Chapter 4

GEOLOGICAL AND GEOCHEMICAL

INFLUENCES ON ESTUARINE ECOSYSTEMS

W. LANGSTON', & J. RIDGWAY^

'Marine Biological Association, Plymouth, UK, ^ British Geological Survey, Nottingham, UK

1. INTRODUCTION
The physical structure of an estuary is governed by geological
circumstance and shaped by a combination of river flows, tidal
characteristics, current speeds and wave action. An over-riding constraint on
estuarine biota is the nature of the variable salinity regime, since the capacity
for ionic and osmotic regulation varies greatly between species and sets the
limits for their distribution. Of equal importance, if the organism is to settle
and survive, are the properties of deposits. Superimposed on these primary
drivers are numerous other factors that influence estuarine biota, either
directly or indirectly. These include light attenuation and oxygenation
patterns (natural characteristics), together with an assortment of
anthropogenic impacts. The current chapter focuses on the ways in which
geological and geochemical features (substrate properties) impinge on
estuarine ecosystems, including modifications made as a result of
contaminant bioavailability and toxicity. We also consider ways in which
biological activity can mobility in estuaries through processes such as
bioturbation and biodeposition.
In simple terms an estuary is a semi-enclosed coastal body of water, with
free connection to the sea, in which salt water is diluted by terrestrial fresh
water flow. Most estuaries are the product of the inundation of river valleys
during Holocene sea-level rise following the end of the last major glaciation.
22 Geological Environment

At first sight estuaries would appear to present biota with severe tests for
survival due to their dynamic tidal nature and accompanying changes in
salinity. Nevertheless, estuaries are among the most productive of aquatic
habitats. The biotas, which make up present day estuarine assemblages, have
their origins in three different systems: the sea, fresh water and the land.
Species of marine origin tend to dominate, however within each of these
categories there are sub-components that reflect varying degrees of success
in penetrating estuaries. As a result of this abundance and diversity, estuaries
are of major importance to fisheries, acting as nurseries to many forms of
aquatic life. Intertidal areas in particular are also significant feeding grounds
that attract a variety of bird life, often being globally important sites for
migratory species.
Positioned between marine and terrestrial environments, estuaries are
zones of sediment transfer between fluvial and marine systems and often
form sinks for sediment from both their hinterland and adjacent coastal zone.
Such sediment may vary greatly in grain size, mineralogy and chemistry, and
there is also the possibility that parts of the estuary will be dominated by
bedrock. These factors, together with salinity, tidal and turbidity regimes, all
may play a part in determining the type of organism and hence community,
capable of colonizing different parts of the estuary. In addition, estuaries
have been the focal point for a variety of human activities, becoming sites of
major port, industrial, urban and recreational development. Agriculture is
often a common feature of the coastal lowlands along their shores. Estuaries
may thus be affected by dissolved and particulate contaminants from
recreational, farming, manufacturing and extractive industries, both on land
and offshore, together with domestic inputs from sewage. Estuarine biota,
therefore, have to adapt to a unique combination of natural and
anthropogenic forcing features.
This examination of geological and geochemical influences is by no
means exhaustive and is largely centered on a few well-studied examples
from temperate estuaries and ecosystems, although mention will be made of
some tropical and sub-tropical estuaries. The emphasis is placed on the
benthic component of estuarine ecosystems since it is here where geological
and geochemical influences are likely to be greatest.

2. GEOLOGICAL INFLUENCES
The most obvious and direct geological influence on estuarine
ecosystems is the nature and extent of various types of substrate. This pattern
is largely pre-determined by the unique geological history of the area.
Clearly, a rock substrate provides a firm base for the attachment of
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 23

epibenthos, which contrasts significantly with the infaunal assemblages

typically associated with mud and sand substrates. This diversity in form and
function will, in turn, affect the biota that depends on bottom-living
organisms for food or protection. A useful account of the relationships of
estuarine organisms to varying substrates is given in Perkins (1974).
By virtue of their genesis through the drowning of pre-existing river
valleys, estuaries inherit features from past landscapes, which in turn have
been strongly influenced by the interplay between geology and climate.
Thus, estuaries may reflect the inundation of relatively narrow, steep-sided
valleys, as in typical arias, or of wide floodplains, leading to the classic,
funnel-shaped development (for a classification of estuaries see Dyer, 1997).
Some of the complexities of estuary formation under different geological
and climatic conditions are described by Lampe (1996) from the shallow,
polymict, tideless estuaries (boddens) of the Baltic coast of Germany,
Woodruffe (1996) from macrotidal estuaries in northern Australia, and
Healy et al. (1996) from a wide diversity of estuary types in New Zealand.
Estuarine environments range from sub-tidal, through intertidal, landward to
the limits of inundation by storms, and include beaches, dunes, eroding
margins, stream and tidal channels, rocky platforms, saltmarshes,
mangroves, sea grass meadows, algal beds and tidal mud or sand flats
(Roman and Nordstrom, 1996). Substrates may be dominantly rocky, sandy
or muddy, with any combination of the three possible.
The nature of the substrate at any particular site is governed by the
complex interaction between tidal forces, freshwater flow and sediment
supply, the latter being particularly dependent on the geology of both the
hinterland and offshore areas. Shoreline exposure to waves is also of great
importance in determining the type of environment that will be established
and maintained at a specific location. For estuaries, fine-grained material is
terrestrially derived, whilst coarser material is of marine origin, transported
landward by tidally driven, bottom currents (Ridgway and Shimmield,
2002). However, in some systems fine-grained muddy material comes from
both terrestrial and marine sources, as in the case of the Scarcies Estuary of
Sierra Leone, West Africa (Anthony, 1996) and the Humber Estuary in the
UK (Dyer et al., 2001). The distribution of mud and sand in an estuary will
clearly be influenced by such factors.
Relationships between sediment type and supply and the hydrodynamics
of the estuary affect the degree of turbidity and this in turn has a strong
influence on the ecosystem, governing the amount of light penetrating
through the water column to the bed. In highly turbid estuaries
photosynthetic activity of phytoplankton can only take place in the surface
(photic) layers and as a consequence productivity may be reduced to a
24 Geological Environment

