Ecological and Toxicological Effects of Suspended and Bedded Sediments 

on Aquatic Habitats

A Concise Review for Developing Water Quality Criteria for Suspended and 

Bedded Sediments (SABS) 

By

Mitra N. Jha, Ph. D.

Bill Swietlik, Ph. D. 

U.S. EPA

Health and Ecological Criteria Division

Office of Science and Technology

Office of Water
2

Table of Contents:

Page

Executive Summary---------------------------------------------------------------------3

Introduction------------------------------------------------------------------------------5

Effects on Aquatic Habitats------------------------------------------------------------5

Effects on Fish Population-------------------------------------------------------------7

Physiological and Behavioral Effects---------------------------------------7

Suspended Sediment Effects------------------------------------------------11

Deposited Sediment Effects-------------------------------------------------14

Effects on Invertebrate----------------------------------------------------------------17

Effects on Mussels---------------------------------------------------------------------21

Effects on Coral------------------------------------------------------------------------23

Effects on Wildlife---------------------------------------------------------------------24

Effects on Aquatic Plants-------------------------------------------------------------25

Other Effects----------------------------------------------------------------------------28

Conclusion------------------------------------------------------------------------------30

References-------------------------------------------------------------------------------31
3

Executive Summary:

The issue of Suspended and Bedded Sediments (SABS) in aquatic environments has been

a topic of concern for decades. SABS are the particulate organic and inorganic matter 

suspended in or are carried by the water and/or accumulate in a 

loose, unconsolidated form on the bottom of water bodies (EPA, 2002). It is a natural 

component of ecosystems and cannot be categorized in the same group as heavy metals, 

volatile organic compounds, or other toxins. The entry of SABS to water bodies from 

upland sources is by both natural and artificial (mechanical) process. Excess SABS in 

rivers and streams has been identified as the largest and most pervasive water pollution 

problem faced by aquatic systems in North America. SABS are the leading cause of 

impairment of water bodies such as rivers, streams, lakes, reservoirs, ponds, and estuaries 

(303 (d) listed sites) in the United States. Anthropogenic inputs of SABS into aquatic 

environments can cause stress to systems and directly and indirectly impact upon the 

behavior and health of aquatic biota. It has been reported that in 1998, approximately 

40% of assessed river miles in the U.S. had problems arising from SABS stress 

(U.S.EPA, 2000a). During the 1980s the number of North American fish species 

considered by the American Fisheries Society to be of special concern, threatened, or 

endangered . Among North American freshwater species, approximately 37% of fishes, 

35% of amphibians, 73% of mussels, and 65% of crayfish have been classified as 

vulnerable, imperiled, or endangered. This review included the effects of SABS 

on fish, invertebrates, mussels, corals, wildlife, microbes, and aquatic plants.

SABS can affect fish populations indirectly by altering the available food supply, or more

directly. Three major direct effects of SABS on fishes include: 

1) physiological and behavioral effects such as suffocation; 

2) bioenergentic effects due to increase in suspended sediments; and 

3) effects due to sediment deposition, leading to increased embeddedness or burial of 

eggs and larvae. 

Physiological responses include but are not limited to impaired growth, histological 

changes to gill tissue, alterations in blood chemistry, and an overall decrease in health 

and resistance to parasitism and disease. Elevated 

levels of sediment may have an adverse effect on fish health, metabolic demand, 

and growth rate of various type of fish. Typical behavior responses include an increased 

frequency of the cough reflex, avoidance of suspended sediments, reduction in feeding 

and temporary disruption of territoriality. The severity of the behavioral response is 

associated with the timing of disturbance, the level of stress, decreased energy reserves, 

phagocytes, metabolic depletion, seasonal variation, and alteration of the habitat. 

Severity of effect caused by suspended sediments on fish population is a function of 

many factors, which, in addition to sediment concentration, duration, particle size, and 

life history stage, may include temperature, physical and chemical characteristics of the 

particles, associated toxicants, acclimatization, other stressors and interactions among 

factors. Reproductive success is severely affected by sediment deposition. Effects include 

egg survival rates, egg death, increased embryo mortality, reduced hatching, and reduced

larval survival These impacts are associated with amount of elevated sediment loads, size 
4

of the sediment particles involved, seasonal variation, rates of sediment deposition and 

reduction in oxygen supply.

There are multiple ecological effects of SABS on aquatic invertebrates including 

alterations in community structure, diversity, density, biomass, growth, rates 

of reproduction, and mortality. Direct impact on the invertebrate are due to abrasion, 

clogging of filtration mechanisms, and in extreme cases smothering and burial resulting 

in mortality. Indirect impact are primarily from light attenuation leading to changes in 

feeding efficiency, and behavior (e.g., drift ), changes in substrate composition. 

Significant relationships between benthic invertebrate communities and sediment 

deposition in streams have been reported. Increased levels of SABS reduce ingestion 

rates of freshwater mussels and reduce their survival . Even low levels of sediment may

reduce feeding and, in turn, have sublethal effects on growth and reproduction. Excessive 

sedimentation can adversely affect the structure and function of the coral reef ecosystem 

by altering physical and biological processes through a variety of mechanisms including 

reduced growth rates, temporary bleaching , and interfering with the settlement of larvae .

Increased concentrations of SABS and increased rates of sedimentation have direct 

impact on aquatic plants and can reduce the primary productivity (photosynthetic 

process) of the impacted area due to increased turbidity. SABS-caused reductions in 

plant species density, biomass, and diversity throughout a trophic level are translated into 

reductions in energy input to the next trophic level. Decreases in plant populations may 

result in decreased populations of zooplankton, macroinvertebrates and higher trophic 

levels Increased SABS concentration and deposition may bury microbes and organic 

matter needed for microbial processes, or alter redox profiles important to biochemical

processes involved in water quality. 

In conclusion, the literature strongly suggests that SABS are significant contributors to a

decline in populations of North American aquatic organisms. Elevated SABS loads 

can alter stream bed composition, change channel geometry, reduce the depth of pools 

and produce a net reduction in riffle areas, reduce available habitat and thereby can cause 

a decrease in the fish or other aquatic organism holding capacity of a stream. 

