In the fall of 1960 thousands of turkey poults and other animals started dying throughout southern England. Veterinarians were at first stymied about the cause of what they came to label “Turkey X disease,” but because so many birds were affected, a major investigation into its origins was undertaken. In 1961 a report from three scientists at London's Tropical Products Institute and a veterinarian at the Ministry of Agriculture's laboratory at Weybridge, entitled “Toxicity associated with certain samples of groundnuts,” was published in the internationally prominent scientific journal, Nature. Groundnuts, as everyone in America knows, are actually peanuts, and peanut meal is an important component of animal feed. It appeared that the turkeys had been poisoned by some agent present in the peanut meal component of their feed. The British investigators found that the poisonous agent was not a component of the peanuts themselves, but was found only in peanuts that had become contaminated with a certain mold.

It also became clear that the mold itself – identified by the mold experts (mycologists) as the fairly common species Aspergillus flavus – was not directly responsible for the poisoning. Turkey X disease could be reproduced in the laboratory not only when birds were fed peanut meal contaminated with living mold, but also when fed the same meal after the mold had been killed.

Chemists have known for a long time that molds are immensely productive manufacturers of organic chemical agents.

Before looking to the future, it will be useful to look back at some issues related to risk assessment and management that have so far been ignored. The issues are not so much technical as they are social and political, and to ignore them completely could leave the misleading impression that all the scientific and policy questions we have discussed are interesting matters for scholarly debate, and not much else. We shall not make this mistake.

Risk assessments reveal public health problems, of greater or lesser magnitude. If a problem is uncovered, we cannot simply hide it (at least not easily); we need to do something to reduce or eliminate it. Somebody will have to pay, no two ways about it. Depending upon the problem, costs could be massive for society as a whole, massive for selected industries, or, at the other extreme, relatively small all round. The latter generally raises only a little smoke, but when costs are heavy, things burst into flames. Because the industries that must bear the cost do not wish to be seen as destroyers of the public health or the environment, some tend to begin by determining whether there are credible ways to attack the scientific quality and accuracy of the risk assessments regulators are relying upon. They may claim risks have been exaggerated, that there is no, or only a minor public health problem.

The central topics of this book have, since its publication in 1992, become permanent occupants of the public health agendas of governments everywhere. The imposition of controls on the production, uses, and environmental releases of chemical products, by-products, and environmental pollutants, whatever their sources, has become the ambition of most societies, and many now claim to rely upon risk assessments to guide their decisions about the necessity for, and extent of, such controls. Indeed, the risk assessment framework sketched out in the first edition, and given far more extensive treatment in this one, has come to be seen as a most powerful tool for evaluating and putting into useful form the complex, diverse, often inconsistent, and always incomplete scientific information and knowledge we have been able to accumulate about the health hazards and risks all chemicals pose if exposures become excessive. So in this edition the reader will be provided with a broader and deeper look at risk assessment as it continues to evolve as a scientific enterprise, and in its role as the bridge between basic and applied research and the many forms of decision-making aimed at risk reduction. Chapters dealing with the early evolution of the risk assessment framework and the several principles and concepts that gave rise to its structure, the ways in which relevant scientific data and knowledge are put to use within that framework, and some of the new challenges risk assessment faces, are almost wholly new in this edition.

It is perhaps too obvious to point out that everything we can see, touch, smell, and taste is a chemical or, more likely, a mixture of many different chemicals. In addition, there are many chemical substances in the environment that cannot be detected with the senses, but only indirectly, by the sophisticated instruments scientists have devised to look for them. The number of different chemicals in and on the earth is unknown, but is surely in the many millions. During the past 125 years scientists have been successful in creating hundreds of thousands of compounds that do not occur in nature, and they continue to add to the earth's chemical stores, although most of these synthesized chemicals never leave the research scientists' laboratories.