minimum. In these conditions salt marsh-plants will dominate overall carbon

production of the estuary.
Where tidal currents are particularly strong, sediment may be unable to
accumulate and the bed of the estuary will be of rock or other well-indurated
material, such as glacial till. The nature of the geology, from hard igneous
and metamorphic rocks, through sandstones to softer limestones, chalks and
very soft clays affects the surface texture on which organisms can live (e.g.
very smooth, highly pitted) and also the larger physical features of the
environment, from even rock platforms to highly fissured and folded terrain.
Rocky substrates provide stable surfaces for the attachment of sessile and
sedentary organisms, whereas the shifting, unstable environments of
estuarine sediments are host to only infauna as permanent residents (Perkins,
1974). In turn, the variable properties of these soft bottom estuarine deposits
(in particular grain size) may be of considerable importance in shaping
estuarine communities.
The Joint Nature Conservation Committee of the UK (JNCC) has
conducted a survey of Britain's marine habitats - the Marine Nature
Conservation Review (MNCR) - and some examples of benthic habitats
from this review (Hiscock, 1996) serve to illustrate the importance of the
geological influence on ecosystems. Four major habitats (with numerous
sub-divisions), all of which can occur in estuaries, are recognised in the
MNCR: littoral rock; sublittoral rock; littoral sediment; and sublittoral
sediment.
Littoral rock environments provide a range of structures (rock platforms,
cliffs, overhangs, caves, pools, boulder fields etc.) that encourage species
diversity. Zonation of species is common, due to tidal immersion and
emersion and biological interactions such as competition for space, predation
and grazing. Stable rocky substrates support a wider range of species than
unstable hard substrates of boulders, cobbles and pebbles. However, stable
bedrock-type communities can develop on shingle and cobbles in areas of
estuaries and sea lochs sheltered from wave action. Rock type is important:
rich algal communities can develop if the rock retains water during tidal
emersion and soft rocks allow animals to bore into them to provide security
from predators; on harder rocks, crevices allow distinctive faunas to develop.
In low or variable salinity zones of estuaries, rocky substrates are
characterized by a low number of species that also occur in full salinity
conditions. Sublittoral rock can support a richer diversity of species than
littoral rock because it is always submerged, placing less stress on
organisms, but the turbid conditions of many estuaries may limit
biodiversity.
In littoral sediment environments, particle size, the mixture of sediment
grades and the stability of the sediment are important controlling factors in
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 25

determining the types and numbers of species present. For example, fine
sediments support more diatoms and bacteria than coarse sediments, because
of the larger surface area for attachment and growth. In contrast, ciliate
protozoa are generally unable to live in sediments of < 0.1 mm diameter.
Moreover, those protozoa species that inhabit sands with a mean particle size
between 0.4 and 1.0 mm are generally rather square in form, whereas those
living in sand of 0.1 to 0.4 mm tend to be long and slender in form (Perkins,
1974). Particle size can also affect burrowing and feeding habits and is of
considerable importance to filter-feeding organisms. Many suspension-
feeders have an optimum range of particle sizes and abundance (in
suspension) that can be retained by gills and other feeding structures. The
predominance of a particle population within the favoured range, at any
given site, will tend to promote colonisation and growth. Depending on the
organism, preferred particle sizes vary over a spectrum from <l|uim to
100|im - equivalent to the material carried in suspension in most estuaries.
Surface chemistry of particles also has an important role in the selection of
appropriate material for digestion. By comparison, deposit-feeders may be
less selective in the choice of particles.
Capilliary lift and water retention are also a function of grain size and are
generally greatest in fine sediments as opposed to coarse sands. This is of
obvious importance to inter-tidal organisms, where water retention during
low tide may be critical for survival. The phenomenon accounts, partly, for
the zonation of species seen in estuaries, though, perhaps with the exception
of saltmarsh plants, this is generally less striking than the more familiar
zonation patterns seen on rocky coastlines. On the upper parts of a mud flat
where there is a risk of drying out, some species have developed restistance
to dessication, in the case of certain diatoms through the ability to secrete
mucilagenous envelopes. Sediment porosity and permeability are related
features of importance in littoral estuarine deposits, since they govern
resistance to the burrowing activities of biota. Sediments with a high water
content generally offer less resistance to penetration by organisms, however
if they become too fluid the permanent burrow structures preferred by some
species cannot be maintained and settlement is impossible. Colonisation by
saltmarsh plant species can also be determined by the firmness of the
substrate: Spartina (cord-grass) may take root in a variety of sediment types
though Salicornia (samphire) is uilable to gain a hold in very fine muds.
Sublittoral sediment environments provide extensive sedimentary
habitats and associated communities. The most species-rich sediments are
those that are stable over time and have a heterogeneous mixture of coarse
and fine sediment grades. Sediment composition varies according to the
strength of wave action, tidal streams and sediment supply. Strong tidal
streams generally lead to a coarse sediment bed, but in the Severn Estuary
26 Geological Environment

(UK) the suspended sediment load is so high that silt settles out and the bed
may be muddy over large areas, despite the presence of very strong tidal
currents (see below and Hiscock, 1996). Deposition in estuaries is
sometimes interspersed with episodes of erosion (associated with high flow,
extreme tidal conditions or variable river channels): clearly, these periodic
alternations in depositional patterns will tend to wash out infauna and
prevent settlement of established communities, especially those with life-
spans greater then a few months.
An example of the distribution of biotopes in the Swale estuary in
southeast England is shown in Figure 4-1. Other examples, from Germany
and New Zealand, respectively, are given in Lampe (1996) and Healy et al.
(1996).

2.1 Geological Influences: Case Study - The Severn

Estuary (UK)
One of the best examples illustrating the importance of natural physical
and geological forcing agents on estuarine communities is that of the Severn
Estuary (Langston et al, 2003a). The exceptional tidal range (in excess of
14.5 metres) and classic funnel shape (Figure 4-2) make the Severn Estuary
unique in Britain and rare worldwide. Large tidal-currents are a dominating
feature, reducing vertical stratification (compared with the rias typical of
south west England) and providing a mechanism for transport of particles up
to sand-size (moving as suspended solids or mobile bed-load). There is a
continual exchange of material from areas of erosion to areas of deposition,
through the turbidity maximum - which occupies the whole of the Estuary
east of Bridgwater Bay. The associated variable frictional stresses result in
variations in bed-types despite the ever-present turbidity of the water. In
turn, the relatively sharp divisions between muddy, sandy and rocky areas
dominate the distribution of benthic organisms. Composition will obviously
differ, fundamentally, between hard- and soft-bottom communities, whilst
among the latter, sediment grain-size is a dominant characteristic
determining the abundance and type of estuarine infauna (Warwick et al,
1989; Moore e^ a/., 1998).
A high proportion of the estuary is subtidal, hosting some rare
communities. The intertidal flats and rock platforms support a range of
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 27