Impacts to lotic fauna may be expressed through pervasive alterations in local food 

chains beginning at the primary trophic level. Decreases in primary production are 

associated with increases in SABS concentration which produce negative cascading 

effects through depleted food availability to zooplankton, insects, freshwater mollusks, 

and fish 
5

Introduction: 

The issue of sedimentation in aquatic environments has been a topic of concern for 

decades. Most sediment is caused by erosion, which became a national issue because of 

the dust bowl and as a result the Soil Erosion Service was established in the mid 1930s. 

Since then, the study of erosion, transport, and deposition of sediment in streams has 

increased exponentially in the United States. The delivery of sediments to water bodies 

from upland sources is by both natural and artificial (anthropogenic) process. When the 

rate of sedimentation from various sources to the water bodies exceeds their capacity to 

transport or assimilate the sediment, an ecological imbalance may occur. Thus it causes a 

stress to the aquatic community and as a result, suitability and/or structure of aquatic

habitats may be changed. This project report provides a collection of information related 

to ecological and toxicological effects to sediment synthesized from the literature. 

Considering all of the past and current research, the present paper focuses on the various 

aspects of sedimentation effects on aquatic habitat to develop the water quality criteria 

for sedimentation. In this report the term “Sediment” has been used to evaluate the 

“Ecological and toxicological effects of sediment on aquatic habitat”. The term 

“Sediment” in the present report is synonymous to the term “Suspended and Bedded 

Sediments (SABS)” (used by EPA’s Office of Water, Office of Science Technology, 

Health and Ecological Criteria Division). Sediment or SABS have been defined as the 

particulate organic and inorganic matter suspended in or deposited on the bottom of 

water bodies (EPA). 

Sediment is not necessarily considered a pollutant because it is indigenous to water 

bodies. Excess sediment is, however, considered pollution, Sedimentation is a unique 

stressor affecting vital components of aquatic ecosystems. It is a natural component of 

ecosystems and cannot be categorized in the same group as heavy metals, volatile organic 

compounds, or other toxins. Excess sediment is a physical stressor that has severe effects 

on aquatic habitats. Sweeten (1995) reported that excess sediment in rivers and streams 

has been identified as the largest and most pervasive water pollution problem faced by 

aquatic systems in North America. In this report sedimentation includes clean sediment, 

suspended sediment concentration (SSC), bed-load, turbidity, dirt, total suspended solids 

(TSS), eroded soil particles including silt, sand, and gravel, and other eroded materials. In 

the literature, sedimentation of substrata has been measured in several ways, including (1) 

the area of streambed covered, (2) depth of coverage, (3) the particle size distribution of 

sediment covering the substrata, (4) weight/volume units (5) turbidity measurement and 

(6) percentage of interstitial spaces filled.

Effects On Aquatic Habitats:

Kuhnle et al. (2001) reported Sediment is the leading cause of impairment of water 

bodies such as rivers, streams, lakes, reservoirs, ponds, and estuaries (303 (d) listed sites) 

in the United States. Turbidity, suspended solids, sediment, and siltation have been 

consistently listed in 305(b) Water Quality Reports in rivers and streams, lakes, 

reservoirs, and ponds, wetlands, and ocean shoreline waters. Karr et al. (1986) estimated 

that 50% of streams and rivers in U.S. are impaired by habitat degradation including 
6

sedimentation. Among North American freshwater species, approximately 37% of fishes, 

35% of amphibians, 73% of mussels, and 65% of crayfish have been classified as 

vulnerable, imperiled, or endangered (Williams et al., 1993; Richter et al., 1997). It has 

been reported that drastic declines have occurred in all major aquatic fauna groups and a 

decline of fish population appears to continue in the country. During the 1980s, the 

number of North American fish species considered by the American Fisheries Society to 

be of special concern, threatened, or endangered increased by 30% (Williams et al 1989). 

For this report, the sediment has been considered as clean sediment (hypothetically 

devoid of other pollutants), which is the stressor of concern. Other pollutants may indeed 

contribute to ecological problems but sediment is the leading problem and is therefore the 

problem we need to address. According to the USEPA (2002), agriculture, construction 

sites, forest industry, roads and urban development all contribute to in-stream 

sedimentation. Vegetation removal and soil compaction resulting from these land uses 

can cause hydrological changes, including alterations in infiltration rates, runoff, and 

stream base-flow. 

Anderson et al. (1996) reported that anthropogenic inputs of sediments into aquatic 

environments can cause stress to systems and thereby, directly and indirectly impact upon 

behavior and health of aquatic biota. Stress due to sedimentation happens because the 

homeostatic or stabilizing processes of the aquatic organism are challenged beyond the 

capabilities of the organism to compensate for the biotic or abiotic disturbance. Increased 

concentrations of suspended sediments can have direct effects on behavior, physiology 

and populations of aquatic biota. Berry et al (2002) suggested that imbalance in loading 

of sediments to aquatic systems is one of the greatest causes of water quality impairment 

in the nation. In 1998, approximately 40% of assessed river miles in the U.S. had 

problems arising from sediment stress (U.S.EPA, 2000a). Griffiths and Walton (1978) 

presented a review of the effects of sedimentation on aquatic biota. The detrimental 

effects of increased suspended and settled sediments on fish, bottom invertebrates, and 

primary productivity are documented in the report. It is shown that the upper tolerance 

level for suspended sediment is between 80-100 mg/l for fish, and as low as 10-15 mg/l 

for bottom invertebrates. Recovery of the aquatic biota from increased sedimentation is 

dependent on the severity of sediment additions and the discharge level of the rivers or 

streams. Recovery from short-term additions of sediment is usually complete within one 

year. 

Sediment movement is part of the natural function of freshwater ecosystems and many 

aquatic species have adapted to life in running water by developing highly complex 

methods for securing themselves to the stubstratum.. Many aquatic species also utilize 

spaces within the bedded sediments. The SABS forms a matrix of particles with waterfilled spaces in between. Some aquatic species spend most of their life in these spaces 

within the bedded sediments, others use it as juveniles or to escape predators, or during 

floods to avoid being washed away. However, increases in sediment input over natural 

levels impact the aquatic habitats and is probably the most serious and widespread issue 

for rivers and streams. When human activity significantly increases the amount of 

sediment, or changes stream flows, it can have significant effects on aquatic flora and 
7

fauna. In the present report the literature review examined the effects of increased 

sediment load and sedimentation on aquatic ecosystems. The review included the effects 

of SABS on fish, invertebrates, mussels, corals, wildlife, microbes, and aquatic plants. 

