Middleton G.V. (Ed.) Encyclopedia of Sediments and Sedimentary Rocks
Подождите немного. Документ загружается.
TILLS AND TIIIITFS
745
Table TI Classilicaiinn of Tills imodilied from Dreimanis, 1988: with pers tomm, van dcr Moor, 2002)
Derivation
of
Glacial Debris
Position ot
Debris in
Transport
Location
Process
Deposits
Till types
by
Subglacial Processes
Subglacial "•
Subglacial
1'
1
•
—* Englacial "•—
.
Ice-Marginal
Extraglacial Processes
^ Supraglacial |
{Supraglaciall
Deformable Bed
Immobilization
-odgement
Till
Melt-out
Till
Flow
Till
Waterlain
Till
Tills may t'orm within several distinct glacial sub-
en vironmcnts: subglacial, englacial. proglacial (terrestrial/
subaquatic) and supraglaeial: and may be deposited from
acti\e temperate wet-based and polar cold-based ice masses,
under stagnant ice conditions, and subaquatic environments of
both grounded and Hoating ice masses (Figure T15{A),(B)).
Table TI illustrates the complexity and interrelationships that
exist within till-lbrming glacial environments.
Under wet-based, active temperate conditions considerable
input from mellwatei- ean be expected, as well as thermal
tluctualions. and transient fluctuations in ice velocity and
stress conditions. Subglacial tills formed under active tempe-
rate conditions are recognized as Lodgement. Klow. and
Melange, with a variety of local till sub-types. Sub-types of tills
are often identified owing to some peculiar characteristic of
grain size, provenance and/or structural attribute. Where
ephemeral passive ice conditions occur within active ice masses
Mell-out tills ean be deposited. Although comparatively rare,
inidei- dry polar climates sublimation or mcltout tills ean form.
Both lodgement and meltout (ills have been extensively
described in the literature (Dreimanis. I9S8: Brodzikowski
and Van Loon. 1991)., whilst subglaeial melanges have only
recetitly been recognized as separate, distinct subglacial
sediments (Curry eitiL. 1994: Hoffman and Piotrow'.ski, 2001;
Menzies and Shilts. 2002). These subglaeial tills are formed as
a consequence of: (1) high shear stress resulting in debris being
released from basal ice in a "smearing on" process where high
pervasive shear stress results in ice pressure melting and debris
release {Lodiicmau liU): or (2) by slow undermelt releasing
debris gradually {Mcfi-ouiiilf) (Shaw. 1987); or (3) as a result
of portions ofa sofl deforming subglacial bed either reaching
shear strengths higher than the mobilizing basal ice stresses
due to onset oi freezing temperatures, or reduced porewater
content, or porewater pressure reduction due to reduced
overburden stress (Menzies. 1989;
1-uller
and Murray. 2000)
producing Mchiiigc: and fmally; (4) by the glaeial crushing of
basal debris and bedrock (H-bed conditions. Menzies. 1989)
resulting in a form of pulverized sediment being produced
{pcr.s.conim. van der Meer, 2002).
Witbin the Supragiacial environment (Figure TI5(A)) tills
have been recognized as Flow tills where sediment approach-
ing an effective stress level of zero slip, slide, and flow off the
ice (Boulton. 1968); and. where temperatures and debris cover
permit, sediments are directly released (melt-out till) from the
ice by melting (Bouton. 1970) or. in rare arid conditions, by
sublimation (Shaw, 1977). The degree to which meltwater is
associated with and inlluences the meehanics and subsequent
deposition of a melt-out till lemains somewhat controversial
(Rappol. 1987; Paul and Hyles. 1990).
A complex set of conditions can be found within Proglacial
environments whether terrestrial or subaquatic. In terrestrial
environments (Figure TI5(A)) the dominant process of till
formation tends to be by mass movement thus flow tills are
formed, and where debris-laden ice. in some instances buried
and re-exhumed in the proglaeial environment, melt-out thus
producing melt-out tills. Where lakes exist deposition may
746
TILLS AND TILLITES
Lodgement Till
Melt-out Flow Tili ^
SUBAOUATIC
Till
Figure T15 Locilions ot till types wilhiti terrestrial ancl equtitic ice
margins (arlaptecl and revised trom Menzies and Shills, 2002).
occur by rain out through the water column from debris-laden
meltwatcr streams or floating ice rafted debris producing
Watfiiain Tills.
As Figure TI5(B) illustrates a complex set of sub-environ-
ments exists where ice masses either float or are grounded
within large bodies of fresh or brackish water. Where a floating
ice mass exists large amounts of debris are likely to come off
the surface of the iec. from ice clitT melting, from ealving ice
bergs,
and from engiacia! and subglacial meltwater tunnels.
The result of these sediment release processes is to cause an
enormous rain-out of sediment through the water column
resulting in a Wateriain Till being deposited on the lake or
ocean floor. Beneath grounded ice masses sediment will be
squeezed, pushed or otherwise extruded from the subglacial
contact at the ice/bed grounding line. Where debris is tnobile
and extruded from the grounding line a Melange, perhaps in
the fomi of a ""til! delta or tongue", may form (Alley etui.,
1986;
King and Fader. 1986; Anderson elai. 2001).
There are several mistaken ideas eoneerning tills and tillites
that persist, viz.. that tills typically exhibit a bi-modal grain
size distribution, are overconsolidatcd, have a very low
permeabihty, are unstratified. homogenous, massive, and
unstructured, and. finally, that tills of differing colors are of
different depositional phases and therefore indicative of more
than one glaciation. As in all niylhs there are occasional truths
but, as a general rule, all are essentially incorrect. Where tills
can be seen tbrming at present the usage of the term till seems
appropriate. Where, as is so often the ease, cither the actual
means of deposition is obseured or the sediments are ancient.
then the till-forming mechanisms and environments within
which they were originally deposited arc unknown: then the
nongenetic terms diamict or diamicton tor unlithitied sedi-
ments and diamictitc for lithified seditnents may be preferred.
A persistent, and perhaps unsolvable problem, remains in
identifying as tills those sediments that are marine as opposed
to glaciomarinc. With the advent of glacial micromorphology
a clearer understanding ofthe internal insitu stratigraphy and
microstructurcs within till has great potential in furthering our
understanding of this complex, and often enigmatic, glacial
sedimeni (Van der Meer. 1996; Metizies. 20D0).
John Menzies
Bibliography
Allc>, R.B.. Bhiiikeiiship. D.D., Bcnllcy. C.R.. and Rooney. S.T..
i9Sf).
Deformatioti ol till beneath ice stream B, West Antarclica.
Sittitrv. 322: 57-59.
America Geological Institute (AGI). 1997. Glossuiyoj Geology. 4th
L'dti. Alexandria. Virginia. 769pp.
Anderson, J.B.. Smilh WclInLT, J., Lowe. A.L.. Mosola. A.B.. and
Shipp. S.S,.
2(K)1.
I ootprim of the expanded WestAnlartic Ice
Sheet; Ice stream history and behavior. GSA Today. 11 (10): 4 9.
Brodi-ikowski. K.. and Van Loon. A.J., 1991. Glaeigenic Sedittictits.
Dc\eloptnents in Sedimentology. 49. Elsevier Seienee Publicalions.
('i74pp.
Boulion. G.S,. 1968, Flow tills and some related deposits on sotne
Vestspitsbergcn glaciers. JotirnalofGhiciohgy. 7: 391 412.
Boiihon. G.S.. 1970. On the deposition of subglacial and mell-oiil lills
at lhe margins of eertain Svalbard glaciers. .lotiniidoKiltteiologv. 9.