Figure 4-1. The distribution of biotopes in the Swale, Kent, Southeast England (reproduced
from, Hiscock, 1996, with permission JNCC).

invertebrate species and the upper Severn Estuary includes an extensive area
of mudflats and sandflats bordered by large fringes of saltmarsh. This variety
of habitats has, in fact, led to its proposed designation as a Special Area of
Conservation (SAC) under the EU Habitats directive. Despite this label,
however, the Severn as a whole supports a relatively impoverished fauna and
flora, characterized by low biodiversity when compared to other sites
(Warwick et al., 1989; Moore et aL, 1998).
28 Geological Environment

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IMS";

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1 i 1 r; It I irrj; jl f iXd '. LA'i S)

GEOLOGICAL AND GEOCHEMICAL INFLUENCES 29

I FtNaiivl^tMJinA tklllr•n«H1l^v^"l;.•l'w-tV<:^

I kiKr- ( bl>tr HjJll

HfciSu, KVHI. WVi. •,t4Ti: NWr-r.V. W li\W .

F/gwre 4-2. The distribution of biotopes in the Severn Estuary, Southwest England. Upper
Severn (top), Lower Severn (bottom) (reproduced from Moore et al, 1998, with permission
JNCC).
Legend to the top figure: Estuarine littoral rock with fucoids, Estuarine littoral rock with
ephemeral algae, Littoral mixed substrata with fucoids. Littoral muddy sand. Littoral sandy
mud and soft mud. Littoral soft mud, Estuarine sublittoral mud.
Legend to the bottom figure: Moderately exposed littoral rock, Estuarine littoral rock with
fucoids. Littoral sand and gravel. Littoral muddy sand. Littoral sandy mud and soft mud,
Littoral/sublittoral mobile sand, Littoral/sublittoral mud. Tide-swept sublittoral hard substrata,
Infralittoral sand, Infralittoral gravel, Infralittoral muddy sand, Infralittoral mud, Estuarine
sublittoral mixed sediment.

High turbidity is largely responsible for this characteristic, impacting on

biota in a number of ways. Generally, the high-suspended solids loading
limits light penetration and hence algal productivity. Thus, a striking feature
30 Geological Environment

of much of the estuary is the absence of a subtidal zone of macroalgae, due

to the effect of the high turbidity, which reduces available photosynthetic
light, coupled with the scouring effect of the silt, which interferes with the
settlement of algal spores. Similar effects may impact on eelgrass (Zosterd)
beds in mid-estuary. Here, additional deposition of sediment during the
construction of the nearby second Severn crossing has coincided with a
reduction in area and density of this important and sensitive biotope.
Invertebrate populations associated with the Zostera bed may also have been
affected by extreme episodes of erosion and deposition of sediment (allied to
coffer dam construction).
Whilst limiting vegetation, high turbidity provides abundant particulate
surface area for microbial processes. As a result, organic carbon may be
enriched (and biological oxidation demand elevated) in fluid muds that can
disperse at spring tides to produce dissolved oxygen sags in the upper
estuary. In those areas frequently covered by turbid layers, colonisation is
likely to be sparse. Much of the sub-tidal Severn mud is impoverished and
even some sandy areas may be depleted because of the mobility of silts at
spring tides. Extreme conditions (sediment instability, turbidity and
scouring) also limit the range and abundance of 'expected' estuarine
organisms to be found inter-tidally (see typical food-web associated with
estuarine mud-flats in Figure 4-3, re-drawn from Green, 1968).
Consequently, productive areas are restricted to the more stable, marginal
regions. Throughout much of the Severn Estuary, the virtual absence of
suspension-feeding bivalves including Cerastoderma edule and Mya
arenaria (and other suspension-feeding invertebrates), can be attributed to
the very high levels of turbidity. The deposit-feeding bivalve Scrobicularia
plana is also uncommon, though probably because of sediment instability.
For similar reasons the colonisation of rooted saltmarsh plants is limited to
areas which are least affected by strong tidal flows.
Sedimentary conditions not only affect ability to colonize, but can also
influence the longevity of those estuarine organisms that do settle. In
comparison with other estuaries, populations of several invertebrate species
in the Severn are dominated by small individuals, suggesting a shorter
lifespan (Warwick e/a/., 1989).
Suspended sediment concentrations in the Severn vary throughout the
diurnal and spring/neap tidal cycles, and modify the distribution of benthos
accordingly. Sediments in the upper reaches tend to reflect deposition of fine
sediment during neaps and erosion of fines (together with deposition of
coarser sediment) during springs. This pattern is in turn exacerbated by
seasonal influences such as conditions of river flow and storm surges,
leading to both short-and long-term cyclical influences on biota.
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 31

STRAND LINE
INSECTS

SMALL
POLYCHAETAI

Figure 4-4. Generalised example of a food web from the mudflats of a UK estuary (adapted
from Green, 1968)

The fate of the large quantity of sediment that is resuspended and

recirculated on each tide will also play a major role in the mobility,
bioavailability and impact of associated contaminants. Because of the
energetic hydrodynamic regime in the Severn, and resultant high turbidity,
there is considerable mixing and redistribution of fines, resulting in
unusually homogenous distributions of contaminant loadings. Since physical
conditions dominate the composition of biota in the Severn, and
communities are already modified as a result, it may be more difficult to
gauge impact due to contaminants than in typical estuarine systems, except,
perhaps, in the immediate vicinity of major discharges.

3. GEOCHEMICAL INFLUENCES
Geochemically, one of the most important elements in the biology of
estuaries is sulphur, which is involved in a complex interplay with other
elements including oxygen. If oxygenation is low, sulphate in estuarine
water is reduced to sulphide. Reduced circulation in pore waters, low O2
levels and the presence of Fe and decaying organic matter will lead to the
formation of ferrous sulphide. The depth and characteristics of this black
reducing layer in sediments influences macrofaunal communities and the
32 Geological Environment

activities of a range of microorganisms. If Fe salts are limiting, excess

hydrogen sulphide liberated by sulphate-reducing bacteria permeates into
pore water and overlying water, compounding the problems for sensitive
species caused by low O2 levels. Generally, sulphur bacteria can tolerate a
wider range of redox conditions and some are capable of oxidising H2S, to
produce free S, which in turn may be metabolised in various ways by other
microorganisms, to complete the sulphur cycle.
Rather than dwelling on this well-reported sequence, the emphasis in this
section is on how geochemical properties, including salinity, can influence
biota in other ways - in particular, the relationships between organisms and
contaminants in sediments (using metals as the primary example). This
encompasses consideration of chemical speciation and bioavailability,
together with an appraisal of assimilation pathways and finally ecological
effects. Firstly, however, it is important to recognise that it is not just the
activities of bacteria and other microorganisms that can shape sediment
characteristics, other biological processes, including bioturbation and
biodeposition, are significant in terms of sediment geochemistry and
transport.