Attempts have been made to relate the concentration and duration of specific sediment 

exposure events to the alteration of aquatic habitats. 

Effects on Fish Population:

Sedimentation can have a significant impact on fish population. Anderson (1996) 

reported that sedimentation can affect fish populations by altering the available food 

supply. According to Berry et al (2002) the conventional wisdom is that both the degree 

of exposure (measured as total suspended solids (TSS) and turbidity) and the duration of 

the exposure impact the fish population. Newcombe and MacDonald (1991) suggested 

that the first, mild, primarily behavioral effects would be seen with low intensity, shortterm exposures. As the duration of exposure and intensity of exposure increase, sublethal 

effects are manifested, and lethal effects begin to show up at more intense exposures of 

longer duration. There are three major types of direct effects of sediment on fishes 

1) Physiological and behavioral effects of sediment, such as suffocation; 

2) Effects due to increase in suspended solids; and 

3) Effects due to sediment deposition, leading to increased embeddedness or burial of 

eggs and larvae. 

Physiological and Behavioral effects of sediment:

In the present report, emphasis has been given to the physiological effects on fish in 

response to the increased stress of sediments. Physiological responses include but are not 

limited to impaired growth, histological changes to gill tissue, alterations in blood 

chemistry, and an overall decrease in health and resistance to parasitism and disease. The 

effects are usually a graded response to increasing sediment dose. Lower dose and 

shorter exposures produce transitory physiological and behavioral effects, while higher 

doses for longer period can be more lasting and severe. Anderson (1996) suggested that 

longer exposure periods and/or higher concentrations of sediments are required before 

physiological responses are expressed. In this respect (high sediment concentration and 

longer exposure), physiological responses are more of a chronic effect.

Growth rates of various type of fish are impacted by increased sediment concentration. 

Anderson (1996) reported that growth is a sensitive physiological response and shorter 

exposure periods of sediment may elicit the responses. Buck (1956), Sigler et al. (1984), 

and McLeay et al. (1987) observed that growth impairment time ranged from a low of 

two weeks for juvenile steelhead trout and coho salmon to a high of six weeks between 

84 and 120 mg/L concentration for young-of-the-year arctic grayling. Anderson (1996) 

stated that growth impairment resulting from increased sediment stress is related to 

metabolic demand rather than from a reduction in feeding. Sutherland (2002) observed

that mean daily growth rates and mean cortisol levels of juvenile C. galactura differ 

significantly among suspended sediment concentration (SSC) treatments, although 
8

responses to the three lowest concentrations were not significantly different. Jowett 

(2001) demonstrated overall effect of high levels of sedimentation on bullies 

(Gobiomorphus breviceps) density was similar to that of providing cover, with the effect 

diminishing as the level of sedimentation decreased. However, Sigler et al. (1984) and 

McLeay et al. (1987) suggested that growth impairment is related more to the metabolic 

demands resulting from stress caused by increased suspended sediment than from a 

reduction in feeding. 

Anderson (1996) reported that increased sediment concentration cause the alteration in 

blood chemistry of fishes resulting from the increased stress. Newcombe (1994) 

observed some other physiological changes such as release of stress hormones (i.e., 

cortisol and epinephrine) and traumatization of gill tissue, these changes presumably 

represent a compensatory response to a decrease in gill function. Servizi and Martens 

(1987), and Wedemeyer et al. (1990) observed the alteration in blood chemistry include 

an increase in haematocrite, erythrocyte count, hemoglobin concentration, and elevated 

blood sugar levels (hyperglycemia), plus decreases in blood chloride content, and 

depletion of liver glycogen. Redding and Schreck (1982) and Servizi and Martens (1987) 

reported blood chemistry of fish is changed when sedimentation concentrations range 

between 500 to 1500 mg/L. Sherk et al. (1973) found changes in physiological effect is 

related with a reduction in the swimming endurance of white perch (Morone americana) 

exposed to 650 mg/L of TSS. Newcombe (1994) observed the changes after four to five 

days of exposure. On the other hand Redding and Schreck (1982) found a significant 

increase in haematocrit volume within steelhead trout after only nine hours of exposure to 

500 mg/L of volcanic ash, clay, and topsoil.

Sediment is harmful to the gills, clogging gill mucus and causing asphyxiation. It has 

been reported that gill tissue may be traumatized by increased sediment concentration. 

The severity of damage appears to be related to the dose of exposure as well as the size 

and angularity of the particles involved (Anderson 1996). These factors could account for 

the large range in responses seen for different exposure rates. For example Herbert and 

Merkens (1961) observed that concentrations as low as 270 mg/L cause gill damage in 

rainbow trout (Oncorhynchus mykiss) after 13 days of exposure. Servizi and Martens 

(1991) observed more damage with larger and more angular particles in Arctic grayling 

that were exposed to concentrations as high as 1300 mg/L for four days. On the other 

hand, McLeay et al. (1987) found no gill damage in young-of-the-year with the primary 

mechanisms of action through physical abrasion of tissue and particle adsorption onto the 

gill. Sherk et al. (1973) observed the types of tissue changes, which include swelling of 

secondary lamella and hypertrophy (cell swelling) of epithelial cells. Hyperplasia or an 

increase in cell number of gill tissue have been reported by Simmons (1984) and tissue 

necrosis by Servizi and Martens (1987). In juvenile rainbow trout, secondary effects were 

observed due to infestation of parasitic protozoans that were exposed to high dose of 

sediments. Sentence does not make sense.

Through phagocytosis process particles could reduce resistance to other stressors by 

impairing fish health. Newcombe and Jensen (1995) discussed the process of 

phagocytosis by which fine particles are enveloped by cells within fish gill and gut 

tissues and are transported to internal repository tissues. According to them the main 

repository organ of fish is the spleen and particles could trigger tumor induction, 
9

especially in circumstances where contaminants were adsorbed to particles in suspension. 

Herbert and Merken (1961) observed more susceptibility in rainbow trout to infestations 

of fin rot when fish were exposed for 121 days to concentrations of 270 mg/L of 

diatomaceous earth. They reported that increased concentrations of sediments are related 

with an overall decrease in the ability to defend against disease and to tolerate chemical 

toxins. For example, Servizi and Martens (1991) found a correlation between the 

prevalence of a viral kidney infection and an increased concentration of suspended 

sediments in coho salmon. McLeay et al. 1987 reported that when concentrations of 

suspended sediments exceeded 100 mg/L, the tolerance of arctic grayling to the toxicant 

pentachlorophenol decreased which suggest a decrease in tolerance related to stress 

associated with sedimentation.

