The risk of herbicides in groundwater

Chemical control in agricultural practice has a recent history, but it has increased greatly in the last few decades. At the end of the last century in Europe, copper sulphate was already being used to control broadleaved weeds in grass crops, but it is only since the 1950s that selective herbicides have been introduced onto the market. The enormous success of these products was an incentive for research and development and this led to a number of new herbicides belonging to different chemical classes being marketed. This was followed by a big expansion of the use of chemical products for weed control worldwide. In developed countries, herbicides are used on 85% to 100% of all main crops. In the 1994 Annual Index of Weed Abstracts, 333 active ingredients are listed, and about 200 of these are widely distributed and used all over the world.

Although the environmental risk from pesticides was recognised relatively early on, the occurrence of pesticides in groundwater was not detected until much later and the major concern was directed towards DDT and other persistent organochlorine compounds. This was due mainly to lack of knowledge about the important features of chemicals movement through the soil, and gave rise to the general misconception that less persistent pesticides could not leach into the groundwater under normal conditions. One of the first references to the discovery of pesticides other than chlorinated hydrocarbons in groundwater was by Richard et al. (1975). In the late 1970s, however, the number of detections increased rapidly, along with concern from public authorities about chemical contamination of groundwater.

Introduction

In Italy, atrazine has been used extensively in maize cultivation since the early 1960s. One of the advantages of using atrazine in maize cultivation is flexibility in the timing of the application. Unfortunately, however, its chemical properties make it a likely source of groundwater contamination, especially in areas with permeable soils. In some areas of the Po Valley in northern Italy, where maize cultivation is extensive and groundwater the dominant source of drinking water, chemical monitoring in the 1980s revealed that concentrations of atrazine exceeded 0.1 μg/l i.e. the maximum admissible concentration set by the Directive on Drinking Water Quality of 15 July 1980 (see Vighi and Zanin, 1994, for examples of monitoring results). Following the implementation of local restrictions in 1986, a nationwide ban on the sale and use of atrazine was introduced in 1990.

An earlier pilot study was conducted to address the question of whether the EC standard, with specific reference to atrazine in Italy, could be justified in terms of social efficiency; that is, is 0.1 μg/l a socially efficient contamination level (Bergman and Pugh, 1997)? Broadly speaking, this question can be answered in the affirmative if 0.1 μg/l is the concentration of atrazine in drinking water for which the marginal cost of reducing the concentration is equal to the marginal benefit of a reduced concentration (see Söderqvist et al., 1995, for a basic introduction to these issues). We confine ourselves here to noting that reduction costs may, inter alia, be due to farmers having to turn to more costly weed eradication methods, and that reduced health risks may be one important constituent of reduction benefits.

Introduction

Chapter 7 showed that market failure (the presence of externalities and imperfect competition in product markets) may give rise to agricultural chemicals which will accumulate in the environment to a socially excessive level. It is the role of governments to intervene to construct institutions to counteract the market failure. Yet all too often, government intervention serves only to exacerbate the original problem by supplying an inappropriate or ineffective institution; such an event is termed a ‘regulatory failure’. In this chapter, various sources of regulatory failure are examined.

In many cases, governments have banned particular chemicals deemed to have accumulated to excessive levels in natural resources. Rölike (1996) reported that by 1993, 78 countries had used this type of regulation. In Germany, 300 of the (approximately) 1000 pesticides registered have been banned individually. Many European countries have responded to a European Commission Directive on drinking water by banning the sale and use of particular chemicals (see p. 218). The objective of this form of regulation is to reduce chemical accumulation by two means: firstly, by removing from use chemicals already present at high concentrations; and, secondly, by providing a signal of society's disapproval of excessive accumulation. It also has the advantage of a degree of certainty (although even this is mitigated by the persistence and continued accumulation of products for several years after the imposition of a ban).

This chapter assesses the combination of maximum acceptable concentrations (MACs) and product-specific bans (PSBs) used by regulators.