3.1 Bioturbation and biodeposition

Estuarine sediments can act as a store for buried contaminants, which
may be isolated from organisms until remobilized by storms, dredging or
changes in currents brought about by natural events, or construction works
such as sea defenses or port installations. However organisms may
themselves modify sediment properties and play a role in remobilising
contaminants from sediment to overlying waters (and vice versa).
Bioturbation and its consequences can vary considerably in nature as a
result of the temporal and spatial heterogeneity in abundance and
composition of benthic communities, coupled with the extent and type of
sediment disturbance. The review by Lee and Swartz (1980) categorises
bioturbatory processes according to guilds (groups of organisms that
function similarly in terms of sediment reworking) of which twelve
dominant types are recognised. This scheme is primarily based on feeding
type (suspension or deposit feeders; injection and ejection of sediments at
differing depths), different degrees of mobility (sedentary to highly mobile)
and burrowing activity (complexity and extent, horizontally and vertically;
epifaunal or infaunal). Within each of these guilds particle reworking can be
further categorised, depending on the relative importance of the different
activities (burrowing, excavation, irrigation, feeding, tube production,
locomotion at the surface and effects on topography). To date, evidence of
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 33

the relative importance of these activities, particularly on contaminant

movements, has largely been qualitative.
Individual reworking rates vary tremendously between species - values
range from ~1 g d"^ up to >200g d"^ depending on category of bioturbation,
organism size and seasonal parameters such as temperature. In terms of total
reworking rate per unit area of sediment (dependent also on abundance of
biota), values are equally variable. They extend over several orders of
magnitude - from a few hundred grams m"^ y"^ for small polychaetes and
bivalves, up to several hundred kg m'^ y"\ and occasionally more, in large
callianassids and holothurians. Care must be taken, however, in
extrapolating to impacts on pollutant budgets - high rates of sediment
reworking may not necessarily result in proportionate effects on sediment or
contaminant transport; more quantitative data is needed on this aspect.
To add to this complexity, bioturbation can affect not only transport of
particulate material, but also the geochemical composition of the sediment
and pore waters - including adsorption/ desorption properties at the
sediment-water interface. In attempting to quantify bioturbation the most
common solution to date has been to assume that it can be described as a
simple diffusive process. However, although there have been some attempts
at modelling sediment redistribution, the signs are that, in reality, the
calculated biodiffusion coefficient does not correlate well with faunal
density or composition since it represents the summation of a complex
mixture of different processes. There is still a need to relate animal activities
with their mixing consequences.
Rhoads (1974) has indicated that biological transport of sediments in
benthic ecosystems is likely to outweigh molecular diffusive fluxes for
some, but not all, ions. Therefore, where sediments are relatively stable and
well populated, it would seem that biological processes could dominate the
distribution of sediment contaminants. In an early demonstration of this,
measured diffusion coefficients derived from mixing curves for ^^^Ru in Irish
Sea sediments were found to be four-to-five orders of magnitude higher than
predicted molecular diffusion coefficients and the disparity was attributed to
biogenic influences (Duursma and Gross, 1971).
Sediment redistribution of pollutants by benthic biota may vary
considerably, depending on conditions. For example, where biological
reworking rates are low (or deposition rates are high), episodic pollutant
inputs will be preserved after a time as a discrete layer buried below the
mixing zone. Cores with this type of structure are most useful in tracing
historical trends in contamination. Where biological reworking is high and
sedimentation relatively low, an equivalent pollution episode will rapidly
become dispersed throughout the reworked layer, though its eventual
disappearance below the mixing zone may be prolonged due to continual
34 Geological Environment

reworking at depth. Bioturbation depths vary considerably among different

types. Probably most activity takes place in the upper 10cm, though larger
burrowing shrimps may be active at depths of over one metre. The potential
importance of biological processes on pollutant distribution in sediments is
evident.
Studies with burrowing shrimp Callianassidae indicate considerable
variability in burrow structure and depth, and therefore bioturbating
properties, which is related partly to the nature of the substrate. In soft muds
there is a single shaft of 30-80cm -much deeper and less complex than the
chambered systems (9-23cm) constructed in coarser substrate (Rowden and
Jones, 1995). These organisms have also been used to demonstrate the
possible geochemical consequences of sediment turnover, which generally
relate to the introduction of oxidising conditions, and resultant changes in
speciation and solubility, at depth in sediments. Effects are most pronounced
for redox-sensitive elements such as Mn, Fe, Pu and As. Transport of
suboxic sediments to the surface by other 'conveyor-belt' species is also
likely to be an effective mechanism for introducing reduced metal sulphide
to the interface; subsequent microbial and chemical oxidation will influence
the speciation and mobilisation of a number of elements. Related
mechanisms by which bioturbation can effect speciation and remobilisation
include the pumping of large volumes of water through tubes and burrows to
the surface (irrigation), and the movement and mixing actions of biota in
surface layers: the burrowing activities oi Nereis diversicolor, for example,
increase fluxes of Zn by between three and sevenfold (Renfro, 1973).
Burrowing organisms can affect the geotechnical properties of estuarine
sediments, which may also influence, indirectly, contaminant transport,
speciation and distribution. The most important of these effects concerns
sediment stability, which tends to decrease in the presence of abundant
macrofaunal populations. This decrease in stability results from an increase
in water content, degradation of organic binding and altered surface
microtopography and is likely to increase erosion and resuspension at the
sediment-water interface. In some cases there may be a direct relationship
between both abundance and type of bioturbation and erodibility.
Consequently, the effect on erodiblity may be seasonal: biological activity is
usually highest in summer, implying that the bed may be less resistant to
erosion, particularly during summer storms. Destabilisation, though probably
only affecting the top few cm of sediment, can result in near-bottom
turbidity zones even in relatively weak tidal currents.
Contaminants such as metals, oil and pesticides may, if present in
sufficient concentrations, influence the rate of bioturbation by affecting
processes such as filtration rates and re-burial times and, in extreme cases,
species composition. Pollutant-induced community alterations could, in turn.
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 35