231 245.
Boulton. G.S.. atid Deynoux. M.. I9K1. Sedimentation in glacial
environments and the idcntitieation of lills and itllites in aneient
sedimentary sequences. Prceanihrian Research. 15; 397-422.
Ciirry. B.B.. Tmost. K.C.. and Berg, R.C.. 1994. Oiialeriiar\ Geology
of the Martinsville Alternative Site. Clark C'ounly. Illinois. Illinois
Slate
Geological
.Stiriey. Circular 556: 85pp.
Drcimanis. A.. 1988. Tills; their geneiic lerminology and classification.
In Goldthwait. R.P.. & Malsch. CL. (eds.l. GeneticClassificationof
Gktcigeniv Deposits. A.A. Balkema. pp.
17-83.
Dreimanis. A., and Lnndqvisl, J.. 1984. What should be called till.'
Striae. 20: 5- II).
I uller. S.. and Murray. T.. 2000. Evidence against pervasive bed
deformation during the surge of an Icelandic glacier. In Maltman.
A.J.. Hubbard. B.. and Hambrey, M.J. (eds.). Defornuilioii
oJ
Gla-
citd Materials. Geological Society of London, Special Publication.
176,
pp. 203 216.
Geikie. A..
1K6.1.
On the phenomena ofthe Glacial Drift of Scotland.
Geologieal
Societv of
Cila.s{>o\v.
Tnittsaetiuns. 1(2); 190pp.
Hambrey. M..I.. and Iiarland. W.B.. (eds.). 1981. Earths Pre-Plcisto-
cene Glacial
Record.
Cambridge University Press.
Hoffman. K,. and Piotrowski. J.A..
201)1.
Till melange at
Amsdorf,
cetitral Germany: sediment erosion, transport, and deposition in a
complex, soft-bedded subglaeial system. SedimemarvGeology. 140;
215-234.
King. L.H.. and lader. G.BJ.. I9K6. Wiseonsinan glaciation of ibc
Atlaniic coniiiienial shelf of southeast Canada. Geological Survey of
Canada Bitltetiii. 363: 72pp.
Menzies. J.. 1989. Subglacial hydraulic conditions and their possible
impact upon subglaeial bed iormalion. Sctlitnentary Genlonv. 62;
125 150.
Menzies.
.1,.
2000. Micromorpholngieal analyses of mierofabrics and
mieroslruetnres. indicative of deformation proeesses in glacial
sediments. In Maltman. A.J.. Hubbard. B,. and lfambre>. M.J.
(eds.).
Dejiirtiiatiiiti ot Glacial Materials. Geological Soeiety of
London. Speeial Publieiition, 176. pp. 245 257.
Menzies, J., and Shilts. W.W.. 2002. Subglacial Enviromnenis, in
Menzies. J.,
(ed.},
Xfodcrti and Past Glacial LtivirontnvtUs. Bulier-
worth-Heinemaii Publishers. Chapter 8. 183 278.
TOOL MARKS
747
Paul, M.A.,
and
Eyies,
N.,
1990. Constraints
on the
preservation
of
diamict facies (Melt-out tills)
at the
margins
of
stagnant glaciers.
Qualernciry Science Reviews,
9:
51-69.
Penck, A., 1906. Siid-Afrika und Sambesifalle. GeographicheZeischrifl,
12:
600-611.
Rappol, M., 1987. Saalian till
in
The Netherlands:
a
review.
In
van der
Meer, J.J.M., (ed.). Tills
and
Glacioieclonics.
A.A.
Balkema
Publishers, pp. 3-21.
Ruszczyriska-Szenajch,
H.,
2001. "Lodgement till"
and
"deformation
till".
Quaternary Science Reviews, 20:
579-581.
Shaw,
J.,
1977. Tills deposited
in
arid polar environments. Canadian
Journal of Earth Sciences, 14: 1239-1245.
Shaw,
J.,
1987.
Glacial sedimentary processes
and
environmental
reconstruction based
on
lithofacies. Sedimentology, 34: 103-116.
Van
der
Meer, J.J.M., 1996. Micromorphology.
In
Menzies,
J. (ed.)
Past
Glacial
Environments. Sediments.
Forms,
and
Techniques.
Glacial
Environments II. Oxford: Butterworth-Heineman, pp. 335-355.
Cross-references
Classification
of
Sediments and Sedimentary Rocks
Deformation
of
Sediments
Glacial Sediments: Processes, Environments,
and
Facies
Melange; Melange
TOOL MARKS
A tool mark
is a
mark produced
by the
impact against
a
muddy bottom
of a
solid object driven
by a
current moving
over
the
bed.
It is
generally preserved
as a
cast, seen
on the
base
ofa
sand
or
silt bed deposited on the muddy bottom soon
after
the
marks have been formed.
The
term
was
originally
proposed
by
Dzulynski
and
Sanders (1962, p.72).
A
large
number
of
different types were named during
the
1960s
by
geologists, mostly from observations
on
the
soles
of
turbidite
sandstones (see Turbidites, Dzulynski and Walton, 1965; Allen
1982,
volume
2,
chapter 13). Tool marks
are
also found
in
other environments, however, notably
in
fluvial deposits
and
marine storm deposits (see Storm Deposits, Gutter and Gutter
Cast).
They
are
valuable indicators
of
paleocurrent direction
(see Paleocurrents)
and
also give some information about
the nature
of the
clay bottom
and the
sediment transport
mechanism operating
in the
flow that deposited
the
over-
lying sand.
Allen (1982, pp. 509-520) classified tool marks into: (i) drag
marks, formed
by
protracted contact
of a
tool moved
by
traction; (ii) roll marks, formed by protracted contact
ofa
tool
rolling over the bed; (m) prod marks, formed by
a
brief contact
of
a
tool with
the bed; (iv)
tumble marks (alsocalled skip
marks) made
by a
saltating tool that produced
a
series
of
marks; (v) skim marks, produced
by
grazing contact
of a
tool
with
the
bed;
and (vi)
marks related
to
near approaches,
a
category which includes chevron marks (see Figure T16;
and
see Allen, 1982,
for
references
to
original sources
for
these
terms).
Some authors (e.g., Ricci Lucchi, 1995, pp.134-139)
distinguish bounce casts, with
an
almost symmetrical shape,
from prod marks which
are
asymmetrical (deeper
on the
down-current side)
and
therefore permit determination
of
the current direction. Brush marks
are
similar
to
prod marks
but have wrinkles produced by drag
of
mud
on
the up-current
side.
For
good photographs see Allen (1982)
or
Ricci Lucchi
(1995).
Gerard V. Middleton
r-PROTRACTED CONTACT
H- BRIEF CONTACT
A-DRAG
\\\\\\\\\\\\\\\V^\\\\\\\\\\\\\\\
Ploughlngs,
smooth,
pinnote fractures
B-ROLLS (e.g. ommonite)
niiiiiinununmiinnmniiiumimnuiimii
I-Prods
(e.g.
plant stems)
A\\\\\\\\\\\\\\\\\\\\\\\\\\V
z
2-Skims
\\\\\\\\\\\\\\\\\\\\\\\\\\\\\^
I-Tumbles (e.g. cube)
2-Soltates
J
<?4)
t<^
. Motion
of
centre
of
tool
Motion
of
point
on
surface
of
tool
.
Figure T16 Classification
of
tool marks, based on tool kinematics (from Allen, 1982, p.510).
