced salt load. As the deep drains removed much more salt than imported with the irrigation water (Table 8), this led to the conclusion that the salinity in the effluent of the deep drains was derived from deeper soil layers (Christen and Skehan, 2001).
Model simulations by Fio and Deverel (1991) confirm these observations. They showed that the base flow towards tile drains increased and the quality decreased with increasing installation depth under non-irrigated conditions. They also showed that this is a result of the path flows becoming deeper and longer with increased drain depth and spacing. On the other hand, during irrigation events the discharge of the shallow and deeper drains became similar as recharge due to deep percolation increased and the proportional contribution of deep groundwater to drain lateral flow decreased.
The major cations and anions making up salinity are sodium, calcium, magnesium, potassium, bicarbonate, sulphate, chloride and nitrate. A lumped salinity parameter is frequently used such as EC in decisiemens per metre (dS/m) or TDS in milligrams per litre. The water quality of surface runoff typically deviates little from the composition of the irrigation water even if it flows over soils with visible salt crusts (Reeve et al., 1955). On the other hand, deep percolation displaces salts accumulated in the soil profile from natural chemical weathering, blown in by salt dust, as well as evapoconcentrated salts derived from the applied irrigation water. Thus, the salt content of the collected subsurface drainage water mainly reflects the salinity characteristics of the soil solution, which in turn is influenced by soil parent material, salinity of the shallow groundwater and salts brought into the soil with irrigation water. In many places, the drainage water composition is further influenced by the mineral composition of deep groundwater which is intercepted by the drains.
Trace elements are commonly present at low levels in nature. Many trace elements are essential micronutrients in very small quantities such as iron, manganese, molybdenum and zinc, but the range between deficiency and toxicity is narrow. Trace elements of concern in drainage water from irrigated lands include: arsenic, boron, cadmium, chromium, copper, lead, mercury, molybdenum, nickel, selenium, strontium, uranium, vanadium and zinc. The heavy metal trace elements are fixed strongly by soil materials and tend to be mobile only in the topmost soil layers. However, some of them form mobile metal-organic complexes in the presence of organic matter. Some trace elements (arsenic, selenium, molybdenum and uranium) are relatively immobile in the reduced form (precipitated or elemental) or are adsorbed while the oxidized and oxyanion species are mobile. For example, selenium is soluble in alkaline and well-oxidized soils.
Similar to the dissolved mineral salts, trace elements are evapoconcentrated in the presence of a growing crop as more or less pure water is lost into the atmosphere and the trace elements remain in the soil solution. Elevated concentrations of selenium, boron and molybdenum may be found in soils formed from Cretaceous shale and marine sedimentary rocks and their shallow groundwaters such as in the western side of the San Joaquin Valley. Selenium typically exists in a reduced form in geologic formations as seleniferous pyrite or organic forms of selenium. When the formations are exposed to the atmosphere, the reduced selenium oxidizes into soluble forms and may be subject to transport from irrigation and precipitation.
Naturally occurring arsenic is commonly found in volcanic glass in volcanic rocks of rhyolitic to intermediate composition (Hinkle and Polette, 1999). It may be adsorbed to and coprecipitated with metal oxides, adsorbed to clay minerals and associated with sulphide minerals and organic carbon (Welch et al., 1988). In natural aquifers, arsenic is especially a problem in West Bengal, India, and Bangladesh but it reportedly also occurs in the United States of America, Hungary, Chile, China, Argentina, Ghana, Mexico, the Philippines, New Zealand and Mongolia.
To evaluate whether trace elements might potentially cause a problem in drainage water management, Westcot (1997) studied ranges and geometric mean values in soils for the priority pollutant trace elements. Comparing actual soil data with these levels would give an initial idea of the potential for trace element leaching.
The two main agropollutants in agricultural drainage waters are nutrients and pesticides. As nutrients were covered above, this sections focuses on pesticides. Once a pesticide enters the soil, its fate is largely dependent on sorption and persistence (NRC, 1993). Sorption is mainly related to the organic carbon content of soils while persistence is evaluated in terms of half-life (the time taken for 50 percent of the chemical to be degraded or transformed). Pesticides with a low sorption coefficient and high water solubility are likely to be leached while pesticides with a long half-life could be persistent.
Pesticides vary widely in their behaviour. Pesticides that dissolve readily in water have a tendency to be leached into the groundwater and to be lost as surface runoff from irrigation and rainfall events. Pesticides with high vapour pressure are easily lost into the atmosphere during application. Pesticides that are strongly sorbed to soil particles are not readily leached but may be bound to sediments discharged from croplands. Pesticides may be chemically degraded through such processes as hydrolysis and photochemical degradation. Pesticides may be biologically degraded or transformed by soil microbes.
Therefore, many pesticides are mainly found in surface drainage water and not in subsurface drainage water as such, due to the filtering action of the soils. However, some pesticides have a tendency to be leached through soil profiles and accumulate in groundwater, such as organophosphates (e.g. DBCP and Atrazine).