influence sediment turnover rate and stability, for example where large sub-
surface burrowers are replaced by less active or sedentary opportunists such
as small polychaetes. This type of effect might be important in estuaries,
particularly close to discharges.
Whilst most bioturbatory activities tend to lead to greater remobilisation,
some biogenic processes associated with benthic ecosystems can help to
decrease erosion. For example, high organic loadings originating from
microorganisms and their exudates tend to increase cohesive properties,
whilst dense tube mats of some polychaetes, algal mats and roots of sea
grasses and salt marsh plants also help to stabilise the benthic sediments. An
illustration of the importance of this occurred during the 1930s when Zostera
was eliminated extensively around the UK by disease: the absence of
stabilising rhizome systems resulted in substantial erosion and collapse of
mud banks. Dense growths of some estuarine salt-marsh plants such as
Spartina and Salicornia may encourage deposition and consolidation by
reducing current speeds. The same applies within the tangled rhizophores
(prop roots) of mangrove vegetation (the equivalent niche to salt-marsh in
many tropical and sub-tropical regions).
High densities of suspension-feeding benthic organisms can remove
significant quantities of suspended matter from the overlying water which,
via faecal pellets or pseudofaeces, eventually becomes incorporated into bed
material. Biodeposition by benthic organisms will therefore increase the rate
at which particulate contaminants are sedimented and reduce dispersal
through currents, correspondingly. Estimates of biodeposition rates in
bivalves (up several kg m"^ yr"^) and descriptions of the influence of factors
such as temperature, suspended solids loadings and substrate type, are given
in the review by Lee and Swartz (1980).
Processing by zooplankton can influence the characteristics and
settlement of particles and associated pollutant loads, though deposit-feeders
probably play the most important role in terms of amounts of material sorted.
In some areas 100% of the surface sediment may be pelleted and in
productive regions with low sedimentation rates it is estimated that the sea
floor may be passed through benthos several times each year (Lee and
Swartz, 1980). These pellets sometimes become enriched in pollutants
compared with surrounding sediment as a result of modification during
passage through the gut (Brown 1986). In terms of transport processes,
however, faecal-pellet production by deposit-feeders may alter
characteristics and microbial activity but does not add new material to
benthic sediments and thus differs from the activity of suspension feeders:
strictly speaking, only suspension feeders are involved in biodeposition.
Pellets are compacted and fuse with the sediment matrix as they are
buried. This may be a slow process in relatively low energy environments
36 Geological Environment

(muds and silts) but elsewhere may take place within one day, aided by
metabolic activities of microbes and disturbance from meio- and macro-
fauna. By comparison, pseudofaeces tend to be less compact and are readily
disaggregated. The effects of feeding on the size distribution of particles
(sorting), rates of processing, rates of biodeposition, and properties of
sediments have been reviewed by Rhoads (1974) and Lee and Swartz
(1980). As an example of the potential scale of effects of pelletisation on
deposition, it has been calculated that faecal pellets settle at a rate which is
four orders of magnitude faster than particles of 2-3|Lim diameter, typical of
the food of many suspension feeders (Haven and Morales-Alamo, 1972).
In general, therefore, suspension-feeding organisms are important agents
of deposition in estuarine ecosystems due to their influence on increasing
apparent grain size (through pellet production) and reducing settlement time.
Once entrained in sediments, other biological processes may come in to play.
Highly pelletised sediments have a further, indirect effect on the transport
processes in that the increase in grain size increases porosity and water
content and can lead to greater erodibility.

3.2 Salinity and chemical speciation

Salinity, along with substrate type, is a dominant environmental variable
directly affecting the distribution of organisms in estuaries, although, as
indicated, other factors may modify communities, significantly. Salinity at
any one point varies with the tides, and will be dependent on stratification,
flushing characteristics and conditions of river-flow, and gradients may be
both horizontal and vertical. The osmoregulatory range of any organism will
determine its ability to survive and colonise any given stretch or tidal level
of the estuary. Significantly, however, it is not only the magnitude of the
salinity regime, but also its rate of change, which affects survival: the more
gradual the change the more likely it is that the organism will adapt. In this
context it is important to recognise that reduced exchange between sediment
pore-waters and the overlying water column can buffer infaunal organisms
from rapid change and extremes of salinity.
Salinity not only exerts direct control on survival but can impact on
ecosystem health, indirectly, via processes such as contaminant
geochemistry and bioavailability. The negative effect of increasing salinity,
on uptake and toxicity of divalent cations such as Cd^^, Cu^^ and Zn^"*", has
been established for a number of estuarine organisms. The assumption made
is that increasing complexation of the free ion at high salinities (principally
with chloride) reduces bioavailability (Engel et al, 1981; Wright and
Zamuda, 1987; Campbell, 1995). Competition from Ca at high salinities,
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 37

together with osmotic and other physiological changes, may also contribute
to these observations.
Most complexing agents would be expected to inhibit metal
bioavailability (by reducing the free metal ion concentration or activity),
although it appears that complexation sometimes enhances uptake. For
example, cadmium bioaccumulation rates in mussels Mytilus edulis are
doubled (when compared with the ionic form) by prior sequestration with
pectin or humic and alginic acids, suggesting that these uncharged forms can
be transported across membranes preferentially (George and Coombs 1977).
Furthermore, binding of Cr^"^ to proteins in tannery waste appears to increase
bioavailability disproportionately, resulting in exceptional body burdens in
mussels collected near to outfalls (Walsh and O'Halloran, 1997).
Significantly, some of the most important pollution events to affect estuarine
ecosystems have been caused by metal-organic moieties - including methyl
mercury, alkyl lead and tributyl tin (TBT). The lipid solubility of these
metals is greatly increased by the presence of the associated alkyl groups,
facilitating entry across biological membranes. This is particularly notable
for tin, which, in inorganic form in estuarine sediments, is relatively inert
and seldom accumulated substantially, even in heavily contaminated
estuaries receiving mine tailings containing cassiterite (SnOa). In contrast,
organotins such as TBT are bioconcentrated from sediments to a significant
degree, especially by some infaunal bivalves (Langston et al, 1990).