748
TOXICITY OF SEDIMENTS
Bibliography
Atlen, J,R,L,, 1982, Sedimentary Structures: Their Character and Physi-
cal Basis, Volume 2, Elsevier,
Dzulynski, S,, and Sanders, J,E,, 1962, Current marks on firm mud
bottoms. Transactions of the Connecticut Academy of Arts and
Sciences, 43: 57-96,
Dzulynski, S., and Walton, E,tC,, 1965, Sedimentary
Features
of Flysch
and
Greywackes.
Etsevier,
Ricci Lucehi, F,,
t995,
Sedimentographica: A Photographic Atlas of Se-
dimentary Structures, 2nd edn, Columbia University Press,
Cross-references
Gutters and Gutter Casts
Storm Deposits
Turbidites
TOXICITY OF SEDIMENTS
Overview
As water quality has improved over ttie past three decades in
North America, diffuse sources of pollution such as storm-
water runoff and sediments are now recognized as long-term,
widespread pollutant sources to aquatic systems. Substantial
impacts on the ecosystem from sediment-associated contam-
inates range from direct effects on benthic communities to
substantial contributions to contaminant loads and effects on
upper trophic levels through food chain contamination (e.g,,
McCarty and Secord, 1999),
Sediment contamination primarily occurs because many
chemicals bind to organic or inorganic particles that eventually
settle to the bottom of our streams, rivers, reservoirs, lakes,
estuaries, or marine waters. Once contaminants bind to a
particle surface or sorb into its interior matrix, they become
less likely to be biotransformed and desorption is usually very
slow; therefore, sorbed contaminants will reside for long
periods in the sediment. Sediment-associated contaminants
tend to accumulate onto small, fine-grained particles and settle
in depositionai areas. This is promoted largely by the very high
surface area and the tendency for higher concentrations of
organic matter in the fine particles.
From a contamination perspective, we understand that
sediments are extremely important to the food web and serve
as a habitat for the benthic community as well as contaminant
reservoir for bioaccumulation and trophic transfer. Once
chemical contamination reaches a concentration whereby it
causes adverse effects to biota, then it is considered polluted.
The greater the evidence of sediment contamination within a
watershed, the more likely a fish consumption advisory exists
(US EPA, t997).
Extent of the problem
There are over 9 billion cubic meters of surface sediments
(upper
5
cm) in the United States of which approximately
1,2 billion can currently be considered contaminated to levels
that pose potential risJcs to fish and to humans and wildlife
who eat fish (US EPA, 1997), This estimate is derived from a
literature search on sediment contamination and represents
65 percent of the watersheds in the continental US, However,
only 11 percent of the nation's rivers had data with any
toxicity information available and of those 77 percent had at
least one station that was sufficiently contaminated such that
adverse effects are probable or possible. This leaves 89 percent
of the data sets with no available toxicity data (US EPA,
1997),
Therefore, the true extent of sediment contamination
remains unknown.
Sediment associated contaminants are found in every type
of aquatic environment within the US from mountain streams
to large rivers, from small lakes to the Great Lakes and in
estuaries and bays (US EPA, 1997), Further, recent investiga-
tions have found sediment pollution in areas previously not
thought to be contaminated (e,g,, estuarine areas of North
Carolina, Pelley, 1999), While the true extent of sediment
contamination is not tcnown, it is apparent that huge quantities
of sediments in industrialized countries are contaminated with
metal and organic chemicals at levels that pose risks to both
aquatic life and to humans consuming aquatic species.
Effects on ecosystems and components
Despite a concentrated research effort over the past two
decades, our ability to assess the impact of contaminated
sediments on ecosystems remains a challenge that is inherent in
the very nature of sediments. The most obvious effect in most
assessments of sediment contamination is that of a direct
adverse impact on the benthic (or bottom-dwelling) commu-
nity, (e,g,, Peterson et al., 1996), Because alteration of the
benthic community structure can result from the mixture of
contaminants present as well as from potential for confound-
ing factors, such as poor habitat, high flows, or low dissolved
oxygen in the water column, it is often difficult to establish
cause and effect for sediment associated contaminants. Studies
of field sites where direct biological effects are obviously
contaminant-induced, are often located in watersheds contain-
ing high levels of industrial and shipping activities. Here,
sediments are likely to be visibly contaminated with petroleum
products and may contain a myriad of metals, pesticides, and
other synthetic nonpolar organic chemicals, e,g,. Grand
Calumet River, Indiana, Thus, assigning cause-effect between
single, specific chemicals and benthic communities becomes
nearly impossible. Most of the detailed biological assessments
dealing with contaminated sediments have dealt with fresh-
water systems, although several sediment risk assessments are
currently underway in coastal harbor areas. Unfortunately,
most of these risk assessments are only published in the "grey"
literature and difficult to obtain.
It is not just infaunal (burrowing) benthic invertebrate
organisms that demonstrate effects from sediment-associated
contaminants. Many bottom-feeding fish have hepatic and
epidermal neoplasms (tumors). There is a strong link between
contaminated sediment and fish neoplasms in all the major
water bodies of the US (Harshbarger and Clark, 1990), The
sediment tintc for these neoplastic growths was clearly
demonstrated in the Black River in Lorain, OH, The higji
incidence of neoplasms in brown bullhead showed a high
correlation with high PAH sediment concentrations. These
neoplasms resulted in age specific mortality in the bullheads.
Once most of the PAH contaminated sediments were removed
by dredging, the incidence of neoplastic growths declined
sharply (Baumann and Harshbarger, 1995),
Sediment-associated contaminants have also become linked
via food web transfer to impacts on upper trophic levels. Such
transfer occurs with mercury and some organochlorines, such
TOXICITY OF SEDIMENTS 749
as PCBs and DDT that are poorly biotransformed and are
hydrophobic; however with other chemicals these connections
are more difficult to establish. From modeling exercises, food
web transfer of persistent contaminants is important for
maintaining the chemical concentrations observed in upper
tropic levels and the benthic component is essential to account
for the observed concentrations (Thomann et al., 1992;
Morrison etal., 1996), Trophic transfer of sediment-associated
contaminants has been documented in both freshwater systems
(e,g,, Lester and Mclntosh, 1994) and marine systems (e,g,,
Maruya and Lee, 1998), In Saginaw Bay, Lake Huron, tree
swallows were found to accumulate PCB from sediments. In
some areas of the Great Lakes and in the Hudson River, NY,
system reproductive damage has been observed for this species
directly linked to PCB (Bishop et al., 1999; McCarty and
Secord, 1999), Further, impacts to birds were observed in the
Saginaw Bay system with reproductive effects on the Caspian
terns following a 100 year flood event suggesting impacts
through food web transfer from a latent sediment source
(Ludwig et al., 1993), Thus, fresh pulses of contaminated
sediments from the watershed can recharge the system
resulting in impacts on the receiving system.
Developing cause-effect links
The complexity of the distributions of contaminants, nutrients,
other sediment/habitat characteristics and resident biota make
accurate determinations of exposure and effects difficult, but
not impossible. This reality suggests that accurate assessments
of contaminant effects use a weight-of-evidence approach
(i,e,,
gathering supporting data from several lines of investiga-
tion) to determine the extent or magnitude and cause of the
contamination problem. The initial ideas on the
weight-of-
evidence approach took data from sediment chemistry, benthic
community structure and sediment laboratory bioassays to
establish a triad of information that could be used to
determine hazard (Chapman, 1986), This approach is con-
tinually improving with additional development on sediment
bioassays (toxicity tests), interpretability of sediment chem-
istry, and the addition of new types of information such as
in situ toxicity tests and residue (tissue chemical concentra-
tion)-effects information.