Sediment contamination is a main concern for surface drainage in hilly areas and in areas with high rainfall. Sediment production in arid zones occurs in improperly designed and managed surface irrigation systems, especially furrow irrigation. Sediments are a direct threat to living aquatic resources and the aquatic environment in general. They also increase the cost of drinking water treatment and maintenance of open surface drainage networks from sediment deposition like in Pakistan and China. In addition, phosphate, organic nitrogen and pesticides bound to sediment particles are a source of pollution. Sediment production can be reduced by minimum tillage practices (NRC, 1993) and limiting surface runoff through sound irrigation practices. Sediment settling ponds may be used to reduce the load of sediments in receiving waters. Polyacrylamides appear to serve as an excellent coagulant for sediments in farm drainage. Sediments are not normally found in subsurface drainage water. However, a drainage pipe filled with soil particles might cause sediment pollution in subsurface drainage water.
The total concentration of salts in drainage effluent is of major concern for irrigated agriculture. Salinity in the rootzone increases the osmotic pressure in the soil solution. This causes plants to exert more energy to take up soil water to meet their evapotranspiration requirement. At a certain salt concentration, plant roots will not be able to generate enough forces to extract water from the soil profile. Water stress will occur, resulting in yield reduction. The extent to which the plants are able to tolerate salinity in the soil moisture differs between crop species and varieties.
For the stability of the soil structure, the composition of the soil solution is an important factor. In the solid phase, soils have a net negative surface charge. The magnitude of the cation exchange capacity (CEC) depends on the amount and type of clay and the organic matter content. Cations such as calcium, magnesium, sodium, potassium and hydrogen are adsorbed on the exchanger sites. Normally, a large fraction of the adsorbed cations is divalent calcium and magnesium. Divalent cations adsorbed to clay minerals provide structure and stability. Where monovalent cations dominate the exchangeable cations (sodium in particular), the soil structure looses its stability and structural degradation occurs easily. As cations are mutually replaceable, the composition of the exchangeable cations is related to the proportion of cations present in the soil solution (Jurinak and Suarez, 1990). Therefore, where drainage water reuse for irrigation purposes is under consideration, not only the total salt concentration should be taken into account, but also the sodium to calcium and magnesium ratio, commonly expressed as the sodium adsorption ratio (SAR). High bicarbonate waters tend to precipitate out calcium carbonate. This may increase the SAR in the soil solution and increase the exchangeable sodium percentage (ESP) on the CEC.
The composition of the salts is also important for crop growth. Dominance of certain ions might cause an imbalance in ion uptake. This results in deficiencies of certain elements and depressed yields. The presence of high concentrations of sodium inhibits the uptake of calcium, causing nutritional disorders. Other ions can be toxic, causing characteristic injury symptoms as the ions accumulate in the plant. Toxic elements of major concern are chloride, sodium and boron (FAO, 1985b).
The extent to which crops suffer from salinity stress depends on several factors. Although yield reductions are defined as a function of the average salt concentration in the rootzone, interactions between soil, water and climatic conditions influence the relationship. Exceedingly high air temperatures may cause a reduced salt tolerance. Cultural practices also determine to a certain extent yield reduction resulting from salinity stress. Other plant characteristics (which differ between plant species, varieties of the same species and growth stages during which salinity stress occurs) determine their ability to cope with salinity stress.
The variations between crops in salt tolerance are attributable to the fact that certain crops can make the necessary osmotic adjustment to enable them to extract more water from saline soils. This adjustment involves two mechanisms: absorption of salts from the soil solution, and synthesis of organic solutes. Halophytes tend to absorb salts and impound them in the vacuoles, while organic solutes serve the function of osmotic adjustment in the cytoplasm. Normal plants tend to exclude sodium and chloride ions. For this reason, these plants need to rely more than halophytes on the synthesis of organic osmolytes. As a result, they are more salt sensitive than halophytes. Annex 1 presents data on crop tolerance to salinity and major ions.
Sensitivity to salts changes considerably during plant development. Most crops are sensitive to salinity during emergence and early development. Once established, most plants become increasingly tolerant during later stages of growth. There is general agreement that the earlier the plants are stressed, the greater the reduction in vegetative growth (Maas and Grattan, 1999).
Not all trace elements are toxic and small quantities of many are essential for plant growth (e.g. iron, manganese, molybdenum and zinc). However, excessive quantities might accumulate in plant tissues and cause growth reductions. Crop tolerance to trace element concentrations varies widely. When accumulated in plant tissue, certain trace elements are also toxic to animals and humans upon eating, e.g. selenium, arsenic and cadmium. As plants do not absorb most of the trace elements that are present in the soil, the trace elements accumulate in the soils.
In 1985, FAO published general guidelines for evaluating water quality for irrigated crop production (FAO, 1985b). These guidelines are general in nature and are based on numerous assumptions. Where the actual conditions differ substantially from those assumed, it might be necessary to prepare a modified set of guidelines. The case studies presented in Part II of this publication present several examples of guidelines developed for local conditions in the context of drainage water management. Pratt and Suarez (1990) provide a list of recommended maximum concentrations of trace elements for long-term protection of plants and animals.