Wilber and Clark (2001) summarized the effects of increased sediment concentration and 

reduced water clarity on the feeding of fishes. Sweeten (1995) observed centrarchids 

(e.g., sunfish and black bass) may be severely impacted in their ability to feed by even 

small increases in SABS concentration and turbidity. Ponton and Fortier (1992) reported 

that fish may be squeezed out of desirable habitat because of increased SABS 

concentration and turbidity. Fish that need to see their prey to feed suffer from the 

reduced visibility in turbid water On the other hand, some fishes are able to hunt better as 

sediment concentration increase, at least up to a point, because of increased contrast 

between the prey and the surrounding water. Some fish, like striped bass, seem to be able 

to feed under extremely turbid conditions, or even complete darkness. Chesney (1993) 

suggested this ability could be very important for a fish that follows the turbidity 

maximum for its abundant food. However, Caux et al (1997) suggested it is very difficult 

to make generalizations about turbidity data, because most of the data are reported in 

turbidity units, which are difficult to convert into suspended solids concentrations. 

Fish can also swallow large quantities of sediment, causing illness, reduced growth and 

eventual death. Barton (1977) reported that many species of fish may simply relocate 

when sediment load is increased, because fish can readily disperse. For instance, 

Newcombe (1994), and Newcombe and Jensen (1996) observed the avoidance of turbid 

water in juvenile coho salmon (Oncorhynchus kisutch), arctic grayling (Thymallus 

arcticus), and rainbow trout (0. mykiss). Horkel and Pearson (1976), Goldes et al (1988), 

and Waters (1995) suggested an increased sediment concentration can reduce dissolved 

oxygen in the water column, and in extreme cases may cause a thickening of the gill 

epithelium and reduced respiratory function. 

Increasing concentrations of sediments can cause changes in fish behavior. Fish that need 

to see their prey to feed suffer from the reduced visibility in turbid water. Deposited 

sediment changes the habitat diversity in rivers and streams. Some species are dependent 

on the rocky bottoms of streams and others live in deep sandy pools or around woody 

debris. Deposited sediment can smother whole stream beds and reduce the variation of 

habitats and flows. Newcombe (1994) reported that behavioral changes are generally 

considered benign and transitory and can be easily reversed and do not exhibit a longlasting impact. Typical behavioral responses include an increased frequency of the cough 

reflex, avoidance of suspended sediments, reduction in feeding and temporary disruption 

of territoriality. The severity of the behavioral response is associated with the timing of 
10

disturbance, the level of stress (and associated energy cost) and the importance of the 

habitat that the fish may be being excluded from. The avoidance of suspended sediment 

plumes is one of the first reactions. Bisson and Bilby (1982) observed this behavior 

evoked in juvenile coho salmon at total suspended sediment (TSS) concentrations as low 

as 88 mg/L, and McLeay et al. (1987) observed that arctic grayling (Thymallus arcticus) 

avoided concentrations greater than 100 mg/L.

Cunjak and Power (1986) observed that by remaining in low velocity areas, fish are able

to minimize their energy expenditures and hence reduce the rate of metabolic depletion 

According to their observation, land-use activities that increase the delivery of fine 

materials to streams can significantly affect the overwintering survival of resident fish. 

Clapp et al. (1990) and Nickelson et al. (1992) reported that during winter fish generally

experience decreased energy reserves and as such search for habitat that allows them to 

reduce energy expenditures. Vondracek and Longanecker (1993) found that preferred

habitats are species dependent, however, for most salmonids, preferred habitats are 

located in low velocity areas such as pools and behind instream cover where focal 

velocities are low (Griffith and Smith 1993; Modde et al. 1991; Heggenes and Saltveit 

1990; Cunjak and Power 1986;). Cunjak and Power (1986) suggested that a mechanism 

of potential sediment impact is a depletion of critical energy reserves as a result of 

increased physiological stress, alterations in behavior and/or exclusion from preferred 

sites of overwintering habitat. This is particularly deleterious to fish species and life 

stages that prefer to overwinter within the interstitial spaces of the stream bed. The net 

loss in energy reserves will depend on the concentration of sediment and the duration of 

impact. Dependent upon existing energy reserves, the fish may be able to tolerate the 

energy depletion attributable to an increase in the cough reflex and reduced feeding, but 

may not be able to tolerate the energy depletion associated with displacement from 

critical habitats. 

Sedimentation effects on fish population are also influenced by seasonal variation. 

Cunjak (1996) and Mason (1976) argued that due to natural factors, the availability 

of winter habitat for fish is generally less than that of summer habitat and may be more 

influential in the determination of the stream’s natural carrying capacity. Cunjak (1996)

reported that preferred winter habitat areas of low current velocity are often predisposed 

to sedimentation, and lower flows often experienced during winter may result in higher

rates of sediment deposition. Bjornn et al. (1977) found that the spring freshet from snow 

melt was rarely sufficient to transport sediment out of pools, therefore, the damage to 

these areas is frequently of a longer-term than sediments deposited in more erosive 

habitats. 

Hartman and Scrivener (1990) observed that a reduction in the abundance of already 

limited winter habitat significantly affects the overall fish population of a watercourse. 

Elwood and Waters (1969) observed that increased sedimentation reduced the population 

of invertebrates and hence the capacity of the stream to support brook trout.. A reduced 

food supply, and a greater expenditure of energy in food search and avoidance of higher 

concentrations of suspended sediments may significantly impair the fish’s ability to

compensate for negative physiological changes and survive the winter.
11

Suspended Sediment Effects: 

Increased concentrations of suspended sediments have the potential to affect fish 

populations. Waters (1995) argued that the severity of effects of suspended sediments is a 

function of many factors, including sediment concentration, duration, particle size, life 

history stage, temperature, physical and chemical characteristics of the particles, 

associated toxicants, and acclimatization. Lloyd (1987) reported lethal effects to fish at 

concentrations ranging from 500 to 6000 mg/L of suspended sediments. Sigler et al. 