3.3 Bioavailability and assimilation pathways

Benthic plants and animals of both the sub-and inter-tidal zones are
important components of estuarine food chains, which ultimately depend on
organic debris, bacteria and diatoms associated with estuarine deposits as
sources of nutrition, nutrients and minerals (Figure 4-3). Inevitably,
bioaccumulation of sediment-bound contaminants may occur via a variety of
pathways and results in their transfer to higher trophic levels, including fish,
birds, mammals and, eventually, humans. Contaminants in sediments, and
their bioavailability, thus have important and widespread implications
throughout the estuarine ecosystem and beyond.
Direct assessments of bioavailability in estuaries usually involve field
surveys with bioindicator fauna, such as bivalves, gastropods or worms.
However, the choice of organism is crucial. Because different organisms
have different ranges in estuaries, and accumulate metals from different
sources, there is no 'universal bioindicator' present throughout the range of
environments found in a typical estuarine-coastal zone. Filter-feeders such as
mussels Mytilus edulis and cockles Cerastoderma edule may give some
measure of the importance of contamination on suspended solids, whilst
38 Geological Environment

infaunal deposit-feeding organisms and detritivors (e.g. clams Scrobicularia

plana, Macoma balthica and ragworm Nereis diversicolor) are often useful
bioindicators of bioavailability in estuarine benthic deposits. In contrast,
being primary producers, seaweeds such as Fucus spp are generally
considered good indicators of dissolved (rather than particulate)
contamination. Nevertheless, there may be situations where even Fucus may
assimilate contaminants directly from sediments, for example where fronds
lay in direct contact with highly contaminated muds during low water
(Luoma et al., 1982). It is important to keep in mind that geochemical and
geological influences may vary among different components of the
ecosystem, and in different conditions.
The determination of metal concentrations in bioindicators is of obvious
relevance in ecotoxicological terms, since, in effect, direct measurements are
made of bioavailable - and hence potentially deleterious - metal. Analysis of
sediment (and water) is undertaken more routinely in statutory monitoring
programmes and, though helpful in terms of defining comparative loadings
in the environment, is usually restricted to measurement of 'total' metal
(including refractory forms which may be of little biological significance).
Metal 'speciation' techniques hold out more promise in terms of
understanding and predicting bioavailability, though they have yet to be
widely validated and adopted.
Surrogate chemical measurements of bioavailable fractions are often
derived from sediment extraction schemes used by soil geochemists -
incorporating for example, dilute mineral acids, hydroxylamine
hydrochloride (reducible metals), EDTA and ammonium chloride (ion-
exchangeable forms), or their biomimetic equivalents (e.g. enzymes and
gastric fluids). Similar attempts have also been made to apply this approach
to synthetic organic contaminants, such as pesticides, PCBs and TBT, using
semi-permeable membrane devices filled with organic solvents to determine
extractable sediment fractions. If used appropriately, such extraction
schemes provide information on how contaminants are bound to sediment -
essential in interpreting the geological and geochemical controls on
bioavailability. Important modifying factors which may need to be
considered, alongside the concentration of the contaminant in extracts,
include: grain size, the role of complexing agents (such as iron and
manganese oxides and organic matter), effects of early diagnosis, and the
role of suspended particulate matter.
Some of the earliest demonstrations of the assimilation of particulate
metals by estuarine organisms, and the importance of geochemical
associations, were provided by Luoma and Jenne (1976, 1977). These
included a study of sediment-Cd uptake in clams Macoma balthica which
demonstrated the importance of organic matter in suppressing Cd
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 39

bioavailability; in contrast, Ag, Co and Zn were assimilated readily from

detrital organics (though Fe and Mn oxyhydroxides inhibited the uptake of
the latter two metals). These and similar findings eventually led to the
general hypothesis that bioavailability declines as the strength of binding to
various sediment components increases (in line with principles derived from
complexation studies with dissolved metals).
Speciation techniques for oxidised sediments are thus based, primarily,
on operational methods of uncertain selectivity to extract specific (labile)
metal forms. Nevertheless, despite drawbacks, these schemes provide some
of the most meaningful and practical assessments of sediment-metal
availability. By examining the goodness-of-fit between extracted metals and
body burdens in ubiquitous benthic species, over a range of sites and
conditions, it is often possible to quantify the influence of sediment
contamination and to evaluate the strength of anthropogenic contributions
from different sources. Scrutiny of outliers in the data can also help to
highlight the more important geochemical parameters which modify
bioavailability (Luoma and Bryan, 1982; Langston, Bebianno and Burt,
1998; Ying et al.,1992; Ridgway et al., 2003).
Particularly useful models for predicting metal bioavailability in
estuarine surface sediments have arisen from field-based studies with
infaunal clams and polychates. These confirm the frequent role of the major
metal-binding components Fe and Mn oxyhydroxides, or organic matter, in
mediating uptake. Thus, for many of the examples, metal burdens in biota
[M^] are best described by the linear equation:

(Mil -J-^.c (1)

[x]
where [Ma]x = the concentration of metal associated with sediment
component x, and [x] = the concentration of the metal binding component.
Such normalising routines quantitatively account for the influence of these
major geochemical parameters and indicate that metal impact is unlikely to
be the same in all sediments. The success of extractable Fe or organic
content as normalisers reflects their importance in the partitioning of
adsorbed (non-detrital) fractions in oxidised, estuarine surface-sediments
(which commonly form a major part of the diet of deposit- and suspension-
feeders).
Notwithstanding the excellent correlations between tissue burdens and
sediment fractions that are sometimes achieved, there is still much
speculation as to the physiological mechanisms involved and the processes
surrounding geochemical controls. The simplest explanations assume that
assimilation of particulate metal takes place in the gut and that metal-binding
40 Geological Environment

sediment phases either compete with uptake sites in the digestive epithelium
or render the metal less labile. Nevertheless, even in estuarine organisms,
which are thought to derive most metals from ingested surface sediments,
accumulation of some metal from interstitial and overlying water cannot be
ruled out (Langston and Spence, 1995).
Provided that the system is at adsorptive equilibrium, it may not be
necessary to separate uptake sources in order to model bioavailability. The
amount of soluble metal [M] will be a function of that which is adsorbed
[Ma] to a complexing sediment phase x (i.e. [Ma]x), together with the
concentration of the solid phase {x\ responsible for binding that metal. Thus,
the ratio