Sediment toxicity testing on whole field collected sediments
and laboratory dosed sediments are an essential
line-of-
evidence that the contaminants in sediment can produce the
effects that are observed in benthic communities. The earliest
sediment bioassay was performed in the early 1970s and
demonstrated avoidance behavior by amphipods for contami-
nated sediments. However, much of the work demonstrating
the effect of sediment-associate contaminants came in the
1980s, By the mid 9O's, standardized methods for whole
sediment toxicity testing occurred within the US EPA,
American Society for Testing and Materials (ASTM), and
Environment Canada, These tests measured acute (short-term
<tO days) toxicity in benthic macroinvertebrates such as
the amphipods Hyalellaazteca. Rhepoxynius
abronius,
Ampelis-
ca ahdita, Eohaustorius estuarius and Leptocheirus plumulosus
and the midges Chironomus tentans and Chironomus riparius
(US EPA, 1994a,b; ASTM, 2000a), The primary measured
response was mortality, but in the case of the midge, growth
was included and reburial was an additional endpoint for the
Rhepoynmius ahronius. Unfortunately, most sediment toxicity
test methods are focused on acute (short-term exposure) and
not chronic (long-term exposure) toxicity. The measures of
acute toxicity are not often adequate to detect the impacts on
benthic communities. For instance, the 10-d test with Rhepox-
ynius ahronius was not sensitive enough to describe the loss of
amphipods from Lauritizen Channel in San Francisco Bay
(Swartz etal., 1994), In reality, chronic toxicity is the more
pervasive problem and it is the chronic responses, such as
changes in reproduction that lead to population level
responses. While recently developed chronic methods (US
EPA, 2000, 2001) will greatly aid our ability to determine if
sediments are toxic, their long duration and increased costs
may impede their widespread adoption.
A large number of other species have been used for
determining the toxicity of sediments, ranging from bacteria
to fish and amphibians (Burton, 1991). The commonly used
test species encompass trophic levels ranging from decom-
posers and producers to predators and representing multiple
levels of biological organization. Comparisons of their sen-
sitivities have shown a wide range of responses to different
types of sediment contamination, with an equally wide range
of discriminatory power (ability to detect differences between
samples) (Burton etal., 1996). This reality suggests that more
than one or two species may be necessary to determine with
certainty whether or not sediment contamination is ecologi-
cally significant.
The freshwater species that have been used successfully and
most often in assessments of sediment toxicity include both
water-column and benthic species: Selenastrurn capriconutum.
Daphnia magna, Ceriodaphnia dubia, Pimephales promelas,
H. azteca, C. tentans and C riparius, and Hexagenia limbata.
Standard guides for whole sediment toxicity testing have
also been published by the American Society for Testing and
Materials for inverteljrates, including H. azteca, C. tentans,
C. riparius, the cladocerans Daphnia magna and Ceriodaphnia
dubia, the mayfly Hexagenia, the Great Lakes amphipod
Diporeia, and the oligochaete Tubifex tubifex (ASTM,
2000a-e), Dredged material evaluations in the US follows
the protocols published by the US EPA and US Army Corps
of Engineers; in 1991 for dredged material proposed for ocean
disposal, and protocols published in 1998 for dredged
material proposed for disposal in "inland" waters of the US
(USEPA-USACE, 1998), Canada's Ontario Ministry of the
Environment has also published solid-phase test methods
(Bedard etal., 1992),
Toxicity tests for marine and estuarine species also range
from fish to bacteria. The marine and estuarine species
frequently and successfully used in studies of sediment
contamination include: the fish Cyprinodum variegatus and
Menidia beryllina, and amphipods Rhepoxynius abronius, Am-
pelisca abdita, Eohaustorius estuarius, Grandidierella japonica
and Leptocheirus plumulosus, the bivalves Haliotis rufescens,
Crassostrea
gigas and Mytilus sp, the echinoderms Strongylo-
centrottispurpuratus, Dendrasterexcentricus, and Arbaciapunc-
tulata, the polychaetes Dinophilus gyrociliatus and Neanthea
arenaceodentata, the mysid Americamysis bahia (formerly My-
sidopsis bahia), the alga Champia parvula, and the bacterial
assay Microtox using Vibrioflscheri. As with freshwater tests,
several "standard" methods for marine and estuarine organ-
isms exist. The US EPA has standardized protocols for toxicity
testing with the amphipods Rhepoxynius, Ampelisca, Eohaus-
torius, and Leptocheirus using a tO-d survival assay, a chronic
28-d survival, growth reproduction assay with Leptocheirus,
and 28-d bioaccumulation assays with Macoma and Nereis
750 TOXICITY OF SEDIMENTS
(US EPA 1994b, 2001). The European OECD (Organization
of Economic Cooperation and Development) is discussing the
possibility of developing standardized sediment tests. Environ-
ment Canada has developed methods for sediments, elutriates,
and leachate samples using Vibrio fisheri luminescence, sea
urchins and sand dollars (eehinoids), and seven amphipod
species. Methods were recently published by the Chesapeake
Bay Program for benthic testing using Lepocheirusplumulosus.
Ampelisca abdita, Lepidactylus dytiscus, and Monoculodes ed-
wardsi. Dredged materials in the US are tested using US
EPA and US Army Corps of Engineers protocols. The
American Society for Testing and Materials has standard
procedures for amphipods, fish, mysids, echinoderm, and
oyster early life stages, polychaetes, and algae (ASTM,
2000b-e).
Another component that is useful in establishing causality is
linking sediment contaminant exposure and tissue-residue
effects in organisms. Because many factors appear to alter
the bioavailability of contaminants in sediments (Hamelink
etal., 1994), approaches to establish links between the body-
residue concentrations and effects in aquatic organisms
provide the insight to better link the toxic response directly
to contaminants. Data has been amassing over the course of
the past several years that allow the direct comparison of some
residue levels with acute and chronic effects (www.wes.army.-
mil/el/ered; Jarvinen and Ankley, 1999). However, the data
base is very limited at this time, thus there is still need to
establish a weight-of-evidence approach for developing the
link between the observed response and the presence of
contaminants in sediments.
The second recent development to assist in the formation of
the link between sediment contaminants, responses in the
laboratory and observed responses in the field is that of insitu
tests.
These are tests that are conducted in the field. One
approach uses confined organisms, such as traditional
surrogates (e.g., Daphniamagtia, Ceriodaphniadubia,
H.
azteca,
C. tentans, Lumbriculus variegatus, Pimephales promelas) or
indigenous species in chambers, cages, bags, or corrals.
Organisms are placed in test chambers, which optimize
exposures to sediments, pore waters, and/or overlying waters,
and are exposed in the field from time periods ranging from
a day to months. This approach has a number of advantages,
in that it can better simulate real-world exposures, which may
fluctuate dramatically over a period of hours, reduces
sampling/experimental related artifacts, and allows for more
natural interactions of potentially critical physical and
chemical constituents than do laboratory conditions (e.g..
Chappie and Burton, 2000).
Traditionally sediment contamination was determined by
assessing the bulk chemical concentrations of individual
compounds often comparing to some background or reference
value. The desire to have chemical parameters to evaluate the
hazard of sediments has focused primarily on two approaches:
(1) a statistical approach to establish the relation-
ship between sediment contamination and toxic response;
and (2) a theoretically based approach that attempted to
account for differences in bioavailability through equilibrium
partitioning.
