|Main authors:||Susanne Klages, Nicolas Surdyk, Christophoros Christophoridis, Birgitte Hansen, Claudia Heidecke, Abel Henriot, Hyojin Kim, Sonja Schimmelpfennig|
|FAIRWAYiS Editor:||Jane Brandt|
|Source document:||Klages, S. et al. 2018. Review report of Agri-Drinking Water quality Indicators and IT/sensor techniques, on farm level, study site and drinking water source. FAIRWAY Project Deliverable 3.1, 180 pp
Agri-drinking water quality indicators (ADWIs) are selected within the cycles of nitrate and pesticides in the agri-hydrogeochemical system. The drivning force (D), pressure (P), state (S), and link (L) indicators are defined in the agri-hydrogeochemical system while impact (I) and response (R) are considered to be outside the system (see Figure 3.1).
The agricultural system, in the ADWI context, is physically defined by the zone between the atmosphere and the rooting depth where all the agricultural activities and reactions occur (Figure 3.1). In the agricultural system, interplay between human activities (e. g. fertiliser and pesticide use, crop production) and natural processes (e. g. nitrogen cycle, soil erosion, adsorption/desorption, denitrification) control leaching of nitrate and pesticides to the underlying hydrogeochemical system. The driving force (D) and pressure (P) indicators represent the agricultural system and the primary focus is to quantify the leaching and run off of nitrates and pesticides into the underlying hydrogeochemical system.
The hydrogeochemical system is the zone from the ground surface to the drinking water abstraction point (Figure 3.1). The hydrogeochemical system governs pathways to deliver nitrate and pesticide from the agricultural system to the drinking water abstraction point. The pathways control the transit time of pollutants, i. e. nitrate and pesticides, and biogeochemical reactions that may change the concentrations and phase of them in the hydrogeochemical system (Figure 3.1). The state (S), and link (L) indicators describe the fate, retention and transport of nitrate and pesticides in the hydrogeological system. The waterworks system describes drinking water production processes (Figure 3.1).
In the waterworks system, nitrate and pesticides in raw waters – groundwater or/and surface water – might be removed by various types and degrees of processes, depending on the water quality and technological possibilitites. The state (S) indicator shows the quality of drinking water.
Pathways in the hydrogeochemical system are the routes of nitrate and pesticides from the agriculture system to drinking water abstraction points. Identifying the dominant pathways are important for different reasons related to:
- Planning and selection of specific agricultural mitigation measures in regard to protection of water ressources as lakes, streams or groundwater taking the lack time into account,
- Planning and selection of drinking water protection strategies and treatment possibilities in order to secure clean drinking water in the short and long term perspective.
Therefore, one of the key roles of the ADWIs should be identifying the dominant pathways of the hydrogeochemical system. Dominant pathways of the hydrogeochemical system are controlled by complex interplay between its hydrogeologic structure (e. g. soil type, soil thickness, soil moisture, surface topography, bedrock type, groundwater table depth, hydrogeology and hydraulic parameters) and climatic conditions (e. g. seasonality, rainfall intensity); therefore, it may spatially vary and seasonally shift.
In the context of development of ADWI, we conceptualised the pathways as follows: two pathways for groundwater and four pathways for surface waters (Table 3.1).
Two pathways for groundwater: To recharge groundwater, water primarily flows vertically via
- Matrix flow pathways and/or
- Preferential flow pathways (Figure 3.1).
This water eventually emerges back to the surface water.
Matrix flow is a pathway through pore spaces in the soil matrix. In the unsaturated zone, matrix flow moves uniformly with a wetting front, therefore it is also called as uniform flow. The transit time of matrix flow can be long (months to years; Table 3.1); therefore, the groundwater table and groundwater chemistry show relatively small variations and slow changes over time.
Preferential flow is a pathway via macro-pores in the soil and fractures in bedrock, bypassing a dense or less permeable matrix (Beven and Germann, 1982, Hendrickx and Flury, 2001). The macro porous spaces in soils can be created along root channels, soil fauna channels, cracks (i. e. freeze-thaw, wetting-drying), fissure, or soil pipes (Beven and Germann, 1982). Preferential flow may be transiently active; however, it can deliver a significant quantity of contaminants to groundwater quickly (hours ~ weeks; Table 3.1).