For both historical and scientific reasons chemists divide up the universe of chemicals into inorganic compounds and organic compounds. The original basis for classifying chemicals as “organic” was the hypothesis, known since the mid nineteenth century to be false, that organic chemicals could be produced only by living organisms. Modern scientists classify chemicals as “organic” if they contain the element carbon. Carbon has the remarkable and almost unique property that its atoms can combine with each other in many different ways, and, together with a few other elements – hydrogen, oxygen, nitrogen, sulfur, chlorine, bromine, fluorine, and a few more – can create a huge number of different molecular arrangements. Each such arrangement creates a unique chemical.

I believe it was Mark Twain who quipped, when asked what he thought of the music of Richard Wagner, that “It's not as bad as it sounds.” Risk assessment might be similarly described.

Some risk assessors, in apparent disregard for the uncertainties associated with interspecies extrapolations and identifying dose–response relationships, have in recent years been making one announcement after another on the health risks associated with chemicals in the environment. “There is a one-in-one hundred thousand chance that ALAR will cause cancer in children consuming apple juice.” “Methylmercury contamination of seafoods leads to 80 000 cases each year of learning disabilities in children.” “Dioxin contamination of animal-based foods causes cancer in one of every 500 000 people.” The media offer such revelations with increasing frequency, usually accompanied by statements from regulatory agencies designed to quell public fears, remarks from manufacturers to the effect that risks have been greatly exaggerated, and professions of outrage from critics of both the regulatory and industrial communities.

It should be obvious by now that statements about risk of the type cited above, if standing alone, are misleading; if it is not obvious, it will be by the end of this chapter. There are no means available to identify these types of risk with the degree of certainty suggested by the language used. Perhaps the best a risk assessor might do, given today's knowledge, is a summary that goes something like this:

INTRODUCTION

With the benefit of hindsight, the last 20 years in microbial ecology will probably be referred to as the census period that dramatically changed our perception of biodiversity within the three domains of life. Bacteria and archaea are no longer viewed as groups of peculiar and morphologically simple organisms that show relatively little diversification despite their long evolutionary history, but have now been recognized to harbour a perplexing number of novel phylogenetic lineages (Rappé & Giovannoni, 2003). Current estimates assume that the number of prokaryotic species ranges in the millions and thus vastly exceeds the fewer than 10 000 described prokaryotic species that have been isolated to date in pure culture (Curtis et al., 2002). This dramatic paradigm shift was only made possible by the development of cultivation-independent molecular approaches for surveying microbial diversity in nature. Whilst it is now evident that most prokaryotes cannot be cultured easily, due to their living in complex communities and their intimate metabolic links with both their abiotic and biotic environments, the powerful arsenal of techniques at our disposal enables us to see beyond the ‘cultured few’ and gain valuable insights into the realm of uncultured microorganisms (Wagner, 2004). It is now relatively straightforward to determine the species richness of natural microbial communities by comparative sequence analysis of environmentally retrieved 16S rRNA gene sequences (Olsen et al., 1986; Schloss & Handelsman, 2004).

INTRODUCTION

Both basic and applied science issues drive our interests in the microbiology of the deep terrestrial subsurface. As an environment that is disconnected from the Earth's surface, the deep subsurface is less subject to variations in temperature and light and, in unsaturated zones, to intense gradients across interfaces created at the microscale level. These characteristics dictate an average growth rate that is very slow, up to thousands of years per cell division (Kieft & Brockman, 2001), and an ecosystem where change occurs over very long time scales (Fredrickson & Onstott, 2001). Thus, the subsurface is one of the most extreme environments on Earth, and identifying what limits life in the subsurface has value as a model for life on other planets (Chapelle et al., 2002; Nealson & Cox, 2002). The inadvertent release of contaminants from industrial processing plants and storage tanks, as well as the possibility of permanently depositing nuclear wastes deep below the Earth's surface (Pedersen, 2001), raise questions about how microbial activities might exacerbate or mitigate contamination problems in the subsurface.

The terrestrial subsurface is the habitat for diverse microbial communities that, together with the oceanic subsurface, may be the habitat for the largest proportion of Earth's biomass (Whitman et al., 1998). As subsurfaces are characterized by a range of physical and chemical properties, from fully aerated sedimentary shallow aquifers to deep igneous rocks devoid of oxygen and elevated in temperatures, their microbial communities are equally varied (Fredrickson & Fletcher, 2001).