The original sediment quality guidelines (SQGs) that were
compared to a reference or to background provided little
insight into the impact of sediment contaminants. Therefore,
SQGs for individual chemicals were developed that relied on
paired field sediment chemistry with field or laboratory
biological effects data. They are often based on frequency
distributions and account for the impact of all chemicals
present, but do not establish cause and effect. These
approaches have also been shown to be useful and predictive
of biological effects in many (but not all) marine and
freshwater systems (e.g.. Long etal., 1998).
However, the issue of bioavailability remains (that portion
of the chemical that is available for biological uptake and
subsequent adverse effects). The equilibrium partitioning
approach attempts to address this issue specifically. This
approach suggested that interstitial water concentrations
represented the equivalent chemical activity for exposure of
aquatic organisms, and chemical concentrations could be
determined from an equilibrium partitioning calculation
(Di Toro etal., 1991). The equilibrium partitioning calculation
was particularly important for attempting to account for the
changes in bioavailability that occur with changes in amounts
of organic matter in sediments.
This method has also been applied to metals by accounting
for the interactions with acid volatile sulfide. Five metals, Cd,
Ni,
Pb, Zn, and Cu, form insoluble sulfides and so their
toxicity is limited by the amount of sulfide in the sediment.
Toxicity is observed when the amount of metal stoichiome-
trically exceeds the amount of sulfide that can bind it. A clear
demonstration ofthe utility of this approach was shown for Cd
toxicity to amphipods, Ampelisca abdita and Rhepoxynius
hudsoni in marine sediments. Like the case for organic
contaminants, the toxicity predicted from this approach has
been sometimes over predicted usually because of the
presence of other ligands that bind the metals (Ankley etal.,
1993).
When comparisons have been made between the equili-
brium-partitioning (EQP) method and other measures of
sediment quality to estimate toxic response, the sediment
quality guideline values and their predictive ability were
comparable, i.e., in greater than 70 percent of samples toxicity
or absence of toxicity were correctly predicted. However, when
comparing the predictive ability for field sites, EQP consis-
tently gave more type tl errors, false negatives, than other
sediment quality guidelines (Ingersoll etal., 1997). This likely
results from failure to account for the wide variety compounds
that exist in sediment and their interactions.
Risk-assessment for sediment-associated contaminants is a
relatively new field and is developing rapidly (Ingersoll etal.,
1997).
The strengths and limitations ofthe various assessment
methods dictates that a weight-of-evidence approach define the
significance and role of sediment-associated contaminants on
the aquatic environment and also through food web transfer
to terrestrial species. This approach should include an
integrated approach, which identifies and ranks stressors
using habitat, physical, and chemical measures, biological
community structure, toxicity, and bioaccumulation descrip-
tors.
Use of these tools can provide essential characteriza-
tions of key watershed sources, sensitive receptors, natural
variation, and both natural and anthropogenic stressors.
Only with the use of multiple tools and an understanding of
their interactions can reliable determinations of sediment
pollution and long-term consequences be made. Then cost-
effective, environmentally protective management decisions
can be made about the type, extent, and need for sediment
remediation.
G. Allen Burton, Jr. and Peter F. Landrum
TOXICITY OF SEDIMENTS 751
Bibliography
ASTM, 2000a. Test method for measuring the toxicity of sediment-
associated contaminants with freshwater invertebrates. El706-00.
In Annual Book of ASTM Standards. Volume 11.05, American
Society for Testing and Materials, West Conshohocken, PA.
ASTM, 2000b. Annual Book of ASTM Standards. Section II Water and
Environmental Technology El367. Biological Effects and Environ-
mental Fate; Biotechnology; Pesticides: Standard guide for
conducting 10-day static sediment toxicity tests with marine and
estuarine amphipods. Section 11.05 American Society for Testing
and Materials. West Conshohocken, PA.
ASTM, 2000c. Annual Bookof ASTM Standards. Section 11 Water and
Environmental Technology El
191.
Biological Effects and Environ-
mental Fate; Biotechnology; Pesticides: Standard guide for
conducting Iife-eycle toxieity tests with saltwater mysids. Section
11.05 American Society for Testing and Materials. West
Conshohoeken, PA.
ASTM, 2000d. Annual Bookof ASTM Standards. Section 11 Water and
Environmental Technology E729. Biological Effects and Environ-
mental Fate; Biotechnology; Pesticides: Standard guide for
conducting acute toxicity tests on test materials with fishes,
macroinvertebrates, and amphipods. Section 11.05 Ameriean
Society for Testing and Materials. West Conshohocken, PA.
ASTM, 2000e. Annual Bookof ASTM Standards. Section 11 Water and
Environmental Technology E1498. Biological Effects and Environ-
mental Fate; Biotechnology; Pesticides: Standard guide for
conducting sexual reproduction tests with seaweeds. 11.05
American Society for Testing Materials. West Conshohocken, PA.
Ankley, G.T., Mattson, V.R., Leonard, E.N., West, C.W., and
Bennett, J.L., 1993. Predicting the acute toxicity of copper in
freshwater sediments: evaluation of the role of acid volatile sulfide.
Environmental Toxicology and Chemistry, 12: 315-320.
Baumann, P.C, and Harshbarger, J.C:., 1995. Decline in liver
neoplasms in wild brown bullhead catfish after coking plant closes
and environmental PAHs plummet. Environmental Health
Perspec-
tives, 103: 168-170.
Bedard, D.C., Hayton, A., and Persaud, D., 1992. Ontario Ministry of
the Environment Laboratory Sediment Biological Testing Proto-
col,
Ontario Ministry of the Environment, Toronto, Ontario,
Canada, draft report.
Bishop, C.A., Mahony, N.A., Trudeau, S., and Pettit, K.E., 1999.
Reproductive success and biochemical effects of tree swallows
(Tachycineta bicolor) exposed to chlorinated hydrocarbon con-
taminants in wetlands of the Great lakes and St Lawrence River
Basin, LISA, and Canada. Environmental Toxicology and Chemistry,
18:
263-271.
Burton, G.A. Jr, 1991. Assessing freshwater sediment toxicity. Envir-
onmental
Toxicology
and Chemistry, 10: 1585-1627.
Burton, G.A. Jr, Ingersoll, C, Burnett, L., Henry, M., Hinman, M.,
IClaine, S., Landrum, P., Ross, P., and Tuchman, M., 1996. A
comparison of sediment toxicity test methods at three Great Lake
Areas of concern. Journalof Great Lakes Research, 22:
495-511.
Chapman, P.M., 1986. Sediment quality criteria from the sediment
quality triad: an example. Environmental ToxicotogyandChemistry,
5:
957-964
Chappie, D.J., and Burton, G.A. Jr, 2000. Applications of aquatic and
sediment toxieity testing in situ. Journal of Soil and Sediment Con-
tamination, 9: 219-246.
Di Toro, D.M., Zarba, C.S., Hansen, D.J., Berry, W.J., Swartz, R.C.,
Cowen, C.E., Pavlou, S.P., Allen, H.E., Thomas, N.A., and
Paquin, P.R., 1991. Technical basis for establishing sediment
quality criteria for nonionic organic chemicals by using equilibrium
partitioning. Environmental Toxicology and Chemistry, 12:
1541-1583.
Hamelink, J.L., Landrum, P.F., Bergman, H.L., and Benson, W.H.,
1994.
Bioavailahility:
Physical.
Chemical, and Biological Interactions.
Boca Raton, FL: Lewis Publishers.
Harshbarger, J.C, and Clark, J.B., 1990. Epizootiology of neoplasms
in bony fish of North America. Science oftheTotal Environment, 94:
1-32.