Four pathways for surface waters: Horizontal flow is the most dominant pathways for surface waters govering the transport and fate of contaminant in the systems. The horizontal flow pathways are
- Overland flow,
- Groundwater discharge and
- Tile-drainage flow (Figure 3.1).
Overland flow is water flowing along the land surface directly into the stream. This occurs in some landscapes, where the groundwater table is near the land surface (peat soils) or the top soil is extremely impermeable e. g., clay rich (Figure 3.1). The transit time of overland flow is extremely short (Table 3.1) and the water will be continuously exposed to fertilisers or pesticides along the pathways. Furthermore, overland flow causes soil erosion, which may transport nitrogen and pesticides in the particle phase. Therefore, a hydrogeologic system with overland flow is expected to be highly dynamic and vulnerable to contamination.
Interflow occurs in the unsaturated zone where infiltrated water flow laterally via preferential pathways and travel directly to the stream (Figure 3.1). Interflow may develop a localised and transiently saturated zone and water may flow relatively fast (e g., days to weeks; Table 3.1).
Groundwater discharge is laterally groundwater flowing directly into surface waters such as streams and lakes. The transit time of groundwater to surface waters may vary but, among the four pathways to surface water, it is the slowest pathway (Table 3.1). Therefore, in a groundwater-dominated system, stream discharge and chemistry may response to the rainfall inputs slowly.
Tile-drainage may be another important pathway in clay rich soils. It may operate in a similar manner to that of interflow (Table 3.1).
Table 3.1: Definitions and qualitative transit time of pathways
|Water resource||Pathways||Definition||Transit time|
|Surface water||Overland flow||pathways along the land surface||very short (~hours*)|
|Interflow||pathways through the unsaturated (or partially saturated) subsurface, mainly via soil macro-pores, fractures, or perched groundwater||short (days~weeks*)|
|Tile drains||pathways via tile drains||short (days ~weeks*)|
|Groundwater flow||pathways through the saturated zone||long (years~decades)|
|Groundwater||Preferential flow||Flow paths via fractures and macro-pores||short (hours~weeks)|
|Matrix flow||Flow paths via matrix (i.e. via pore spaces)||long (years~decades)|
*) These time specifications are approximations to compare the paths to each other. These approximations are only valid when the flow of water is sufficient during wet conditions/heavy rainfall. This is not always the case. The transit time is dependent on the water flows (in this case rainfall).
Input and output of N in the agri-hydrogeochemical system
N is introduced to the agri-hydrogeochemical system as fertilisers (mineral and organic fertilisers), atmospheric deposition, and biological N fixation. N is removed from the agri-hydrogeochemical system by crop and animal production, manure export, and biological denitrification (Figure 3.2).
Fate of N in the agri-hydrogeochemical system
N is present in soils in both mineral (e. g. nitrate or ammonium,) and organic forms (e. g. urea, crop residue, manure, soil organic matters). Some organic fertilisers (like urea) rapidly hydrolyze into ammonium after application and are therefore sometime consideded as mineral fertiliser. The total N stock in soil (Nt) is almost equal to organic N. Mineral N (Nm) is readily soluble and bioavailable, and so it can be taken up by crop. There are, however, a range of situations where the crops can not absorb all nitrogen. This may enhance leaching of nitrogen.
Organic N must be mineralised to NO3- or NH4+ first via microbial organisms to be utilised by plants and transported with water (Figure 3.2). Rates of microbial N mineralisation may depend on environmental conditions (e. g. climate) and types of organic N. For example, the mineralisation rate is higher in warm and humid climate conditions. Liquid manure with a narrow carbon/nitrogen ratio results in higher mineralisation rates than farm yard manure containing straw.
Mineral N can also be immobilised by soil micro-organisms (Figure 3.2). For example, immobilisation of mineral N occurs when cerial straw (high C/N ratio) is incorporated into the soil after harvest.
Denitrification reactions are the main pathways to remove nitrate from the agri-hydrogeochemical system as gaseous N, i. e. dinitrogen (N2), nitrous oxide (N2O), nitrogen oxides (NOx). Types and rates of microbial N denitrification reactions depend on available energy source (e. g. organic carbon, pyrite) and the redox condition. Denitrification reactions occur only under the redox condition (no oxygen present). Such conditions develop mainly in the root zone layer and deep saturated zone. Denitrification also occur in the upper part of the root zone during wet condtions (e. g. heavy rainfall).