(1984) observed mortality in young coho salmon and steelhead trout at 500 to 1500 

mg/L. Stober et al. (1981), Servizi and Martens (1992) and Smith (1940) reported that 

short term increases in suspended sediment concentrations between 11000 and 55000 

mg/L appear to be the point at which salmonid mortality significantly increases. McLeay 

et al. (1987) reported that arctic grayling exposed to sediment concentrations greater than 

1000 mg/L were slower to recognize food. . Feeding rate may be a function of prey 

visibility. However, Sigler et al. (1984) suggest that reduced feeding is more complex 

than reduced ability to see prey items, as many fish species (especially benthic feeders) 

do not use sight to identify prey items, but still exhibit reduced levels of feeding in 

response to elevated sediment loads. 

Juvenile and adult fish generally appear to be more resilient to stress from suspended 

sediments than other life history stages. Wilber and Clarke (2001) reviewed the data on 

the effects of suspended sediment on fish with an emphasis on . impacts of dredging on 

estuarine organisms. The authors synthesized dose-response relationships of estuarine 

aquatic organisms to suspended sediments and then related those findings to sediment 

conditions associated with dredging projects. They added to the data from Newcombe 

and Clark (1996) and provided a useful way of plotting the empirical data, such that all of 

it can be seen, and compared with expected exposure concentrations. This display of data 

provides a powerful tool for the estimation of expected effects from a given suspended 

sediment exposure scenario. Wilber and Clarke (2001) plotted the adult estuarine fish 

data separately from the freshwater and salmonid data, to show how little data there were 

for the fishes, and that most of those data were from short duration tests at very high 

exposures. However, the two data sets do not appear to be much different between the 

estuarine fishes and their salmonid counterparts. Wilber and Clarke (2001) also provide 

tables that depict biological response as a function of suspended sediment dosage 

(sediment concentration/unit time). Biological Response categories reported by Wilber 

and Clarke (2001) include no effect, behavioral, sub-lethal, and lethal effects. It is 

difficult to tell from the data, however, whether this is because estuarine fishes and 

salmonids really have similar sensitivities to suspended sediment, or because the 

estuarine fishes have almost exclusively been tested at very high sediment loads for short 

periods of time (Berry et al., 2002). Interestingly, if the egg and larval data for the 

estuarine fishes are plotted and compared to these data to the egg and larval data from the 

salmonid and other freshwater fishes, there does seem to be a differences between the 

two habitats, except that the estuarine fish eggs and larvae were less sensitive to 

suspended sediments than the salmonid and freshwater fish eggs and larvae.

High concentration of suspended sediments has also been associated with the loss of 

territoriality and interruption of movements of salmonids. Berg and Northcote (1985) 
12

found that territorial behavior was lost at concentrations exceeding 30 NTUs. There is no 

transition here from talking about suspended sediments and bed permeability.

The permeability of the stream bed is determined by size composition of the substrate 

material, viscosity of the water (temperature dependent) and the packing of the substrate 

material (Stuart1953; Cooper 1965). A small increase in the proportion of fine material 

can severely reduce the porosity and permeability of the gravel bed (Lisle and Lewis 

1992) and the ability of alevins to receive adequate oxygen and emerge from the gravel. 

Newcombe and MacDonald (1991) first realized that the appropriate way to report data 

for the effects of suspended sediment on aquatic organisms was to include information on 

duration of exposure, as well as exposure concentration. Up until that point, the 

importance of duration of exposure had been largely overlooked. They summarized, in 

graphical and tabular form, much of the available data on the effects of sediments on fish 

and invertebrates. Newcombe and Jensen (1996) presented an extensive data table of the 

effects of suspended sediments on fish, and went a step further, developing empirical 

models of the effects of sediment on fish with matched biological effect, concentration 

and duration information. The effects were scored on a qualitative “severity of ill effect” 

(SEV) scale, that included everything from “no behavioral effects” (lowest on the scale) 

to behavioral effects (low on the scale); to sublethal effects (higher on the scale); to lethal 

effects (highest on the scale). 

Newcombe (1994) and Newcombe and Jensen (1996) observed that increased mortality 

has been associated with increased suspended sediment for arctic grayling (T. arcticus), 

Atlantic silverside (Menidia menidia), rainbow trout (0. mykiss), fourspine stickleback 

(Apeltes quadracus), white perch (Morone americana), yellow perch (Perca flavescens), 

American shad (Alosa sapidissima), striped bass (M. saxatilis), coho salmon (0. kisutch), 

chinook salmon (0. tshazuytscha), chum salmon (0. keta), striped killifish (Fund&s 

majalis), sheepshead minnow (Cyprinodon variegatus), and even carp (Cyprinus caqio). 

Newcombe also developed a model for the effects of sediment with the effects of 

diminished water clarity on fish. Different models were developed for different groups of 

fishes: juvenile and adult salmonids together, adult salmonids, juvenile salmonids, eggs 

and larvae of salmonids and non-salmonids, adult estuarine non-salmonids, and adult 

freshwater non-salmonids. The empirical data were displayed in tabular form. The 

models were presented both in visual form (as 3-dimensional response surfaces) and as 

linear regression equations, and are also used to interpolate and extrapolate from the 

empirical data. Gardner, 1981; Berkman and Rabeni, (1987I) reported that the 

abundances of visual feeders, like sunfish and trout, decline with elevations in turbidity. 

Gardener (1981) found that the feeding rate of bluegills (L. macrochirus) decreased at a 

turbidity of 60 NTU, and Breitburg (1989) noted striped bass larvae consumed less D. 

pulex with increased sediment concentrations. Ryan (1991) reported that species that 

depend on drifting insects as a major food source, including trout and salmon, have also 

shown depressed feeding rates. 

Stowell and co-workers (2001) reported that deposition of 25% fine sediments in 

spawning rubble or gravel reduces fry emergence by 50%. Sedimentation also has 
13

negative impacts on the invertebrate food supply of many fish. Sediment deposition 

affects rearing habitat of juvenile fish since young salmonids frequently use the 

interstitial spaces in the stream bed for cover. Berkman and Rabeni, (1987) suggested that 

reproductive strategies that involve parental care, such as fin fanning and egg nipping and 

mouthing, appear to be more successful in sedimented habitats, than those species that 

rely on substrate or pelagic broadcasting of eggs. Thus, a reduction in the suitability of 

potential rearing habitat as a result of sediment introduction is related to a reduction in 

the space available for occupancy (Reiser et al. 1985). When pools and interstitial spaces 

in gravel fill with sediment, the total amount of habitat available for rearing is reduced 

(Bjornn et al. 1977). Griffith and Smith (1993) found that numbers of juvenile rainbow 

trout and cutthroat trout decreased due to lack of available cover in heavily embedded 

gravel substrata. Bjornn et al. (1977) observed that interstitial space is important during 

winter because juvenile fish live in these areas making them especially susceptible to 

impacts from increased sedimentation. Without these inter-gravel refugie, young fish may 

abandon the stream or move to less suitable areas where survival rates may be reduced. 