This book arises from a meeting of the same name held at the Postgraduate Research Institute for Sedimentology, Reading, UK, in April 1994. The meeting was sponsored by the Challenger Society and the Geochemistry Group of the Mineralogical Society: we are grateful to both these organisations for their generous support. Eight scientists active in research on the biogeochemistry of intertidal systems were invited to present papers on their own particular interests within the field. We deliberately selected a mixture of established and young UK scientists and invited them to both briefly review the field and to present some new data from their own research. These papers have now been formalised into this book together with an overview chapter written by ourselves. The various authors have taken different approaches to the task we set them: some have achieved a balance between review and current research, whereas others have chosen to concentrate mainly on either review or recent research. As editors we have encouraged their diversity, allowing the authors the freedom to deal with the subject as they see best. We wish to thank all the authors for their good humour and hard work. All the chapters have been peer reviewed so we are additionally grateful to D. Cooper, A. Grant, J. Hamilton-Taylor, D. Hydes, S. Malcolm, R. Mortimer, J. Scudlark and S. Wakefield. (Chapter one was reviewed by the authors of the other book chapters.)

Introduction

Many inorganic and organic pollutants are of great concern to water quality managers owing to their persistence, toxicity and liability to bioaccumulate (Tanabe, 1988; Robards & Worsfold, 1991; Bryan & Langston, 1992; McCain et al, 1992; Förstner, 1993). The major temporary or ultimate sink for such pollutants in estuaries is the sedimentary reservoir, including intertidal deposits (Hanson, Evans & Colby, 1993; van Zoest & van Eck, 1993; Kennicutt et al., 1994), and definition of the biogeochemical mechanisms by which they absorb onto, desorb from and repartition amongst natural, heterogeneous particle populations is essential in order to assess their environmental fate. In estuaries, prediction of the distribution of pollutants is further compounded by intense temporal and spatial gradients of the reaction controlling variables such as salinity, dissolved oxygen concentration and particle composition, occurring both within the sediment and in the water column during particle suspension.

This chapter focuses on an empirical technique to study the sorptive behaviour of trace metals and trace organic compounds in estuaries. Thus, the partitioning of constituents between particles and solution is determined experimentally, without identifying the inherent reaction mechanisms or reactant speciation, under controlled laboratory conditions using natural samples spiked with radiotracer analogues of the constituent of interest (Turner et al., 1993). Adsorption and desorption processes may be modelled as a function of particle concentration, and the controlled variables of salinity and dissolved oxygen concentration, by incorporating empirically derived results into simple mass balance equations.

In some places the boundary between land and sea is in the form of abrupt and often spectacular cliffs but, elsewhere, the boundary can take the form of a complex environment of intertidal sediments. These environments include shingle banks, sandy beaches, mud flats, saltmarsh and mangrove (or mangal) communities. In some cases one or other of these environments will occur, in others they will be associated with one another. For example, on many North Sea and North American East Coast shorelines, mud flats grade into saltmarshes behind the shelter of shingle spits and sand dunes. In general, saltmarshes and mangroves occupy similar ecological niches with mangroves at lower latitudes (winter temperatures greater than 10°C) and saltmarshes at high latitudes, though in some locations both communities coexist (Chapman, 1977).

The considerable global significance of these intertidal systems is evident from Figure 1.1. Around the coast of Britain alone there are 44 370 hectares (ha) of saltmarsh (Allen & Pye, 1992) and 589 429 ha on the US East Coast (Reimold, 1977) with a total global area of 3.8 × 107ha (Steudler & Peterson, 1984 and references therein). There are 365 500 ha of mangal forest around the Indian subcontinent and another 250 000 ha in the Mekong delta (Blasco, 1977). The total global area of mangal is about 2.4 × 107ha (Twilley, Chen & Hargis, 1992).

Introduction

Trace metals, for example elements from the first row of transition metals including cobalt, nickel, copper and zinc (Co, Ni, Cu and Zn), occur naturally in intertidal sediments. The natural (or ‘background’ inputs) reflect a combination of the composition of drainage-basin rocks and marine-derived sediment. The global anthropogenic input of many trace metals to the environment currently equals or exceeds the amount released by weathering (de Groot, Salomons & Allersma, 1976), and the extent of anthropogenic influence in intertidal sediments is particularly high, since intertidal areas are often considered as convenient dumping grounds for industrial and other waste.