In the soil layer, microorganisms reduce nitrate by oxidising organic carbon (Figure 3.2). In this layer, the influx of fresh organic matters from the ground surface fuel the microorganisms and the reduced condition may develop at a micro-scale.
In the reduced saturated zone, in contrast, organic matter usually is less available to fuel the microorganisms. In this layer, microorganisms use different energy sources. A coupled pyrite oxidation and nitrate reduction is one of the well-known reactions to remove nitrate in the deep saturated zone (e. g. Figure 3.2; Postma et al., 1991).
In the unsaturated zone and the oxic saturated zone, nitrate travels conservatively. In the unsaturated zone, the soil air usually contains oxygen, and the oxic groundwater is defined by the presence of detectable dissolve oxygen. Therefore, nitrate concentrations in water in the unsaturated zone and the oxic groundwater may show the similar level to that of water leached out of the root zone.
Pathways of N in the hydrogeochemical system
Nitrates travel vertically via matrix flow and/or preferential flow in the hydrogeochemical system (Figure 3.2). For groundwater, nitrate concentrations will vary with depth. Despite intense denitrifying activities in the root zone, nitrate concentrations will be highest in the root zone due to high N input. In the unsaturated zone and oxic groundwater, nitrate concentrations in water may be similar to the concentrations in the leaching from the root zone. NO3 concentrations becomes negligible in the reduced saturated zone.
In the saturated zone or groundwater nitrate reduction occurs in a transition zone with anoxic condition often called the redox interface. Here the nitrate concentrations are lower than the concentrations in the leaching from the root zone and often nitrite is used as an indicator anoxic nitrate reducing condtions.
For surface water, nitrate concentrations will vary depending on the dominant pathways. The pathways through the zone between the soil surface and oxic groundwater – overlandflow, interflow, tile-drainage, and oxic groundwater – will deliver nitrate while the reduced groundwater will not be a pathway of nitrate to the surface water (Figure 3.2).
Reasons for N leaching/runoff from the agricultural system
An important reason for the leaching of nitrates below the root zone is the fact, that in majority of soils most of the soil nitrogen (Nt) is organically bound as soil biomass (humus) and slowly released over time by microbial degradation and transformation into mineral nitrogen (Nmin). Only the microbial biomass itself is fast degradable (Nfast) (Beisecker et al., 2015). The release rate of the organically bound nitrogen is linked to the microbial activity in soil: it is high under warm and humid climate conditions and low during cold and dry weather. Small changes in the climatic frame conditions lead to a relatively large alteration of the N release rate. The larger the Nt-stock in soil, the less predictable the amount of N which will mineralise during one growth season (Figure 3.3). With advancing climate change, the situation may even become worse as climate conditions become less predictable.
N-fertiliser planning usually takes into account a certain amount of Nmin in soil at the beginning of the vegetation period and a certain amount of mineralised nitrogen during the plant growth. The difference between plant need and this soil borne nitrogen should be met by N fertilisation (Figure 3.4). Nitrates leach out of the agricultural system due to the unpredictablility of both, the plant development during the growing season and and of the amount of nitrogen which will be mineralised.
A reasonable efficiency of N fertilisation (quota of N in harvested field products versus the amount of fertilised N) is around 50 to 60 % under central European conditions, which means around 40 to 50 % of the applied nitrogen is not harvested as crop product. In Denmark the N use efficiency in 2014 was around 40 % (Hansen et al, 2017). Besides the above explained factors this is due to the following reasons:
- Ammonia losses, especially if fertiliser application techniques are used which do not insert nitrogen fertilisers into the soil.
- Denitrification losses from the top soil, e. g. after application of nitrate containing fertilisers to soil rich in organic matter (high denitrifcation potential), such as grasslands and peat soils.
- Plant growth/development and in consequence N plant need is not exactly predictable, as it depends on a range of factors (e.g., climate, plant health etc.).
- An imbalance between nitrogen demand and supply: this may be absolute, in case nitrogen (as mineral fertiliser or manure) is applied under positive yield estimates.
- The imbalance may be relative, when nitrogen supply and plant demand do not match in course of time (due to N mineralisation or not appropriate timing of N fertilisation).
- The harvested crop contains not all fertilised N, part of it remains in roots and plant parts which remain on the field (i.e. straw, leaves).
- The type of organic fertiliser used: liquid manure or digestate with a narrow carbon/nitrogen ratio posess a higher N-release rates than solid organic fertilisers, such as compost (Gebauer and Schaaf, 2017).