Ringler and Hall (1975) reported that increased temperature and reduced dissolved 

oxygen levels of intragravel water in salmon and trout spawning beds because of clearcut logging practices. An associated reduction in resident cutthroat trout populations was 

attributed to this reduction in spawning habitat suitability. However, they suggested the 

failure to document serious reductions in coho salmon could be related to sediment 

clearing and removal by these larger fish during operation work (redds construction). 

Scott et al (2002) examined the relationship between relative abundance of endemic 

stream fishes and cumulative sediment load in an analysis of data from five 

physiographic regions. Objectives were to determine the functional form of the 

relationship and assess the usefulness of endemic fishes as ecological indicators. At 137 

sites where fishes were sampled, sediment load was modeled from watershed land cover 

and calculated as percent increase over background rates. The shape of the relationship 

suggested that sediment may be a limiting factor on abundance of endemic stream fishes. 

Quantile regression models of the upper 10% of data points along the range of sediment 

loading indicated that for each 1,000% increase in sediment load over background rates, 

the maximum relative abundance of endemic fishes supported by a stream was reduced 

by 2.2%. These results need further evaluation with additional data on a region-by-region 

basis. Griffith and Walton (1978) have shown that the upper tolerance level for 

suspended sediment is between 80-100 mg/l for fish, and as low as 10-15 mg/l for 

benthcinvertebrates. Recovery of the aquatic biota from increased sedimentation is 

dependent on the severity of sediment additions and the discharge level of the rivers or 

streams. Recovery from short term additions of sediment is usually complete within one 

year.

Effect due to Sediment Deposition, leading to increased embeddedness or burial of 

eggs: 

Sediment can settle in crevices where fish deposit their eggs. High levels of sediment can

completely fill in deep pools, destroying the entire habitat for some species. The 

freshwater blackfish, for example, lays sticky eggs, which only attach to clean, unsilted 
14

surfaces. Interstitial spaces are also critical for the eggs and juveniles of Macquarie 

Perch, and the Broad-finned and Spotted Galaxias. Recent studies on several species of 

native fish have found that all of their eggs die even when lightly coated in fine sediment. 

Waters (1995), and Wilber and Clarke (2001) suggested fish reproduction is severely 

affected by sediment deposition. According to Burkhead and Jelks (2002) excessive 

sedimentation is a ubiquitous problem in southeastern rivers. They inferred that 

sedimentation could negatively affect reproductive success in benthic spawning fishes. 

The primary mechanisms of action are through increased egg mortality, reduced egg 

hatch and a reduction in the successful emergence of larvae. The primary cause of egg 

death is from burial by settled particles. Bodaly and Fudge (1984) reported that 

sedimentation rates of 0.03 to 0.14 g dry weight sediment/cm2(i.e., 1-4 mm depth of silt 

and clay) significantly reduced the survival of lake whitefish (Coregonus clupeaformis) 

eggs. Thin coverings (a few mm) of fine particles are believed to disrupt the normal 

exchange of gases and metabolic wastes between the egg and the water. The effects upon 

egg mortality appear to be more closely related to the sedimentation of particles and less 

related to the concentration of suspended sediments. 

Leedy (1975) found a 94% reduction in numbers and weight in large game fish due to 

sedimentation. Salmonids are especially vulnerable to sedimentation because they lay 

their eggs in gravel and small rubble with water flow sufficient to maintain oxygen 

supply. Fine sediments may cement spawning gravels, impeding the construction of 

redds. Increases in fine sediments also reduce the availability of oxygen to eggs and 

increase embryo mortality. Alabaster and Lloyd (1982) and Ryan (1991) reported that 

sedimentation decreases available spawning habitat, reduces spawning activity, and 

increases egg and larvae mortality. As sedimentation occurs, spawning habitat also may 

be smothered. This is especially true for substrate spawners (Muncy et al., 1979). If 

sedimentation occurs after spawning, then oxygen supply to eggs and sac fry in the 

substrata may be decreased due to reductions in water circulation. Waters (1995), Argent 

and Flebbe (1999), and Reynolds et al (1989) showed an elevation in sac fry mortality of 

arctic grayling (T. arcticus) with increases in turbidity.

Bjornn et al. (1977) found that the number of juvenile salmon that a stream can support in

winter was greatly reduced when the inter-cobble spaces were filled with fine sediment. 

The decreased carrying capacity was a function of both a loss of substrate cover for 

juvenile fish and a reduction in food as benthic invertebrate communities changed. 

Bjornn et al. (1977) suggested that the summer rearing or winter holding habitat may be

more influential to the carrying capacity of a stream reach than embryo survival.

Shaw and Maga (1943), and Cordone and Kelley (1961) demonstrated an increase in

percent of fine material in the stream bed can have impacts on egg survival rates since it 

reduces streambed permeability. Lowered permeability reduces the interchange between 

stream flow and water movement through the redd, resulting in a reduction in the supply 

of dissolved oxygen to the egg and a hindrance to the removal of metabolites. Slaney et 

al. (1977) reported that rainbow trout egg survival was significantly reduced when 

spawning gravel contained more than 3% of fines (0.3 mm diameter ). Hall and Lantz 

(1969) determined that hatching success of coho salmon and cutthroat trout 
15

(Onchorhynchus clarki) was reduced by 40 to 80% when spawning substrates contain 20 

to 50% fines (1-3 mm diameter). Zallen (1931) observed that concentrations of 1000 to 

3000 mg/L had no effect upon the survival of mountain whitefish eggs (Prosopium 

williamsoni). Campbell (1954), however, found 100 percent mortality in rainbow trout 

eggs exposed to TSS concentrations of 1000 to 2500 mg/L. Slaney et al. (1977) noticed 

that hatching success for rainbow trout was reduced after 2 months of exposure to 57 

mg/L. A significant reduction in the hatching success of white perch and striped bass 

(Morone saxatilis) was observed in only 7 days after exposure to about 1000 mg/L TSS 

(Auld and Schubel 1978). Burkhead and Howard (2002) used a simple ANOVA tool to 

test the effects of suspended clay fines on reproductive success in the tricolor shiner 

Cyprinella trichroistia. They reported that differences in egg mortality effects 

associatedwith elevated sediment loads is related to the size of the sediment particles 

involved and rates of sediment deposition. 