Anthropogenic inputs to intertidal environments are often direct, through point-source waste disposal, but they are also indirect, from riverine, marine and/or atmospheric sources. Trace metals are partitioned between each component of the intertidal sediment–water system: they are found in solution (‘bulk’ water or interstitial water) and associated with suspended and deposited sediments. This chapter is concerned with the biogeochemistry of trace metals in deposited intertidal sediments. Two main sections follow: in the first, an overview of surface sediments and sediment depth profiles is presented, and in the second, a case study is given of the historic record of Zn from saltmarsh sediments in the Severn Estuary, UK.

Introduction

In the context of saltmarsh/coastal flux studies, a flux is the net exchange of a substance between these two ecosystems. The fluxes of several types of material, including nutrients, trace metals, herbicides and radionuclides, have been studied by researchers. Although this paper primarily appraises the methodology of nutrient flux estimations, many of the problems are also shared by researchers working on other types of material fluxes.

Since the first detrital flux studies conducted on Sapelo Island, Georgia, USA (Teal, 1962; Odum & de la Cruz, 1967), there has been a research interest in nutrient fluxes between saltmarshes and the adjacent coastal water. The major drive behind this work was to ascertain whether saltmarshes affect the concentration of dissolved inorganic nutrients and detritus (and hence the ecology) of coastal water. The link between saltmarsh processes and coastal productivity may be of relevance to coastal managers who wish to predict the impact of future saltmarsh loss (due to reclamation, erosion or drowning due to sea level rise) or possibly saltmarsh gain (if managed retreat becomes an extensively used strategy for coastal defence in rural areas).

The published flux results indicate that some saltmarshes appear to process nutrients in a very different way to others (Nixon, 1980; Carpenter, 1993). These discrepancies could be due to actual differences in nutrient budgeting caused by variations in saltmarsh characteristics. However, the disparity may be artificial, caused by differences in field methodology and the large uncertainties involved in the calculation of tidal and annual fluxes.

Introduction

The focus of this chapter is on the behaviour of radionuclides in the coastal and intertidal sediments of the Irish Sea, in particular those in the north-east. A review of previous research concerned with the fate of radionuclides in and around the Irish Sea is presented along with some recent research the authors have performed.

Boundaries of the Irish Sea

The Irish Sea is a semi-enclosed body of water, bounded by the eastern coast of Ireland and the coasts of north Wales, north-west England and south-west Scotland. It connects with the Atlantic Ocean through St George's Channel in the south and the North Channel to the north. Water enters the Irish Sea both through St George's Channel and the North Channel, although the former dominates, and leaves via the North Channel (Howorth, 1984; Dickson, 1987). The flowrate of water out through the North Channel is about 5 × 109m3d-1 and the residence time of the water in the northern Irish Sea is about one year (Pentreath et al., 1984).

In relation to radionuclide behaviour, the distribution of bed sediments according to grain size is of particular importance. Figure 8.1 shows the disposition of bed sediments according to the grain size description of Pantin (1977, 1978). The two main areas of muddy sediments, one extending southwards from St Bees Head, Cumbria, towards Morecambe Bay and the other lying between the Isle of Man and the north-east coast of Ireland, are of particular significance.

Introduction

Intertidal sediments, especially fine-grained estuarine sediments, may be a net sink, or at least a transient repository of hydrophobic organic chemicals (HOC) from a number of sources. HOC have a strong, though variable, affinity for sediment minerals, especially for clays coated with natural organic matter (OM). Measurement of this adsorptive capacity is thus an important goal for geochemists wishing to study the budget and fate of HOC in the intertidal environment.

Although a considerable literature of adsorption measurements exists, the cocktail of HOCs which reaches estuarine waters is so diverse and originates from so many sources (including from municipal discharges, from effluents from industrial works sited on estuaries, and from riverine inputs) that inevitably there are important gaps in our knowledge. Included in this is information about a number of important agrochemical HOCs which may enter estuaries, via riverine transport. This chapter focuses particularly on one group of HOCs (pyrethroids), the chapter by Turner and Tyler discusses the behaviour of some other groups.

The pyrethroid agrochemicals represent a quarter of the world insecticides market and are widely used in the UK, continental Europe, the USA and elsewhere as household, public health and agricultural insecticides (Hill, 1989; Perrior, 1993). The chemicals are described by the generic name ‘pyrethroids’ because they are all man-made analogues of the esters of the natural cyclopropanoic acid, chrysanthemic acid, found in the plant family pyrethrums, but in fact they exhibit a range of physico-chemical properties in the environment.