- Point sources, e. g. grazing animals with access to a small stream or lake, may cause nitrogen pollution to surface waters (Bohner et al., 2007).
Inputs and outputs of pesticide to/from the agri-hydrogeochemical system
Pesticides are introduced to the agri-hydrogeochemical system by pesticide application, atmospheric deposition and drift (Figure 3.5). Pesticides can be removed from the agri-hydrogeochemical system by crop production (accumulation in crops); however, removal by crops may be a minor loss compared to the overall pesticide cycle (Figure 3.5).
Fate of pesticides in the agricultural system
There are approximately 250 active substances approved by the EFSA for use in the European Commission (2018a) and these substances show a wide range of physical, chemical, and biological properties. Nitrogen, for instance, is transported predominant as nitrate in water. Pesticides, on the contrary, are transported as gas, particles, and solutes. To a less extent, pesticides can be transported via biota. In addition, pesticides undergo more complex physicochemical and biogeochemical transformation than nitrogen does.
Here, we provide a general overview of the pesticide cycle in the agri-hydrogeochemical system from the ADWI perspectives. We focus on two properties of pesticides that may control the appearance of pesticides in water: persistence (or degradability) and mobility.
The fate of pesticides is controlled by three types of processes: transfer, transport, and degradation processes (Figure 3.6: Fent 2005; Gavrilescu, 2005). Via transfer processes, pesticides move among different environmental media such as air, soil, water and biota. In each medium, pesticides undergo different degradation processes. Transfer processes are responsible for moving pesticides from the initial sources. Persistence and mobility of a pesticide are governed by interactions between the pesticide’s property and these processes.
Persistence. Persistence is degradability of a pesticide by transformation and degradation processes. Via these processes, the structure of a pesticide breaks down and its toxicity usually decreases. The physical and chemical characteristics of pesticides may be the first order control for their degradability. In general, more reactive ones (e. g., soluble, small-sized compound, aliphatic structure) are more easily degradable. Table 3.2 summarises the pesticides properties that may control the degradability of organic pesticides (Gavrilescu, 2005).
Table 3.2: Physical, chemical and structural characteristic that may control degradability of organic pesticides (modification of Table 5 from Gavrilescu, 2005)
|More easily||Less easily|
|Solubility in water||Soluble in water||Insoluble in water|
|Size||Relatively small||Relatively large|
|Functional group substitutions||Fewer functional group||Many functional groups|
|Compound more oxidized||In reduced environment||In oxidized environment|
|Compound more reduced||In oxidized environment||In reduced environment|
|Created||Biologically||Chemically by man|
|Structure||Aliphatics (branch structure)||Polyaromatic (ring structure)|
The degradation processes are divided into three categories: microbial degradation, chemical degradation, and photodegradation (Gavrilescu, 2005).
Microbial degradation is the primary process to degrade pesticides in soil and water. Soil biota, such as microorganisms, bacteria and fungi may use pesticides as a source of energy or degrade pesticides while using other energy sources such as organic carbon. The rates of microbial degradation will be highest under a favorable condition for soil biota such as a warm, moist and neutral pH environment (Gavrilescu, 2005).
Chemical degradation is an abiotic process, including hydrolysis, oxidation-reduction reactions, and ionization (Gavrilescu, 2005). Pesticides can be degraded by sunlight. Photodegradation occurs not only in the air but also in the shallow soil where photons can penetrate (Gavrilescu, 2005).
Mobility. Pesticides in soils exists both as particles and as solute and the phase influences their mobility and transport mechanisms. Table 3.3 summarises the key properties of pesticides and environmental conditions that affect the mobility.
Table 3.3: Key pesticide and environmental characteristics that control pesticide mobilisation (modified from Gavrilescu, 2005)
|mobilised as solutes||mobilised as particles|
|organic carbon-water partitioning coefficient (Koc)||low||high|
|organic matter content||low||high|
The organic carbon-water partitioning coefficient (Kow) of pesticide is the ratio of the concentration of a chemical compound in the n-octanol phase to its concentration in the aqueous phase at equilibrium in a two-phase octanol/water system. Kow is usually expressed as logKow, which is inversely related to water solubility. It is frequently used to predict the distribution of a substance in water and soil.