Ingersoll, C.G., Dillon, T., and Biddinger, G.R., 1997. Ecological Risk
Assessment of Contaminated Sediments. Pensacola, FL: Society of
Environmental Toxicology and Chemistry Press.
Jarvinen, A.W., and Ankley, G.T., 1999. LinkageofEffectstoTissuere-
sidues: Development ofa Comprehensive Datahasefor Aquatic Organ-
isms Exposed to Inorganic and Organic Chemicals. Penseola, FL:
Society of Environmental Toxicology and Chemistry Press.
Lester, D.C, and Mclntosh, A., 1994. Accumulation of polychlori-
nated biphenyl congeners from Lake Champlain sediments by My-
sisrelicta. Environmental Toxicology and Chemistry, 13: 1825-1841.
Long, E.R., Field, L.J., and MacDonald, D.D., 1998. Predicting
toxicity in marine sediments with numerical sediment quality
guidelines. Environmental
Toxicology
and Chemistry, 17: 714-727.
Ludwig, J.P., Auman, H.J., Kurita, H., Ludwig, M.E.,
Campbell, L.M., Giesy, J.P., Tillitt, D., Jones, P., Yamashita, N.,
Tanabe, S., and Tatsukawa, R., 1993. Caspian tern reproduction in
the saginaw bay ecosystem following a 100-year flood event. Jour-
nal of
Great
Lakes Research, 96-108.
Maruya, K.A., and Lee, R.F., 1998. Biota-sediment accumulation and
trophic transfer faetors for extremely hydrophobic polychlorinated
biphenyls. Environmental
Toxicology
and Chemistry, 17: 2463-2469.
McCarty, J.P., and Secord, A.L., 1999. Reproductive ecology of tree
swallows {Tachycineta hicolor) with high levels of polychlorinated
biphenyl contamination. Environmental
Toxicology
and Chemistry,
18:
1433-1439.
Morrison, H.A., Gobas, F.A.P.C, Lazar, R., and Haffner, G.D.,
1996.
Development and verification of a bioaecumulation modei
for organic contaminants in benthic invertebrates. Environmental
Toxicology
and
Technology,
30: 3377-3384.
Pelley, J., 1999. North Carolina considers controls to protect
contaminated waters. Environmental Toxicology and Technology,
33:
lOA.
Peterson, C.H., Kennicutt II, M.C, Green, R.H., Montagna, P.,
Harper, D.E. Jr., Powell, E.N., and Roscigno, P.F., 1996.
Ecological consequences of environmental perturbations associated
with offshore hydrocarbon production: a perspeetive on long-term
exposures in the Gulf of Mexico. Canadian Journalof Fisheries and
Aquatic Sciences, 53: 2637-2654.
Swartz, R.C, Cole, F.A., Lamberson, J.O., Ferraro, S.P.,
Sehults, D.W., DeBen, W.A., Lee II, H., and Ozretieh, R.J.,
1994.
Sediment toxicity, contamination, and amphipod abundance
at a DDT- and Dieldrin-contaminated site in San Franeisco Bay.
Environmental
Toxicology
and Chemistrv, 13: 949-962.
Thomann, R.V., Connolly, J.P., and Parkerton, T.F., 1992. An
equilibrium model of organic chemical accumulation in aquatic
food webs with sediment interaction. Environmental Toxicology and
Chemistry, 11: 615-629.
US Environmental Protection Agency, 1994a. Methods for measuring
the toxicity and Bioaceumulation of Sediment-associated Con-
taminants with Freshwater Invertebrates. Office of Research and
Development. EPA/600/R-94/024. Washington, DC
US Environmental Protection Agency, 1994b. Methods for assessing
the Toxicity of sediment-associated contaminants with estuarine
and marine amphipods. £'/M/600/R-94/025, Office of Research and
Development, Washington, DC
US Environmental Protection Agency, 1997. The Incidence and Sever-
ity of Sediment Contamination in Surface
Waters
of the United States.
Volumes 1-3. EPA 823-R-97-006. Scienee and Teehnology Office.
Washington, DC.
US Environmental Protection Agency, 2000. Methods for assessing
the toxicity and bioaccumulation of sediment-associated contami-
nants with freshwater invertebrates. Second edition, EPA/600/R-
99/064, Duluth, MN.
US Environmental Protection Agency, 2001. Method for assessing the
chronic toxicity of marine and estuarine sediment-associated
contaminants with the amphipod Leptocheirus plumulosus, 1st
edn, EPA/600/R-01/020. Washington, DC.
USEPA-USACE, 1998. Evaluation of dredged material for
discharge in waters of the US testing manual. EPA-823-B-98-
005.
Washington, DC.
752
TRACERS FOR SEDIMENT MOVEMENT
TRACERS FOR SEDIMENT MOVEMENT
For the purposes of studying sediment dynamics, tracers are
defined here as marked particles introduced into a river,
estuary, or coastal system, to obtain general information on
the characteristics of sediment movement within such environ-
ments. Tracers provide a relatively simple means of over-
coming technical and sampling problems without the need for
a detailed kinematic study of the sedimentary regime
(Crickmore etal., 1990). The type of tracer to be used will
largely depend on the objectives, environment characteristics,
and intended observation period of the experiments (Hassan
and Ergenzinger, in press). In addition, one must consider each
method and materials, the sample size and expected recovery
rate,
detection, data collection and analyzes.
Both artificial and natural tracers are available for tracking
coarse and fine material, tn selecting an artificial tracer, the
geochemical and physical properties, and hydrodynamic
conditions must be considered. Some tracing techniques can
be used for fine and coarse sediment, though tagging silt and
clay sizes is complex. With fine material, physical character-
istics make tracing and detection difficult for several reasons:
(i) the fine sediment is usually carried in suspension, and its
movement dispersion over large areas is rapid; (ii) coating or
attaching tags onto the fine sediment may change the physical
characteristics of the particle; and (iii) the cohesive nature of
clay and silt causes fiocculation, and labeling may lead to
changes in the chemical properties of the sediment.
A variety of detection methods have been tried, ranging
from simple visual identification of the tracers, to magnetic or
metal detection, radioactive detection, neutron activation, and
radio wave detection. Some detection methods, using paint,
magnets, and radio transmitters, permit individual identifica-
tion of tracer particles whilst others such as radioactive
material do not. The approach will clearly influence how much
information can be gathered, and hence subsequent data
analyzes and interpretations. Detection strategies include:
individual location of particles; fixed automatic detection of
particle passage; periodic dynamic detection and automatic
dynamic detection (Hassan and Ergenzinger, in press).
A brief background and explanation is provided below on
the most commonly-used tracers.
Painted particles are probably the simplest method of tracing
gravels. Paint is easy to apply and can be detected visually, but
field detection is limited to the surface. In the fluorescent
method a special dye is applied to a set amount of sediment.
Upon light stimulation, the treated sediment emits light of a
wavelength characteristic of the dye (Crickmore etal., 1990).
Samples can then be taken within the study area to assess the
dispersion of sediment. The fluorescent method is suitable for
sand tracing, and even larger material. Suitable fluorescent
dyes include rhodamine, auramine, eosine, primulin, fluor-
escein, and anthracine.