INTRODUCTION

Methane is an important link within the global carbon cycle and has become a major focus for scientific investigations over the last decades, especially since the discovery of large deposits of methane hydrates in continental margins. The majority of recent methane production is biogenic, i.e. produced either by thermogenic transformation of organic material or by methanogenesis as the final step in fermentation of organic matter carried out by methanogenic archaea in anoxic habitats (Reeburgh, 1996). There are also abiotic sources of methane, e.g. at mid-oceanic ridges, where serpentinization takes place. In marine environments, the bulk of the methane is produced in shelf and upper continental-margin sediments, which receive large amounts of organic matter from deposition (Reeburgh, 1996). As methane builds up, it migrates upwards and may reach the sediment surface. Here, its ebullition and oxidation can lead to the formation of complex geostructures, such as pockmarks or carbonate chimneys and platforms, as well as large-scale topographies, such as mud volcanoes and carbonate mounds (Ivanov et al., 1991; Milkov, 2000). In most of the deeper continental margin and the abyssal plain sediments, methane production is low, as only 1-5 % of the surface primary production reaches the bathyal and abyssal seabed, due to degradation processes in the water column (Gage & Tyler, 1996).

INTRODUCTION

This chapter will consider biogeochemical cycling in the coastal zone. This is defined as that area of estuarine and coastal, relatively shallow water where there is strong benthic-pelagic linkage and exchange between the water column and the underlying sediment. In deeper water this connection becomes increasingly tenuous as the exchange between the euphotic zone and the benthic layer declines. Longhurst et al. (1995) recognized the coastal boundary domain, divided into 22 provinces, as often bounded by a shelf-break front, and included coastal upwelling regions. The coastal zone generally exhibits high rates of primary production compared with the open ocean (Table 1), and there is the greatest impact from inputs from the land to the coastal sea through estuaries. Estuaries and coastal seas are highly heterotrophic systems which are net exporters of CO2 to the atmosphere due to the mineralization and recycling of both autochthonous and allochthonous organic matter (Borges, 2005).

PHYSICO-CHEMICAL DIFFERENCES BETWEEN LATITUDINAL REGIONS

The physical-biological interactions that influence marine phytoplankton production have been reviewed by Daly & Smith (1993). Because of the spherical shape of the Earth, more solar energy falls per unit area of surface in equatorial regions than at the poles (Fig. 1a), and the incidence of light at the equator is vertical to the surface, but oblique at the poles. Furthermore, the distance radiation travels through the atmosphere is longer at the poles, thus reducing the irradiation incident at the poles compared with equatorial regions.

INTRODUCTION

Although it has been known for over a century that micro-organisms have the potential to reduce metals, more recent observations showing that a diversity of specialist bacteria and archaea can use such activities to conserve energy for growth under anaerobic conditions have opened up new and fascinating areas of research with potentially exciting practical applications (Lloyd, 2003). Micro-organisms have also evolved metal-resistance processes that often incorporate changes in the oxidation state of toxic metals. Several such resistance mechanisms, which do not support anaerobic growth, have been studied in detail by using the tools of molecular biology. Three obvious examples include resistance to Hg(II), As(V) and Ag(II) (Bruins et al., 2000). The molecular bases of respiratory metal-reduction processes have not, however, been studied in such fine detail, although rapid advances are expected in this area with the imminent availability of complete genome sequences for key metal-reducing bacteria, in combination with genomic, proteomic and metabolomic tools. This research is being driven forward both by the need to understand the fundamental basis of a range of biogeochemical cycles, and also by the possibility of harnessing such activities for a range of biotechnological applications. These include the bioremediation of metal-contaminated land and water (Lloyd & Lovley, 2001), the oxidation of xenobiotics under anaerobic conditions (Lovley & Anderson, 2000), metal recovery in combination with the formation of novel biocatalysts (Yong et al., 2002a) and even the generation of electricity from sediments (Bond et al., 2002).