Aquatic organisms have different requirements with respect to the chemical and physical characteristics of a water body. Dissolved oxygen, adequate nutrient levels, and the absence of toxic concentrations of hazardous elements are essential factors for sustaining aquatic life. Drainage water disposal can disturb the chemical and physical characteristics of the aquatic habitat.
In natural water, the levels of trace elements are normally very low. Elevated concentration levels have a negative impact and harmful effects on aquatic life. Some trace elements such as mercury and selenium are of particular concern because of their bioaccumulative nature, even at very low concentrations (Westcot, 1997). For example, in the United States of America the regulatory maximum contaminant level for selenium for aquatic biota in freshwaters is 2 ppb and for drinking-waters for humans, 50 ppb. The former is lower due to the bioaccumulation of selenium through the aquatic food chain.
Pesticides may also cause toxicity problems in aquatic organisms in surface waters. While pesticide use is currently highest in North America and Europe, it is expected to increase at a faster rate in developing countries in the near future. Many of the synthetic organic compounds are persistent and bioaccumulate. They magnify up the food chain and are often absorbed in body fat, where they can persist for a long time. In the case of fish tissues and fishery products, some of these compounds may also reach consumers. As fish are an important source of protein, it is essential to prevent and avoid accumulation of contaminants in fish or shellfish (Chapman, 1992).
All aquatic organisms including fish or other aquatic resources living in contaminated water bodies are being exposed daily to a multitude of synthetic chemical compounds that disrupt the development of the reproductive, immune, nervous and endocrine systems by mimicking hormones, blocking the action of hormones, or by other unknown interference with the endocrine system (Rutherford, 1997). Fish have different life stages: egg, larvae, fingerling and adult. Various pollutants may have different effects on their life cycle and on their functions and abilities (e.g. capacity to reproduce, nurse, feed and migrate).
The greatest threat to the sustainability of inland fishery resources is degradation of the environment. According to the GEO-I prepared by the UNEP, access to and pollution of freshwater are among the four key priority areas (FAO, 1999a). Various guidelines have been proposed for water important for fisheries or protecting aquatic environmental quality in general (EIFAC, 1964; British Colombia Ministry of Environment, Land and Parks, 1998; CCME, 1999a; and Chapman, 1992). Water quality guidelines established for temperate regions should not be applied without caution to other climate conditions as toxicity, persistence and accumulation rates might differ substantially (Biney et al., 1994).
Water for livestock watering should be of high quality to prevent livestock diseases, salt imbalance, or poisoning by toxic constituents. Many of the water quality variables for livestock are the same as for human drinking-water resources although the total permissible levels of total suspended solids and salinity may be higher (Chapman, 1992). Annex 2 presents water quality guidelines for livestock drinking-water quality. The guidelines consist of two parts. The first part consists of guidelines for the use of saline water for livestock and poultry. Unsafe levels of salinity and ions depend on the amount of water consumed each day, and on the type, weight, age and physical state of the animal (Soltanpour and Raley, 2001). The second part contains maximum recommended limits of both chemical and microbiological variables. These limits are based on animal health, quality of the products and taste.
The quality of water has a major influence on public health. Poor microbiological quality is likely to lead to outbreaks of infectious water-borne diseases and may cause serious epidemics. Chemical water quality is generally of lower importance. The impact of chemicals on human health tends to be of a chronic long-term nature, and there is time available to take remedial action. However, acute effects may be encountered where major pollution events occur or where levels of certain chemicals (e.g. arsenic) are high from natural sources (WHO, 2000).
Increases in salinity, related to drainage water disposal on a shared water resource, may threaten its use for domestic and drinking-water supply. Although the WHO (1993) has not formulated any guidelines based on TDS, high salt concentrations can cause taste problems. Concentrations of less than 1 000 mg/litre (1.56 dS/m) are normally acceptable to consumers. For the majority of the major ions, no health guidelines have been derived. Present guidelines are based on taste and other side-effects of individual ions, such as staining of laundry by iron, or the rotten egg smell of sulphidic water. Most of the toxic trace elements are included in the health criteria for guidelines for the quality of drinking-water as some of them are carcinogenic. For example, arsenic contamination of drinking-water supplies is of major concern in Bangladesh. Expected concentrations in natural waters are generally well below 1 mg/litre. Where concentrations are exceeded, expensive treatment processes are required to make the water acceptable for human consumption. Some well-known chemical pollutants that affect health include nitrate, arsenic, mercury and fluoride. In addition, there is an increasing number of synthetic organic compounds released into the environment whose effect on human health is poorly understood, but appears to be carcinogenic (WHO, 2000). Annex 3 presents WHO water quality guidelines.
Humans also use water resources for bathing and recreation. Such activities in contaminated waters pose a health risk due to: the possibility of ingesting small quantities; contact with the eye, nose and ear; and contact through open wounds. Health risks related to recreation are mainly related to pathogenic contamination. The potential risks from chemical contamination of recreational waters are usually small. Even repeated exposure is unlikely to result in ill effects at the concentrations of contamination found in waters and with the exposure patterns of recreational users. However, the aesthetic quality of recreational water is extremely important for the psychological wellbeing of users (WHO, 1998).