The
radioactive
technique is very versatile and can be used in
both fine and coarse sediment (Crickmore et al., 1990). It
allows in situ continuous detection of exposed and buried
particles and, hence, permits immediate data processing and
rapid evaluation of the tracking strategy. The presence of the
radioaetively labeled sediment is detected from the radiation
emitted by the tag (e.g., Crickmore etal., 1990). Wide ranges of
radioisotopes provide a choice of tracers with a half-life to
match most objectives. According to the method of labeling,
the radioactive tracers can be grouped into three categories:
manufacture of radioactive grains of the same density and
shape as the bed material, application of the isotope to the
outside or within the natural grains, and introduction of
natural radioactive minerals. The most commonly used
isotopes are '""La (half-life of 40.2 hours), ^'Cr (half-life of
27.8 days), "^Sc (half-life of 83.9 days), and "°Ag (half-life of
253 days). The method is toxic and dangerous to the
environment and therefore it is very difficult to obtain
permission to use.
In the neutron activation
method,
a chemical element is fixed
to the sediment and its concentration is measured after
activation in a nuclear reactor. The sediment is labeled by an
element with a high thermal neutron which, after irradiation, is
detectable above the natural radioactive background. Studies
have used, among other elements, cobalt, tantalum, iridium,
rhenium, and gold (e.g., Crickmore etal., 1990).
Ihs fingerprinting teehnique can be used to define sediment
sources and transfer within a drainage basin, and sedimenta-
tion rates in lakes and reservoirs. This method is based on (i)
the selection of physical and/or chemical properties that
distinguish a particular source material, and (ii) the compar-
ison of measurements of the same properties of the sediment
with the equivalent values obtained for the source material
(e.g., Oldfield etal., 1979; Walling and Woodward, 1992).
Fallout
radionuclides have been used to pursue a wide range
of studies within drainage basins (e.g.. Walling etal., 1999). On
reaching the earth's surface by either wet or dry faJlout,
radionuclides are adsorbed onto the soil and become useful for
tracing the movement of fine sediment. Applications include
the study of suspended sediment sources, sedimentation rates
and sediment yield, gully erosion, floodplain sedimentation,
sediment budget, residence time of fine sediment, redistribu-
tion of soil, and lake and reservoir sedimentation. Three
fallout radionuclides have primarily been used: '^^Cs, un-
supported ^'"Pb, and ''Be, of which Cs is the most favored.
Caesium-137 is an artificial radionuclide produced as a result
of global fallout from the testing of nuclear weapons. With a
half-life of 30.1 years, "'Cs provides an average value of
surface erosion for a period of about 45 years. It has been
found that eroded soils lose "^Cs in proportion to the severity
of erosion. At erosion sites '^'Cs will be present in quantities
smaller than the reference value, whereas at deposition sites the
opposite is observed. Both ^'"Pb and ^Be are natural radio-
nuclides that reach the earth's surface via atmospheric fallout
at a constant rate through time. Because of a relatively short
half-life, the ^Be technique is useful for short-term measure-
ments that can be related to specific events.
Eerruginous tracers include metal strips, acid coating,
artificial magnets, and mineral magnetite (Hassan and
Ergenzinger, in press). The underlying principle is that when
a tracer passes over an iron-cored coil of wire, a measurable
electrical pulse is generated. Magnet or metal locators are used
to find tagged particles, and automatic detection systems are
also available. The latter type of system can provide in situ
continuous measurements of sediment movement to record an
unlimited number of clasts. Due to the high recovery rate,
magnetic methods are the most widely used technique. The
method permits the location of buried particles, as well as
those at the surface. Both natural and artificial magnetic
tracers have been used. If no naturally magnetic material is
TUFAS AND TRAVERTINES
753
present and suitable for use where the research is to be carried
out, then it is necessary to use artificially magnetized tracers.
For tracing gravel larger than
16
mm, a magnet can be inserted
into pre-drilled holes. Manufactured synthetic magnetic
clasts,
made from resin or concrete with a magnetic core in
the center, have been used for tagging small and large particles
(>8 mm). This approach allows the manufacture of clasts of
different shapes and densities. Magnetic enhancement is
another method, which is based on the enhancement of
natural magnetism by strong heating of naturally iron-rich
pebbles. Through the heat treatment the particle mineralogy is
altered and the magnetism is enhanced up to 300 times its
original strength to a level that can be detected and
distinguished from the bed material. Magnetic properties of
the soil can be used to trace sediment sources and fine sediment
transport in alluvial streams (e.g., Oldfield etal., 1979). The
upper part of the soil profile has a higher concentration of
magnetic content than the lower part or the parent material
due to pedogenic processes and burning. The enhanced
magnetism (see artificial magnetic enhancement, above) helps
to differentiate between the surface and subsurface soil and
hence the suspended sediment sources.
The Radio transmission method permits the active detection
of tracers in cobble size material (Schmidt and Ergenzinger,
1992).
A transmitter capsule, about 65 mm long and 20 mm in
diameter and with a battery life expectancy of about three
months, is inserted into the clast. Within a set of tracers, the
stones will each emit slightly different frequencies so that they
can be individually identified. The detection system consists of
a transmitter, an antenna, receiver, and a data logger.
Marwan A. Hassan
Bibliography
Crickmore, M.J., Tazioli, G.S., Appleby, P.G., and Oldfield, F., 1990.
The
Use
of Nuclear
Techniques in
Sediment
Transport
and Sedimenta-
tion Problems. Technical Documents in Hydrology, IHP-III Project
5.2, Paris: UNESCO, 170pp.
Hassan, M.A., and Ergenzinger, P., in press. Use of tracers in fluvial
geomorphology. In
Kondolf,
G.M., and Piegay, H. (eds.). Methods
in
Fluvial Morphology. Wiley and Sons. Expected to be published in
April
2003.
Ingle, J.C Jr, 1966. The Movement of Beach
Sand.
Elsevier.
Oldfield, F., Rummery, T.A., Thompson, R., and Walling, D.E., 1979.
Identification of suspended sediment sources by means of magnetic
measurements: some preliminary results.
Water
Resources Research,
15:211-217.
Schmidt, K.-H., and Ergenzinger, P., 1992. Bedload entrainment,
travel lengths, step lengths, rest periods—studies with passive (iron,
magnetic) and active (radio) tracer techniques. Earth Surface
Processes
and Landforms, 17: 147-165.
Walling, D.E., He, Q., and Blake, W., 1999. Use of 'Be and '"Cs
measurements to document short and medium term rates of water
induced soil erosion on agricultural lands. Water Resources
Research, 35: 3865-3874.
Walling, D.E., and Woodward, J.C, 1992. Use of radiometrie
fingerprints to derive information on suspended sediment sources.
In Bogen, J., and Walling, D.E. (eds.). Erosion and Sediment Mon-
itoring Programmes in River Basins. IAHS Publication Number 210,
pp.
153-164.
Cross-references
Attrition (Abrasion), Fluvial
Attrition (Abrasion), Marine
Coastal Sedimentary Facies
Deltas and Estuaries
Grain Size and Shape
Rivers and Alluvial Fans
Sediment Transport by Unidirectional Water Flows
Sediment Transport by Tides
Sediment Transport by Waves
TUFAS AND TRAVERTINES
Tufas and travertines are rare in Paleozoic and Mesozoie
successions, but become frequent in the Cenozoic. Never-
theless, the majority of recorded examples are Quaternary and
Holocene (Waring, 1965; Pentecost, 1995; Ford and Pedley,
1996).
The term "tufa" is recommended for the description of
all ambient temperature freshwater carbonate deposits with
"travertine", an alternative name, being reserved for all
hydrothermal freshwater deposits (Pedley, 1990). In contrast,
many employ the term "travertine" for all types with
distinction into high temperature "thermogene" and ambient
temperature "metogene" deposits (Pentecost, and Viles, 1994).