INTRODUCTION

Whilst chemolithoautotrophic micro-organisms are found in nearly every environment on Earth, they are more abundant in dark habitats where competition by photosynthetic organisms is eliminated. Caves, particularly, represent dark but accessible subsurface habitats where the importance of microbial chemolithoautotrophy to biogeochemical and geological processes can be examined directly. At Lower Kane Cave, WY, USA, hydrogen sulfide-rich springs provide a rich energy source for chemolithoautotrophic micro-organisms, supporting a surprisingly complex consortium of micro-organisms, dominated by sulfur-oxidizing bacteria. Several evolutionary lineages within the class ‘Epsilonproteobacteria’ dominate the biovolume of subaqueous microbial mats, and these microbes support the cave ecosystem through chemolithoautotrophic carbon fixation. The anaerobic interior of the cave microbial mats is a habitat for anaerobic metabolic guilds, dominated by sulfate-reducing and -fermenting bacteria. Biological controls of speleogenesis had not been considered previously and it was found that cycling of carbon and sulfur through the different microbial groups directly affects sulfuric acid speleogenesis and accelerates limestone dissolution. This new recognition of the contribution of microbial processes to geological processes provides a better understanding of the causal factors for porosity development in sulfidic groundwater systems.

Karst landscapes form where soluble carbonate rocks dissolve by chemical solution (karstification), resulting in numerous geomorphic features, including caves and subterranean-conduit drainage systems (e.g. White, 1988; Ford & Williams, 1989). This has traditionally been viewed as an abiotic, chemical process that occurs near the water table, with biologically produced CO2 as the principal reactive component.

INTRODUCTION

This chapter is intended to provide a brief overview of the key concepts underlying the emerging area of environmental microbial metal geochemistry, rather than an exhaustive synthesis. The reader is referred to the following more comprehensive reviews on the biogeochemistry of metals (Warren & Haack, 2001), metal-mineral reactions (Brown & Parks, 2001), emerging molecular-level geochemical techniques (O'Day, 1999; Brown & Sturchio, 2002) and a recent synthesis of how genetic expression in the environment can underpin geochemical reactions (Croal et al., 2004). The relevance of micro-organisms to metal behaviour arises from the overlap of the biosphere with the geosphere and the transformations that occur because of their interactions. Microorganisms have evolved in intimate association with the rocks, soils and waters (i.e. geosphere) in which they find themselves. In order to grow and survive, they have adapted to these environments and use the inorganic components to drive their metabolic machinery; the myriad functional pathways by which they do so ensure that they influence a number of key elemental cycles in the process. As a consequence, many important geochemical processes are ultimately shaped by life, rather than strict geochemical equilibria, a fact that is increasingly recognized as strict geochemical principles fail to constrain observed environmental behaviour.

Trace-metal behaviour in the environment is of increasing global concern as water and soil contamination with these toxic substances continues and the detrimental effects on ecosystems and human health emerge.

INTRODUCTION

Primary productivity in the ocean amounts to the net assimilation of CO2 equivalent to about 50 Pg (petagram, i.e. 1015 g) C year–1, while on land this is approximately 60 Pg C year-1 (Field et al., 1998). Almost all of this primary productivity involves photosynthesis, and in the ocean it occurs only in the top few hundred metres, even in waters with the smallest light attenuation (Falkowski & Raven, 1997). About 1 Pg C of marine primary productivity involves benthic organisms, i.e. those growing on the substratum (Field et al., 1998), in the very small fraction of the ocean which is close enough to the surface to permit adequate photosynthetically active radiation (PAR) to allow photolithotrophic growth. This depth at which photosynthetic growth is just possible varies in time and space, and defines the bottom of the euphotic zone (Falkowski & Raven, 1997). The remaining ∼49 Pg C is assimilated by phytoplankton in the water column (Field et al., 1998). This chapter will concentrate on the planktonic realm, while acknowledging the importance of marine benthic primary producers and their interactions with micro-organisms (e.g. Dudley et al., 2001; Raven et al., 2002; Raven & Taylor, 2003; Cooke et al., 2004; Walker et al., 2004).