According to Anderson (1996) the magnitude of the effect of sediments exposure may 

also be influenced by the timing of sediment exposure with respect to the stage of embryo 

development. The dose of sediment required to induce egg mortality is greatly influenced 

by the physical characteristics of the stream which, in turn, affects sediment transport 

capabilities and the capacity to maintain sediments in suspension or otherwise to result in 

their deposition. The magnitude of impact upon fish resulting from increased 

concentrations of suspended sediments and levels of sedimentation can vary seasonally. 

Early live stages (i.e., eggs, alevins) of many salmonids are found in the stream bed 

during the winter months. These stages are particularly sensitive to the effects of 

increased concentrations of sediments and the deposition of fines sediments. The 

introduction of sediments during the winter, therefore, has the potential to appreciably 

influence these early life stages. River spawning salmonids typically deposit their eggs in 

gravel beds commonly found in the upper reaches of river systems. Scullion and Miller 

(1979) reported that brown trout typically bury eggs in interstitial spaces of the substrata 

to depths of 9 to 12 cm. Alevins remain in the interstitial spaces until the start of 

exogenous feeding. Lisle and Lewis (1979) reported that the percolation of water through 

the incubation substrate is an essential factor in determining the survival rate of 

incubating eggs.

Koski (1966) and Phillips et al. (1975) reported that even if intergravel flow is adequate 

for embryo development, sand that plugs the interstitial areas near the surface of the 

stream bed can prevent alevins from emerging from the grave. Weaver and Fraley (1993) 

observed that the emergence success of west-slope cutthroat trout was reduced from 76% 

to 4% when fine sediment was added to redds. Female stream-spawning salmonids 

typically clean an area of the stream bed in which they bury their eggs. This nest building 

activity flushes sediments and increases the stream bed permeability. Wickett (1954),

and McNeil and Ahnell (1964) suggested that, with time, sediment conditions within the 

redd gradually return to ambient levels. Under normal conditions, this slow increase in 

sediment intrusion is not a problem, however, increased levels of sediment within a 

system as a result of anthropogenic disturbance increase the rate and level of sediment 

intrusion and reduces the period of time in which the redd is clean. The period of time 

before sediment intrusion into the redd is very important with respect to the survival

of salmonid larvae. Wickett (1954) observed that sediment accumulation during early 
16

embryonic development may result in higher egg mortalities than if deposition occurs 

after the circulatory system of developing larvae is functional. According to Shaw and 

Maga (1942) this may be due to thehigher efficiency in oxygen uptake by the embryo or 

alevin with a functional circulatory 

system. 

At Allen Brook (2001), it was observed that many species of fish and macroinvertebrates require the interstitial spaces at the bottom of streams to lay their eggs. 

Most fast flowing, lower order streams like Allen Brook tend to have mostly cobble 

bottoms with large interstitial spaces. Excess sedimentation fills in the spaces between 

cobbles on the streambed and limits the reproductive success of these aquatic species. 

Excess fine sediment can also lead to increased levels of total suspended solids (TSS). 

Higher than normal levels of TSS can cause increased stress in aquatic biota, leading to 

decreased productivity. Macroinvertebrates serve the role of primary consumers in Allen 

Brook. As such they serve as a food source for secondary consumers (fishes). 

Wilber and Clark (2001) observed that increased sedimentation in other habitats (e.g., 

estuaries) can cause burial of eggs. An increased supply of fine sediment to a stream can 

cause the gravel interstices of a stream bed to be filled in. This process can cause 

reduced hatching due to the reduction in flow through the stream bed and the resulting 

decrease in dissolved oxygen. It can also cause reduced larval survival because of 

armoring of the sediment surface, which traps the larvae. Even a small amount of 

deposited sediment can cause a problem. David Nelson (NMFS, unpublished data) 

observed that winter flounder eggs suffer reduced hatching if buried to only one half an 

egg diameter. Of all of the taxonomic groups, fishes, particularly salmonids, have 

received the most attention from sedimentation researchers. According to Waters (1995) 

this is because of the commercial and recreational importance of salmonids, and the 

obvious impact that logging and other land use activities have had on salmonid fisheries, 

particularly in the Pacific northwest 

 

Effect on Invertebrates:

There are multiple ecological effects of sedimentation on aquatic invertebrates. In the 

literature it has been reported that high sediment levels resulted in decreased levels of 

benthic organisms. Increased sediment loads in streams also have a significant effect on 

zooplankton. Gammon (1970) and Lenat et al. (1981) suggested that sediment impact the

density, diversity and structure of invertebrate communities. Wagener (1984) and Mend

(1989) in their studies demonstrated decreases in invertebrate densities and biomass due 

to sedimentation. Increases in sediment input reduce the density of invertebrates due to 

physiological activities or by habitat alteration. Tsui and McCart (1981) reported that 

sediments have an abrasive effect on invertebrates and interfere with the respiratory and 

feeding activities of benthic animals. Cordone and Kelley (1961) and Tsui and McCart 

(1981) reported that increased sediment deposition reduces the biomass of invertebrates. 
17

After reviewing the literature, sedimentation effects on invertebrates can be categorized 

as: Direct impacts due to abrasion, clogging of filtration and respiration mechanisms and 

smothering and burial resulting in mortality. 

Indirect impacts from light attenuation leading to changes in feeding efficiency and 

behavior (e.g., drift ) and alteration of habitat from changes in substrate composition

Waters (1995) identified three major relationships between benthic invertebrate 

communities and sediment deposition in streams: 

1) Correlation between abundance of macroinvertebrat and substrate particle size, 

2) Embeddedness of substrate and loss of interstitial space, and 

3) Change in species composition with change in type of habitat (substrate composition).

The other important observations by the Water (1995) is that sedimentation alters the 

structure of the invertebrate community by causing a shift in the proportion from one 

functional group to another. For example, streams with clear water normally contain a 

high proportion of invertebrates in the shredder group. Bode (1988) and Wagener (1984) 

reported when sediment deposition is systematically increased, alteration to other groups 

such as grazers or collector- gatherers is resulted. Rosenberg and Snow (1975), Dance 

(1978), and Lenat et al. (1981) observed that increased sediment deposition cause a 

change towards chironomid-dominant benthic communities.