{Md\x

(used in equation 1 as a predictor of bioavailable sediment metal) would also

represent a surrogate of dissolved metal. If more than one sediment phase is
important in complexation, then additivity of this term, for each phase, is
assumed (Tessier e^ a/., 1993).
This hypothesis receives some support from experimental studies with
artificially manipulated particulates (e.g. Luoma and Jenne, 1977) in which
sediments exhibiting the greatest rates of sediment to water desorption
(lowest K j values) were also those from which metal bioaccumulation was
greatest. Where such labile sedimentary sinks occur in the field,
bioavailability of particulate metals could be enhanced both through
increased assimilation from ingested material and from the higher
concentrations of desorbed metal. However, whilst this concept might be
expected to work well with Cd and other metals whose particle reactivity is
relatively low, the assumption of steady -state conditions would be less
realistic for metals whose behaviour is dominated by strong and complex
interactions with sediment.
Not surprisingly, attempts to predict metal bioavailability, based on
equilibrium partitioning, are sometimes unsuccessful, even after
normalisation with respect to Fe or organics. This applies particularly when
sediments of widely differing characteristics are compared. Intuitively, other
site-specific geochemical parameters such as sulphides would be expected to
exert more control under anoxic conditions and might conceivably determine
bioavailability of metals from reducing sediments. The validity of traditional
extraction schemes is questionable in these circumstances, however, and the
relevance of buried, reducing sediments as a metal source for many
organisms is uncertain and perhaps highly variable.
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 41

Anomalously high levels of Cu have been observed in estuarine clams

S.plana and M. balthica from relatively anoxic sites, even though
contamination with Cu is not evident in the sediments themselves. An
increase in bioavailable cupric ions during bouts of anoxia is one possible
explanation; alternatively, enhanced Cu burdens in clams could result from
immobilisation in tissues, as CuS, following an influx of hydrogen sulphide
(Bryan and Langston, 1992). The presence of high levels of acid volatile
sulphide (AVS) in sediments is reported to modify the impact of Cu and Ni
in amphipod bioassays (Di Toro et al, 1992), though, in contrast to the clam
scenario, high levels of sulphide in this case are thought to reduce
accumulation and toxicity (where sulphur-to-metal ratios in sediment are >1,
metal is assumed to precipitate as insoluble, unavailable, metal sulphide). A
high total S content of sediments has also been shown to lower Pb
availability to mussels M.edulis, close to a Pb/Zn smelting complex, though
whether this is a competitive phenomenon, or the result of the anomalously
sulphide-rich nature of ore-impacted sediment, is unknown (Bourgoin et aL,
1991).
In summary, concentrations of metals (and presumably other
contaminants) in estuarine biota are significantly influenced by the major
geochemical parameters. For aerobic surface sediments, metal availability is
often determined, predictably, by adsorption/desorption characteristics on
Fe/Mn oxyhydroxides and organic coatings. The bioavailability of metals in
anoxic (usually sub-surface) sediments is more uncertain but appears to fall
under the control of sulphide reactivity: metal ions whose solubility is less
than FeS (e.g. Cd, Ni, Cu, Zn, Hg and Pb) may be precipitated from pore-
water as insoluble metal sulphides which are relatively unavailable, at least
in some organisms. In contrast, in other estuarine infaunal species anoxic
conditions appear to enhance bioaccumulation.
In situations where adsorptive equilibria cannot be assumed, predictions
of bioaccumulation may be confounded by the fact that estuarine organisms
accumulate metals from a combination of water and diet (including
sediments) in varying proportions. Development of kinetic models, where
individual pathways are treated additively, may help to interpret uptake
routes and efficiencies, and transfer through aquatic ecosystems. Metals in
solution are usually considered to be more bioavailable than solid-phase
metal; however, the higher concentrations in the latter often render dietary
vectors more important. This may be further enhanced in filter feeders
through the selection of certain particle types, relative to bulk sediments.
The importance of assimilation efficiencies and particle type in
determining bioaccumulation of solid-phase metals has been illustrated in
estuarine sediment-dwelling clams where, for example, highly efficient
assimilation of Se (93% from epipelic diatoms) accounts for a predominantly
42 Geological Environment

particulate uptake route. Similarly, the presence of adherent bacteria and

extra-cellular polymers on the sediment particles selected by M balthica
enhances the digestibility and bioavailability of Ag, Cd and Zn. Cr
assimilation also appears to be much more efficient from bacteria and
polymers than from diatoms or purely inorganic sediment fractions (Decho
and Luoma 1991,1994, 1996; Harvey and Luoma, 1985; Luoma et al„
1992). Further insights into assimilation pathways and efficiencies, from
particulates of different type, would be useful.

3.4 Geochemical Influences: Case Studies - Ecological

effects of contaminated estuarine sediments.
Tributyltin (TBT) is worth singling out as a 'model' pollutant which has
been responsible for significant impairment to estuarine ecosystems (as
result of leaching from antifouling paints on boat hulls), imposex in
dogwhelks Nucella lapillus being a particularly notable effect (Bryan et al.,
1986; Minchin et al., 1995). In the 1980s, recognition of damage to non-
target organisms, particularly molluscs, instigated legislation to ban the use
of TBT paints on boats <25m length (which encompasses the majority of the
leisure market). TBT levels in waters of many marinas and small-boat
harbours diminished considerably, following restrictions. However, a slow-
down in the rate of disappearance of TBT has been described for some ports
and estuaries. Delays to further improvements in water quality are partly
related to sedimentary sinks of the compound, together with certain
dockyard operations and the continuing presence on larger vessels (still
entitled to use existing TBT antifouling until 2008, though no longer
permitted to apply new coatings, according to edicts of the International
Maritime Organisation). There is also a possibility that TBT paints are still
being used on some craft, illegally.
TBT concentrations in sediments are highest, as expected, close to
dockyards, marinas, and hull-cleaning facilities, whilst chronic
contamination may be detected in deposits at considerable distance from
TBT sources. Partitioning is reversible, allowing release back to the water
column, but because of the relatively high affinity for particulate matter,
residence times for TBT in sediment are usually prolonged in comparison
with overlying water, especially in unperturbed, organic-rich fines. Paint
particles containing TBT may also become entrained in sediments,
increasing the persistence of the biocide (Thomas et al, 2003). Degradation
of TBT can occur in surficial sediments through the activity of
microorganisms, and involves stepwise debutylation to inorganic tin.
Nevertheless, TBT removal rates are relatively slow, with half times of the
order of 1-5 years. In undisturbed anaerobic muds, and at sites where there is
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 43