It is related to the soil adsorption coefficient (Kd) which is described as:
Kd = Concentration of compound in soil / Concentration of compound in water
Kd usually varies greatly because the organic content of is extremely variable also. Nevertheless, adsorption occurs predominantly into the organic matter of the soil, therefore it is more useful to express the distribution coefficient in Koc.
Koc is also known as organic carbon-water partition co-efficient and is described as:
Koc = (Kd * 100)/ % Organic carbon
Koc may be the most important property to determine whether a pesticide is transported as particles or as solutes (Gavrilescu, 2005). The range may not be precise though, in general, pesticides with high Koc values (>100) – the herbicides trifluralin, paraquat and glyphosate – are likely adsorbed onto soil particles and lost via erosion. Such pesticides may accumulate in the soil and degrade over time releasing daughter compounds. While pesticides with intermediate Koc values (0.1< K<100), which are most herbicides today, are primarily lost with water (Fawcett et al., 1994).
Water solubility may be another important property to determine how easily a pesticide can be transported in water. In general, a pesticide with water solubility of less than 1 ppm is likely to adsorb onto soil particles (Gavrilescu, 2005). If a pesticide is not persistent, because it is transformed into different forms, it will less likely be mobilised either by solutes or particles.
Soil texture, pH and organic contents may also be important environmental factors to determine the degree of adsorption of pesticides (Gavrilescu, 2005; Table 3.3). The soil texture controls the available surface areas where adsorption can occur. Soil pH affects the pesticide solubility and microbial degradation rates; consequently, the adsorption rate will change. In general, in acidic soil, a pesticide is more soluble and microbes degrade the pesticide faster. Organic matter in soil provides binding sites to pesticides and serves as energy source for microbial reactions/degradation.
Pesticide cycle in the agri-hydrogeochemical system
The persistence and mobility of pesticides are mainly determined in the root zone. Below the root zone, pesticides are mainly conservatively transported. Microbes play the dominant role in degrading and transforming pesticides (Fenner et al., 2013; Gavrilescu, 2005). Although some researchers reported that microbial degradation in groundwater is potentially possible (Janniche et al., 2012), due to scarcity of energy source and nutrients, the rates of microbial degradation below the soil layer is generally insignificant compared to that in the soil layer. Therefore, once pesticides leach out of the root zone, they are redistributed without any significant transformation or degradation (Figure 3.5).
After a pesticide is applied in an agricultural field, it can be released back to the air via evaporation and volatilisation. In addition, it can be degraded by light in the shallow soil layer and be emitted to the atmosphere (Figure 3.5). Some fractions of the pesticide may accumulate in crops. Depending on its property, the pesticide can be adsorbed onto the soil particles (PES in Figure 3.5) or dissolved in soil pore water (PEW; in Figure 3.5). MSoil microbes can degrade the pesticide, producing daughter compounds (PED; Figure 3.5).
Pathways of pesticides in the agri-hydrogeochemical system
A potential pathway of pesticides through the atmosphere is spray drift (Figure 3.5). When a pesticide is applied as spray, it can be drifted directly into the adjacent surface water especially under windy conditions (Carter, 2000).
Surface erosion is a pathway to transport pesticides that are adsorbed onto soil particles (Figure 3.5). The dissolved phase of pesticides and the daughter compounds are transported via water pathways: pesticides are transported vertically via matrix flow or/and preferential flow and laterally via overland flow, interflow, tile-drainage, and groundwater (Figure 3.5).
The application of pesticides for plant protection purposes on agricultural fields is a diffuse source (Carter, 2000). On the contrary, point sources are localised situations such as tank-filling respectively cleaning, farmyard-runoff and spills from agricultural sources, fruit washing facilities or even sewage plants (Carter, 2000). Point sources are mainly due to misuse or inadequate management. Especially after heavy precipitation events, farmyard runoffs together with field runoffs produce contamination peaks and account for most of the contaminant load of small streams. In a catchment area with individual agricultural farms scattered and no other possible contamination source present, farmyard runoff accounted for 89.8 % of pesticide contamination, especially fungicides and insecticides (Neumann et al., 2002).
Pesticides of diffuse pollution will be transported via the pathways mentioned above, depending on their properties and environmental conditions, but pesticide pollution from point sources are mostly directly transported into water (Carter, 2000). For instance, poor management of filling/cleaning facilities may result in discharge of pesticides along the impermeable surface or via pipes (Wenneker et al., 2010), acting like overland flow. Illegal discharge of pesticides will directly flow into stream or groundwater, bypassing all the pathways (Carter, 2000).