Introduction

In most marine-type sediments (including intertidal ones) organic carbon is the only reducing agent to enter a sediment column. The remainder of the sediment load arrives in its oxidised form, and, with the exception of straightforward compaction, early diagenesis (i.e. the process of change during burial) results directly or indirectly from the flow of electrons. The initial source of the electrons (organic matter) is sequentially oxidised in microbially mediated reactions, using a range of available oxidising agents, and results in some degree of vertical zonation in sediment chemistry (Richards et al., 1965; Froelich et al., 1979). This is because microbial communities outcompete each other for organic carbon (Stumm & Morgan, 1970; Claypool & Kaplan, 1974). However, the resolution of these sediment horizons is likely to be poor because:

1. (a) although the most important substrate shared by each community is organic carbon, it has been suggested that different communities metabolise different fractions of it, each with a first order rate constant (Berner, 1977). This would produce overlapping or coexistence of zones, unless the competition between communities was for another shared substrate (e.g. nitrate).

2. (b) the ‘dead cat in the mud’ effect (Coleman, 1985), where the substrates for microbial metabolism do not enter the system as disseminated and reactive particles to give a homogeneous reaction mixture, but rather as poorly sorted, highly concentrated sources of metabolites that may survive beyond zones where they would have been exhausted in a simple layered model. For solid substrates such as organic carbon, iron oxides and manganese oxyhydroxides, this may be an important control on their rate of supply to the reaction.

3. […]

Introduction

The relationship between microbial processes, diagenetic reactions and mineralisation within intertidal sediments, including the development of diagenetic carbonate concretions, tubules and localised cemented layers, has attracted major interest in recent years. Such phenomena are relatively common in intertidal sediments around the British Isles and on other European coasts (see e.g. Pye, 1981; Andrews, 1988; Al-Agha et al., 1995; Duck, 1995). The processes associated with their formation are important in relation to the cycling of carbon, sulphur and iron, as well as the fate of industrial and other pollutants. Early diagenesis within saltmarshes in particular is of importance because of the high biological productivity which is characteristic of such environments, and their position at the interface between marine and terrestrial systems (see Vernberg, 1993, for a recent review of saltmarsh processes).

The role of microbial activity in early diagenesis has been recognised for some time, and a variety of approaches has been used to investigate these processes, including, for example, studies of C–S–Fe interactions (Berner & Westrich, 1985), rates of sulphate reduction (Jorgensen, 1978; Jorgensen & Bak, 1991), nitrate reduction (Sorenson, 1987) and Fe/Mn reduction (Lovley, 1991). Recently, significant progress has been made in quantifying the rates of microbial reactions and the interactions between biotic and abiotic factors (e.g. Blackburn & Blackburn, 1993; Parkes et al., 1993a,b), partly as a result of the application of biochemical techniques which allow detailed specification of microbial biomass, community structure and nutritional status (e.g. White, 1993; Eglinton, Parkes & Zhao, 1993).

Introduction

Intertidal sediments occupy a unique position between land and sea giving them an importance which is only now becoming apparent. Most studies have looked in detail at the nature of the biogeochemical processes (e.g. Klump & Martens, 1989) but have not had the impetus to place the processes into a quantitative context. The priority and scale of study has changed due to an awareness that intertidal areas are the front line in dealing with both the effects of climate change and the movement of materials from land to sea (Nowicki, 1994). There are now global, regional and national research programmes examining interactions at the land/sea boundary (e.g. LOICZ (Pernetta & Milliman, 1995), NECOP (Ortner & Dagg, 1995), JoNuS (Anon., 1994), LOIS (Anon., 1992)).

Concern about the impact of nutrients on the marine environment is providing a sharp focus for some of these studies (e.g. Nowicki & Oviatt, 1990; Anon., 1994). Intertidal sediments are a part of the sediment system that has long helped to maintain the productivity of the coastal zone via the storage and recycling of nutrients which were imported from offshore waters (Nixon, 1986; Nixon, 1992). Human population increases, coupled with aquatic sewerage systems and the development of intensive agriculture, have resulted in a significant change in the quantities of nutrients supplied from land sources. Now the main source of nutrients is from the land and not from the sea (Nixon, Hunt & Nowicki, 1986).