a continuing input of TBT from antifouling, temporal reductions in TBT

sediment-loadings may be undetectable; half-lives under such conditions are
likely to be of the order of decades (Langston and Pope, 1995).
Consequently, the retention of TBT in fine-grained muds is seen as one
of the major reasons for the continuing threat from the compound in poorly
flushed estuaries and harbours. The concerns are two-fold: firstly, slow
release of TBT held in the sediment reservoir, to the water column, may
extend exposure to concentrations above no-effects thresholds for pelagic
species (including larval forms). Secondly, some burrowing species,
particularly deposit-feeding clams such as Scrobicularia plana, derive
potentially deleterious burdens of TBT directly from sediments and, as a
result, have been in decline at TBT-contaminated sites. Where TBT persists
in sediments, clam populations are unlikely to recover quickly. This
scenario, where sediments increasingly become a major source, is probably
representative for a broader range of contaminants, following the widespread
introduction of measures to reduce discharges to estuarine waters in recent
years.
Areas exhibiting metal contamination from mining or smelting operations
often provide further examples of ecological impact and serve to illustrate
both geochemical and geological influences on estuarine ecosystems.
Reduced biodiversity is attributable to metals in such cases, though in fact
this phenomenon is symptomatic of most forms of anthropogenic
disturbance (Langston, 1990). The Fal Estuary in south-west England is a
case in point of such a metal-impacted system (Figure 4-4).
Compared with most estuaries in south-west England, the Fal, as a whole,
has a very low abundance of a number of sensitive taxa, notably benthic
crustaceans and molluscs, whilst certain small annelid worms are more
abundant (Rostron, 1985). Metal pollution is implicated as a major factor
responsible for these differences (Langston et aL, 2003b). The conspicuous
absence of bivalve and gastropod molluscs from highly metal-contaminated
sites in the Fal (notably Restronguet Creek) is a consequence of the long
history of metal mining in the region (Bryan et aL, 1987). Here, for example,
high levels of Cu and Zn in sediment inhibit the settlement of juvenile
bivalves, including Scrobicularia plana, Cerastoderma edule and Mytilus
edulis. In contrast, enhanced metal tolerance is observed in some organisms
(e.g. the polychaete Nereis diversicolor), and is effective in ameliorating
impact. Different sensitivities to metal-laden sediments, between taxonomic
groups, may thus translate into community-level effects and Restronguet
Creek has developed a distinct macrofaunal community composition
44 Geological Environment
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 45

. r*. X, i.j'iM

' T ^ f i 'n^AiUT.-'t*^-,

... ,1. ... * J .;i..> * ti ifk^»V\ . .) Mv..M

wi \Sf •-v.v) *i;. /Is ^ til«JO
1. ,n*r;l »r: Wid ;,il^ .Sr-it-w l.^ilh«f; AiH^i.

• j-/fc*i)tr-yl»vwV-elr*>-\T'h -Mil-»'V'' Stflj > ^, tht iil i ^o I « sifT, ^• sf ' , .',

F/gwre ^-5. The distribution of biotopes in the Fal Estuary, Southwest England. Upper Fal
(top), Lower Fal (bottom) (reproduced from Moore et al, 1999, with permission JNCC).
Legend to the top figure: Sheltered littoral rock with fucoids, Sandy mud shores with Hediste
and Macoma, Littoral soft mud with Hediste, Shallow sublittoral rock with kelp and sponges
and mixed gravel with sabellid worms, anthozoans, ascidians and polychaetes. Shallow
estuarine mud. Oysters Ostrea edulis beds. Legend to the bottom figure: Exposed/moderately
exposed bedrock shore with fucoids. Sheltered littoral rock with fucoids. Sheltered littoral
rock and mixed substrata shores. Mixed substrata shores. Steep upper shore bedrock and
sheltered lower shore mixed substrata with fucoids, Sandy mud shores with Hediste and
Macoma, Littoral soft mud with Hediste, Sublittoral moderately exposed rock with Laminaria
hyperborean, Sand-scoured rock outcrops with mixed kelps, Sublittoral mud with bedrock
outcrops. Sheltered sublittoral rock with Laminaria saccharina, Sheltered sublittoral rock
with kelp and sponges and muddy gravel with sabellid worms, anthozoans, ascidians and
polychaetes, Sublittoral marine mixed sediments with sponges and ascidians, Sublittoral
sediments with Zostera marina beds, Sublittoral muddy gravel, Sublittoral estuarine mud with
kelp on available hard substrata, Sublittoral estuarine mud, Sublittoral beds.
46 Geological Environment

compared to other, less-contaminated creeks in the Fal system (Warwick, et

al 1998). Sediment copper concentrations in Restronguet Creek are in the
region of 2,500 |Lig g"\ whereas in other creeks in the Fal, Cu concentrations
range from 100-1,200 |Ligg\
There is a gradation in meiofaunal (nematode and copepod) community
structure in different parts of the Fal, which, as with macrofauna, is
consistent with increasing metal concentrations, particularly Cu (Warwick et
al, 1998). Large differences in the Cu tolerance of nematode communities
from different creeks have also been shown to correlate with their previous
history of exposure to Cu in sediment (Millward and Grant, 2000). A level of
200|ag g'^ sediment Cu (IM HCl extractable) has been suggested as the
threshold above which 'pollution-induced community tolerance' is initiated.
It is uncertain whether these thresholds would be of universal significance
however, as many communities in the Fal may be partially adapted to
survive high metal concentrations.
Similar ecological effects have been observed in polluted sediments from
other locations. For example, studies in metal-contaminated Norwegian
Fjords indicate considerable reductions in faunal diversity at sediment Cu
concentrations above 200 jiig g'\ with sensitive species, including molluscs,
lost, leaving a high proportion of (tolerant) polychaetes (Rygg, 1985). Large
reductions in shellfish production have been linked to impact from
metalliferous mine spoils in estuaries in Goa (Parulekar et al, 1986) and in
Australia, smelting wastes (hazardous because of their chemical and
biological reactivity) have been shown to result in changes in seagrass
communities (Ward et al, 1984).
Despite recent trends towards improved water quality, contamination in
estuarine sediments often remains consistently high (as in Restronguet
Creek, following closure of the mining industry). Benthic communities also
appear to retain their modified characteristics for long periods. The
persistence and behaviour of contaminants in sediments is clearly an
important geochemical feature, which may govern the rate of 'recovery' of
impacted estuarine ecosystems long after man's polluting influence has
abated. There is still much to be learned concerning timescales and
mechanisms involved in this process.

4. CONCLUSIONS
In conclusion, estuarine ecosystems are significantly influenced in a
variety of ways by geological and geochemical characteristics, particularly
GEOLOGICAL AND GEOCHEMICAL INFLUENCES 47

the nature of benthic deposits. The over-arching effects of varying tidal and
sahnity regimes introduce a further level of intricacy to an already
complicated and dynamic environment. Understanding the key processes and
interactions responsible for spatial and temporal variability in biological
communities is an essential requirement for predicting future change in the
functioning of these most complex of ecosystems, and for distinguishing
between natural and anthropogenic causes.