Issues following an inappropriate usage of pesticides and alternative entry paths into the environment are addressed in the EU guideline 2009/128/EG. The Member States are obliged to transform the guideline into National Action Plans, to introduce measures to protect aquatic environments and to organise the education of pesticide applicants in the correct handling, disposal and cleaning of pesticide application devices.
Challenges in pesticide monitoring and regulation
Up to present, the pesticides that are found in the different environmental compartments can only sporadically be related to application data of the pesticides, since a regional differentiated data compilation of application data and a consequential estimation of the pesticide inputs is missing (SRU, 2016).
According to the approval procedure and a proper usage of pesticide products, no pesticide transport to surface or groundwater and no accumulation in soil should take place. However, EU-wide, a number of pesticides are detected in surface- and ground waters, the most abundant are listed in Table 3.4.
Pesticide contamination in surface waters being reported by EAA (WISE-databank, reports by Member States) are in their large majority due to substances, which are withdrawn from the market some time ago. In most sites, this is due to occationally high quantities of pesticides contained in the water table that feeds surface water. In some sectors, also fraudulent use of pesticides had been proven (Laurent, 2015).
Table 3.4: Most abundant pesticides being detected in surface- and ground waters: number of waterbodies (WB) not achieving a good chemical status due to pesticides
and number of Member States (MS) affected (EEA, 2018a; University of Hertfordshire, 2017)
|Substance||CAS- or EEA-No.||Chemical group||Type||Examples for product with AS||Introduced to the market||Current market situation||No. of WB not achieving good chemical status||No. of MS with WB not achieving good chemical states for the listed substance|
|Isoproturon||34123-59-6||Urea derivative||Herbicide||Arelon, Azur, Alpha IPU, Alpha Isoproturon, Koala, Trump, Javelin, Javelin Gold, Protugan, Tolugan Extra||1971||approved EU-wide since 2003, with exeption of CY, DK, EL, FI, MT; withdrawn in 2016||199||8|
|Hexachlorhexane||608-73-1||Halogenated hydrocarbon, organochlorine||Insecticide, acaricide||Lindan||1945||on the market since 1945; acording to EC 1107/2009 not approved||120||11|
|Trifluralin||1582-09-8||Dinitroanaline||Herbicide||Alpha Trifluralin 48EC, Ardent, Fargro Axit, Treflan, Uranus, Elancolan||1961||approval withdrawn in 2007 acording to EC 1107/2009||12||6|
|Chlorfenvinphos||470-90-6||Organophosphate||Insecticide, acaricide, veterinary substance||Vinylphate, Birlane, Steladone, Supona, Apachlor, Haptarax||1962||not approved||10||4|
|Atrazine||1912-24-9||Triazine||Herbicide||Gesapri, Fenamin, Atrazinax, Weedex, Primaze, Atratol, Radazine||1957||not approved||9||4|
|Simazine||122-34-9||Triazine||Herbicide||Sanazine, Simanex, Amizina, Eagrow, Derby||1960||not approved||5||2|
|Alachlor||15972-60-8||Chloroacetamide||Herbicide||Lasso, Alanex, Pillarzo||1936||not approved||5||3|
|Pentachlorphenol||87-86-5||Organochlorine||Insecticide, Herbicide, fungicide, molluscicide, plant growth regulator, wood preservative||1936||not approved||3||3|
|Pesticides||EEA_34-01-5||Active substances in pesticides, including their relevant metabolites, degradation and reaction products||345||11|
|Bentazone||25057-89-0||Benzothiazinone||Herbicide||Basagran, Zone 48, Troy 480, Herbatox, Leader, Laddox||1972||approved in all EU-countries||31||5|
|Atrazine||1912-24-9||Triazine||Herbicide||Gesaprim, Fenamin, Atrazinax, Weedex, Primaze, Atratol, Radazine||1957||approval expired||60||8|
|Desethylatrazine||6190-65-4||Dealkylated atrazine metabolite||Plant growth inhibitor||69||5|
|Terbuthylatrazine||5915-41-3||Triazine||Herbicide, microbiocide, algicide||Calaris, Skirmish, Gardo Gold||1967||expired in DK, EE, FI, FR, LT, LV, MT, SE||25||4|
|Bromacil||Uracil||Herbicide||Hyvar X bromoacil, Borocil 1V, Cynogan, Borea, Krovar II, Urgan||1961||expired||13||5|
|Simazine||122-34-9||Triazine||Herbicide||Sanazine, Simanex, Amizina, Eagrow, Derby||ca. 1960||expired, except for ES||17||5|
|Metholachlor||51218-45-2||Chloracetamide||Herbicide||Dual, Bicep, Pennant, Pimagram||1976||expired||58||3|
|Alachlor||15972-60-8||Chloracetamide||Herbicide||Lasso, Alanex, Pillarzo||1969||expired||63||1|
|Acetochlor||34256-82-1||Chloracetamide||Herbicide||Harness, Trophy, Trophee, Acenit, Guardian, Sacemid, Surpass||1985||expired||32||1|
|Dicamba||24-00-9||Benzoic acid||Herbicide||Di-Farmon R, Foundation, Prompt, Relay P, Banval||ca. 1963||expired in MT and SE||22||2|
Using the WISE-databank, the causes of pesticide contamination of ground waters cannot be identified precisely, because obviously some of the Member States reported under a collective term, other reported the analyses of certain active substances. Most of the pesticides being reported contamining groundwater are not any more approved by EFSA-authorities (Table 3.4).