The global net primary productivity of the oceans is less than that on land, despite about 70 % of the Earth being covered in ocean and primary productivity over considerable areas of land being limited by water supply.

INTRODUCTION

A fungal component of the marine biota was only recognized as recently as 1944 (Barghoorn & Linder, 1944), and it was not until the 1960s that studies commenced to assess the extent and diversity of fungi in marine systems. Since this time, considerable effort has been exerted to uncover marine fungal diversity, with high decadal discovery indices in the 1970s and 80s (Hawksworth, 1991), resulting in around 1000 fungal species known today from marine environments. Nevertheless, it is hardly surprising that, with the extent of marine environments globally, we probably have a very incomplete view of fungal diversity, together with their frequency and function in these ecosystems. The objective of this review is to assess the extent of our present knowledge and to highlight future directions to further elucidate their biology and ecology.

THE NATURE OF MARINE ENVIRONMENTS

Marine ecosystems are globally extensive, and account for around 70 % of global surface area. They can be defined generally as aquatic systems influenced by substantial concentrations of salts, particularly sodium chloride, from existing oceanic systems. Seas and oceans divide between regions bordering and influenced by terrestrial regions and the open ocean, which is strongly zoned through the water column. These broad boundaries are illustrated in Fig. 1, which also details linkages between marine compartments.

INTRODUCTION

The most important perceived environmental roles of fungi are as decomposer organisms, plant pathogens and symbionts (mycorrhizas, lichens), and in the maintenance of soil structure through their filamentous growth habit and production of exopolymers. However, a broader appreciation of fungi as agents of biogeochemical change is lacking and, apart from obvious connections with the carbon cycle, they are frequently neglected within broader microbiological and geochemical research contexts. While the profound geochemical activities of bacteria and archaea receive considerable attention, especially in relation to carbon-limited and/or anaerobic environments (see elsewhere in this volume), in aerobic environments fungi are of great importance, especially when considering rock surfaces, soil and the plant root-soil interface (Gadd, 2005a). For example, mycorrhizal fungi are associated with ∼ 80 % of plant species and are involved in major mineral transformations and redistributions of inorganic nutrients, e.g. essential metals and phosphate, as well as carbon flow. Free-living fungi have major roles in the decomposition of plant and other organic materials, including xenobiotics, as well as mineral solubilization (Gadd, 2004). Lichens (a symbiosis between an alga or cyanobacterium and a fungus) are one of the commonest members of the microbial consortia inhabiting exposed subaerial rock substrates, and play fundamental roles in early stages of rock colonization and mineral soil formation.

INTRODUCTION

The contribution of micro-organisms to amorphous silica precipitation in modern geothermal hot-spring environments has been the topic of intense study in the last three to four decades. Here, we present a review on the field and laboratory studies that have specifically addressed bacterial silicification, with a special focus on cyanobacterial silicification. Studies related to the biogenic silicification processes in diatoms, radiolarians and sponges are not discussed, despite the fact that, in the modern oceans (which are undersaturated with respect to silica), the diagenetic ‘ripening’ of such biogenic silica controls the global silica cycle (Dixit et al., 2001). It is well-known that the amorphous silica in these organisms (particularly in size, shape and orientation) is controlled primarily by the templating functions of glycoproteins and polypeptides (e.g. silaffin and silicatein). For information on these issues, we refer the reader to the extensive reviews by Simpson & Volcani (1981), Kröger et al. (1997, 2000), Baeuerlein (2000), Perry & Keeling-Tucker (2000), Hildebrand & Wetherbee (2003) and Perry (2003). In addition, in terrestrial environments, a large pool of amorphous silica is cycled through higher plants (grasses and trees) that are believed to use silicification as a protection mechanism against pathogens and insects. Information on these processes can be found in the papers by Chen & Lewin (1969), Sangster & Hodson (1986) and Perry & Fraser (1991).