Tufas commonly contain abundant evidence of mosses and
higher plants, whereas thermal sites encourage physico-
chemical precipitates and colonization is restricted to mi-
crobes. Travertine (hydrothermal) deposits frequently contain
unusual geochemical signatures (e.g., high sulfur and carbon
dioxide components). Many deposits have been dated using
radiocarbon methods (Srdoc etal., 1986), and uranium series
isotopes (Hennig etal., 1983). Stable isotopes of oxygen and
carbon have proved valuable as proxy-environmental indica-
tors (Andrews etal., 2000).
Speleothems (^.
v.)
are similar to travertines but are physico-
chemical cave and cavity precipitates.
Tufas
Tufas typically are rich in plants, but may also contain insects,
pollen, molluscs, and ostraeods (Preeee, 1991; Vadour, 1994;
Taylor et al., 1998). Mosses, liverworts, and semi-aquatics
commonly provide a framework upon which fringes of
low-magnesian calcite cements precipitate. Precipitation is
partly controlled physico-chemically by CO2 degassing
whicli renders the dissolved calcium carbonate less soluble
(Lorah and Herman, 1988). However, there is also an
important biomediated contribution triggered as a by-product
of photosynthesis and associated metabolic processes
connected with diatoms, cyanobaeteria, and heterotropic
bacteria. These microorganisms are ever present within
surficial (procaryote-microphyte) biofilms and many tufas
and travertines are commonly entrapped within the associated
fringe cements. Lithification is virtually instantaneous in many
fluvial, waterfall, and lake margin settings and forms
laminated, freshwater plant-dominated framestone reefs or
"phytoherms" (Buccino etal., 1978; Pedley, 1990). Oncoids are
common and are often associated with allochthonous tufa
debris (phytoclast tufa) in channels and point bars. Fine lime
mud-grade tufa formed in lakes is derived either from marginal
stromatolite microherms and phytoherms or is a by-product of
754
TUFAS AND TRAVERTINES
microbial ("picoplankton") metabolism within the water
column (Thompson, 2000).
Tufa classifications are diverse (e.g., Chafetz and Folk,
1984;
Ordonez and Garcia del Cura, 1983; Pedley, 1990).
Confusion is reduced if studies are based on facies associations
(Buccino et al., 1978; Pentecost and Viles, 1994; Ford and
Pedley, 1996). These demonstrate that lithofacies complexity is
a function of slope angle, irregularities in substrate, and
velocity and flow-frequency of carbonate-enriched water
supply. Although there is a facies continuum, probably four
common models can be derived from the spectrum.
Fluvial tufa model
This may be flow dominated (braided fluvial tufas) and
dominated by oncoid and phytoclast channel fills. Small
phytoherms and microherms (especially stromatolites) may
occur on the stream bed or along the channel margins.
Alternatively, gorges may be dominated by several cycles of
fluvial barrage tufas (Golubic, 1969). Barrages are narrow,
downstream arcuate dams of phytoherm framestone that create
lakes.
The pool depocenters commonly contain thick lime mud
and organic-rich deposits (Pedley etal., 2000). The contained
pollen assemblages are useful proxy-climatic indicators.
Perched springline tufa model
This consists of a multilobate mound that develops in close
proximity of valley side resurgences. Typically, such mounds
have a virtually flat top and steep distal face, the latter often
associated with cascades and trailing mosses (Violante etal.,
1994).
Detrital (phytoclastic) deposits, developed distal to the
lobe,
typically contain land snails and paleosols.
Lacustrine tufa model
This model develops in association with large standing water
bodies. Their margins are sites of "bacterioherms" which
develop characteristic flat-tops controlled by the position of
the air/water interface. In the shallow marginal waters oncoids
occur and Chara stands are common. Lime silts and muds
dominate the deepest waters.
Paludal tufa model
Poorly drained slopes colonized by hydrophytic vegetation
(especially moss hummocks and grass cushions) are sites of
lime mud precipitation (e.g., Belgian "Crons"). More extensive
tufa sheets are developed in valley bottom situations and may
be associated with shallow ephemeral pools and ill-defined
watercourses. Small tufa mounds may form where lime-rich
springs issue into otherwise organic dominated ponds.
Frequent desiccation encourages paleosol development
throughout many deposits.
Saline tufas
Tufa variants which do not fit into the four models are the
"mound tufas" and "towers" which develop within saline
lakes (e.g.. Pyramid, USA, Benton, 1995; also in China Arp
et al., 1998). They consist of bacterial microherms. They
develop at freshwater resurgence points on the lake bed where
microbial colonization is encouraged by local salinity dilution.
Marginal microbial phytoherms can also occur.
Travertines
Waring (1965) lists several hundred thermal (travertine-
forming) springs. Many deposits are physico-chemical calcium
carbonate precipitates from simple degassing and cooling of
lime-rich waters close to thermal springs. Cyanobacteria live in
the near boiling water and trigger further calcium carbonate
precipitation. However, toxic sulfides can inhibit biofilm
colonization. Temperature and toxicity generally decline
down-flow. Consequently, higher plants are able to colonize
these cooler water (tufa) domains. The two commonest
associations are:
Fissure ridges
In tectonically active areas calcium-rich thermal waters can
rise along fault-lines and crystalline travertine are deposited
to form elongate mounds with cleft-like median valleys.
Pinacles and mounds can develop from the same process
where the water is point sourced (e.g.. Mammoth Hot Springs,
Wyoming). The deposits are often composed of thin laminae,
which dip steeply away from the fissure.
Terraces
Alternatively, lobate, planar sheets can develop into an
intricate series of microterraces and gutters associated
descending flights of small rimstone pools (e.g., Guo and
Riding, 1998).
Diagenesis
The vast majority of tufas and travertines are composed of
low magnesian calcite. They are prone to early "spar-
micritization" by microborers, however, many retain sharp
internal fabric boundaries. The addition of physico-chemical
cements further reinforces older deposits. Metastable carbo-
nates (e.g., aragonite in some travertines) soon neomorphose
to calcite spar. Selective dissolution within tufas and traver-
tines appears common. However, many voids are primary
cavities lined by later cements.
Martyn H. Pedley
Bibliography
Andrews, J.E., Pedley, H.M., and Dennis, P., 2000. Palaeoenviron-
mental records in Holocene Spanish tufas: a stable isotope
approach in search of reliable climatic archives. Sedimentoiogv,
47:
961-978.
Arp,
G., Hofmann, J., and Reitner, J., 1998. Microbial fabric
formation in spring mounds ("Microbialites") of alkaline salt
lakes in the Japan Sand Sea, PR China. Paiaios, 13(6): 581-592.
Benson, L., 1995. Carbonate deposition. Pyramid Lake subbasin,
Nevada:2. Lake levels and polar Jetstream positions reconstructed
from radiocarbon ages and elevations of carbonates (tufas)
deposited in the Lahontal basin. Palaeogeography Palaeoclimatol-
ogy Paiaeoecology, UT: 1-30.
Buccino, G., D'Argenio, B., Ferreri, V., Brancaccio, L., Panichi, C,
and Stanzione, D., 1978. I travertini della bassa Valle del Tanagrio
(Campania): studio geomorphologico, sedimentologico e geochi-
mico.
BoUitino
Societa
Geologica
Italia, 97: 617-646.
Chafetz, H.S., and Folk, R.L., 1984. Travertines: depositional
morphology and the baeterially-constructed constituents. Journal
of Sedimentary Petrology, 54: 289-316.