Tauchnitz et al. (2017) investigated in the German Harz foreland pesticide concentration in soils and surface waters. In surface waters, Glyphosate, Bentazone, AMPA, Diflufenican, Tebuconazol and Terbutylazin were detected. There was no correlation between agricultural application and detection of pesticides, possibly due to pesticide use/leaching from residential areas. Agricultural activities were clearly the reason for accumulation of pesticides in soils underneath agricultural activities. Especially Glyphosate and MCPA were found in depths of nearly five meters, S-Metolachlor and Pendimethalin in around one meter depth.
Ulrich et al. (2018) report the accumulation of herbicides and their transformation products in small water bodies or catchment areas. The autumn sampling focused on the herbicides Metazachlor, Flufenacet and their transformation products – Oxalic acid and – Sulfonic acid as representatives for common pesticides in the study region.
Agriculture has a direct influence on the concentrations of nitrates and pesticides in the raw water used for drinking water production. However, the raw water, coming from either groundwater or/and surface water, might be treated at the waterworks before delivered as drinking water to the consumers. Therefore the specific water treatment procedure at the waterworks is important for the final concentration of nitrates and pesticides in drinking water.
In Europe, drinking water is produced with different degree of treatment (Van Der Hoek et al., 2014; amended, Table 3.5):
- Without treatment
- With conventional treatments such as aeration and sand filtration;
- With advanced treatments such as active carbon filtration, advanced oxidation process (e.g., UV/H2O2, ozonation), desalination; and
- With combination of conventional and advanced treatments
- Excluding of polluted wells
Because nitrate is highly soluble and pesticides are persistent, the conventional treatments cannot remove them. Nitrate in water is removed via ion exchange, reverse osmosis, electrodialysis, biological/chemical/catalytic denitrification and combination among them (Kapoor et al., 1997).
The ion exchange process is running NO3- containing raw water through exchange resins, which contain strong base anions such as Cl- and HNO3-; therefore, nitrate is replaced with these anions. The reverse osmosis process is filtering out ions by pushing water through a semipermeable membrane. The electrodialysis process is transferring ions in a diluted solution to a concentrated one through a membrane with a direct electric current. The biological, chemical, and catalytical denitrification processes are denitrifying nitrate by biological (e.g. microbes), chemical (e.g. Fe (II)), and catalytical (e.g. palladium-alumina; Pd-Al2O3) agents, respectively.
To remove pesticides in water, advanced oxidation processes, coagulation-flocculation-sedimentation, nanofiltration, and active carbon adsorption are used (Ormad et al., 2008). The advanced oxidation process is to break down pesticides into biodegradable compounds using strong oxidants such as chlorine or ozone. This process is often combined with biological treatments; therefore, it is also called as a ‘preoxidation process’. The coagulation-flocculation-sedimentation is a physical and chemical process to remove pesticides by forming larger particles so that they can easily be separated out. The nanofiltration is filtering out pesticides using a filter membrane that has extremely small pore sizes. Pesticides can be removed by adsorption on the active carbon. These techniques are often used in combination.
Table 3.5: Summary of drinking water treatament methods in Europe (modification of Table 1 in Van Der Hoek et al. 2014)
|Water process method||Groundwater||Surface|
|Conventional treatment||Aeration and/or Rapid Sand Filtration (RSF)||Coagulation, sedimentation, and filtration (CSF)|
|Advanced treatment||Carbon filtration, advanced oxidation process, membrane desalination, ion exchage, reverse osmosis, electrodialysis, nitrification, coagulation-flocculation-sedimentation, nanofiltration||Carbon filtration, advanced oxidation process, membrane desalination, ion exchage, reverse osmosis, electrodialysis, nitrification, coagulation-flocculation-sedimentation, nanofiltration|
|Conventional + advanced treatment||Aeration and/or RSF + advanced treatment||CSF + advanced treatment|
Table 3.6: Cost of water treatments in France (Juery, 2012)
|code1||Water intake + disinfection||0.05|
|code2||Pretreatment + coagulation/flocculation + sedimentation + sand filtration + disinfection||0.13|
|code3||Pretreatment + coagulation/flocculation + sedimentation + preoxidation + sand filtration + O3/activated carbon filters treatment + disinfection||0.20|
|code4||Code 3 without preoxidation + biological denitrification||0.38|
|code5||Pretreatment + coagulation/flocculation + sedimentation + sand filtration + disinfection+microfitration+nanofitration||0.43|
|code6||Ultrafiltration + disinfection||0.50|
The level and type of water treatment of each country may depend on the quality of raw water, its financial circumstances, and strategic political decisions (WHO Regional Office for Europe, 2002). In general, advanced treatments (e. g. nanofiltration, advance oxidation process) are more expensive than the conventional treatment techniques. An example of treatment cost is shown for France in Table 3.6 (Juery, 2012). Disinfection alone costs in average 0.05 (€/m³) wheras ultrafiltration + disinfection cost in average 0.50 (€/m³) (Table 3.6). The costs of water treatment in other countries will differ from those in France (Table 3.6), because of differences in local conditons. Table 3.7 shows main types of drinking water treatments used in some countries (not specific to the case study sites) that are part of FAIRWAY (WHO Regional Office for Europe, 2002).
Table 3.7: Main types of water treatment methods for drinking water production (modification of Table 4.5 in WHO Regional Office for Europe, 2002)
|Country||Groundwater/spring water||Surface water|
|Germany (UBA 2016, 2018)||
One of the challenges of drinking water production is disinfectant by-products (DBPs) such as thrihalomethanes (THMs) and haloacetic acids (HAAs; e.g. Van Der Hoek et al., 2014; WHO Regional Office for Europe, 2002). Disinfection processes are intented to eliminate pathogens in drinking water (Safe Drinking Water Committee, 1980) and it is particularly important for surface water. Van Der Hoek et al. (2014) reported that nearly 88 % of drinking water production – nearly 99.99 % from surface water and more than 70 % from groundwater – in Europe employes a disinfection process. Raw water is disinfected primarily via oxidation reactions using strong oxidants such as chlorine compounds (e. g. chlorine, hypochlorite, chlorine dioxide), ozone (O3), UV/H2O2 (World Health Organisation, 2000). Although it may be a minor risk in comparison to preventing waterborn diseases (WHO Regional Office for Europe, 2002), oxidation of some pesticides during the oxidation process can produce DBPs (e. g. Adams and Randtke, 1992; Chiron et al., 2000; Huang et al., 2009; Li et al., 2016). Nitrate may indirectly play a role in producing DBPs: excessive nitrate, along with phosphorus, in raw water (i. e. surface water) can trigger algae blooms. Then, while the algae cells are destroyed at the disinfection process, DBPs can be produced as well (e. g. Huang et al., 2009; Plummer and Edzwald, 2001).
In Denmark, groundwater protection has a high priority in order to secure clean drinking water for the population provided directly from the taps in the houses. Accordingly, the Danish groundwater protection policy is based on prevention rather that treatment at the waterworks.
Waterworks in Germany, according to a recently conducted nationwide survey, tend to avoid the implementation of expensive treatment measures in order to reduce high nitrate concentration in drinking water (Oelmann et al., 2017). Figure 3.7 shows clearly, that the so called “Preventive measures” (consulting farmers, cooperation between farmers and water works, including – financial – support, purchase or lease of land) are far more common than the so called “Reactive measures” (mixing, avoiding. e. g. excluding the polluted well, advanced treatment).
Note: For full references to papers quoted in this article see