Davies and Nelson (2001) studied the impact of increased sedimentation due to lack of 

riparian buffers on macroinvertebrate abundance. They tested thirty-four streams, of 

which all those lacking riparian buffers upstream (which help control sedimentation) had 

lower macroinvertebrate abundance. Cordone and Kelly (1961) Peddicord (1980), 

Waters (1995), and Wilber and Clarke (2001) reported that elevated levels of sediments 

have a wide range of effects on both water column and benthic invertebrates. Simon et al 

(2001) demonstrated that a significant increase in the amount of sediment in a stream’s 

water column or on its bed surface is detrimental to aquatic organisms. Sedimentation can 

lead to embeddedness, which blocks critical macro-invertebrate habitat by filling in pores 

between rocks on the stream bottom. The State of Vermont (2001) observed that 

sedimentation has increased embeddedness in Styles Brook in Vermont’s West River 

Basin, which has degraded the brook’s macro-invertebrate habitat and caused a shift in its 

macroinvertebrate community composition. 

Angradi (1999) lists numerous studies on the effects of fine sediment upon 

macroinvertebrate colonization of introduced substrates, non-manipulative sediment 

impact studies, and influences on macroinvertebrate behavior. In the northern 

Appalachians, Angradi (1999) examined fine sediment effects upon macroinvertebrates 

and tested the reliability and sensitivity of metrics in the detection of these effects. This 

study combined a survey of 15 streams with a sediment manipulation experiment. The 

choice of manipulating the proportion of fine sediment less than 2 mm in size was 

determined by previous research at the Fernow experimental forest. Angradi (1999) 

found three metrics, EPT (Ephemeroptera, Plecoptera, and Trichoptera) taxa richness, % 

Orthocladiinae (Diptera: Chironomidae), and % Chironominae (Diptera: Chironomidae) 

to respond to fine sediment in both the stream survey and field experiment. The field 
18

experiment reduced some confounding environmental variables found in the field survey 

while revealing Ephemeroptera, Plecoptera, and Trichoptera decline in diversity relative 

to sediment-specialists when sedimentation levels become excessive. The response of 

these metrics in both the field survey and experiment indicated they were reliable and 

sensitive metrics for sediment assessment. Tebo (1955) found logging increased the 

proportion of inorganic sand and silt within the stream resulting in lower abundances of 

macroinvertebrates. The lower abundances were attributed to a lack of food in inorganic 

silt and the unstable nature of the sand substrate. Furthermore, Tebo (1955) found the 

unstable sand substrate to be subject to loss during flooding causing the associated 

macroinvertebrate community to be swept downstream in the process. The lower 

abundances were attributed to a lack of food in inorganic silt and the unstable nature of 

the sand substrate. 

Invertebrate drift is directly affected by increased suspended sediment load in freshwater 

streams and lakes. Herbert and Merkens (1961) reported that increases in suspend 

sediments (e.g., 120 mg/l) can result in increased drift which significantly changes the 

distribution of benthic invertebrates in streams. Waters (1995) considered the effects of 

increased deposited sediments on benthic invertebrates as one of the most important 

concerns within the sediment pollution issue, especially in regards to the dependence of 

freshwater fisheries on benthic productivity. In northern Idaho, Luedtke and Brusven 

(1976) examined the effects of sand accumulation on benthic and drifting 

macroinvertebrates. They found that while sand was not a barrier to drifting insects, it 

was not colonized by the drift. Areas of sand accumulation acted as barriers to dispersal 

by benthic macroinvertebrates who were unable to crawl or burrow through the sand due 

to its unstable nature. This created isolated regions of cobble and gravel with high 

macroinvertebrate abundances separated by these areas of sandy accumulation. Since 

mating dispersal flights are typically of short distance, limitations to genetic diversity 

may occur if the distances between isolated macroinvertebrate populations become too 

great. Extreme cases of isolation may lead to extirpation through limited food resources 

for predators andgenetic bottlenecks. In an Indiana stream, Lamberti and Berg (1995) 

suggested community functional group changes may have been due to sedimentation. In 

areas of sedimentation, they found a 292% increase in gatherers while observing an 83% 

decrease in filterers in comparison to reference sites. They believe this reflected an 

interruption in CPOM processing due to sedimentation possibly at the shredder stage.

As deposited sediment increases, changes in community structure and diversity can 

occur. These changes generally involve a shift in dominance from EPT taxa to 

oligochaetes (Lenat et al. 1979), especially Naiadae and Tubificidae (Matter and Ney 

1981), and to amphipods and gastropods (Hall et al. 1984). Zweig and Rabeni (2001) 

have shown that density of benthic invertebrates is significantly lowered before 30% 

increase in deposited sediment is reached. Other laboratory studies indicate that 

invertebrate densities are not affected by embededness levels under 33% (Walters 1995). 

A strong correlation between EPT taxa and sediment loading has been suggested. 

Increased embededness can result in decreases in aquatic insect densities and small 

increases in siltation can directly affect caddisfly pupa survival. Number of EPT taxa, 

number of species, and ratio of EPT to chironomids, all indicators of a pollution 
19

intolerant assemblage, are negatively associated with embeddedness and positively 

associated with width-to-depth ratio and percent riffle (Nerbonne and Vondacek 2001). 

Zweig and Rabeni, (2001) examined the response of benthic infauna to deposited fine 

sediments in four Missouri streams. Five biomonitoring metrics were significantly 

correlated across streams with deposited sediments. Deposited-sediment tolerance values 

were developed representing responses to deposited sediments for 30 taxa. Tolerance 

values were used to develop the Deposited Sediment Biotic Index (DSBI). The DSBI was 

calculated to characterize sediment impairment in the four streams. DSBI values for each 

site examined were highly correlated with depth and degree of embeddedness of 

deposited sediment.

Gore (1985) reported that macroinvertebrate density and diversity are directly related to 

substrate diversity. Lenat et al (1981) reported that as sediment settles, interstitial spaces 

between coarse substrata are filled, which reduces available habitat for macroinvertebrates. If enough sediment fills interstitial spaces, then an impermeable sediment 

barrier may form that diminishes available living habitat and causes reductions in 

hyporheic oxygen levels due to the inhibition of interstitial water circulation (Beschta and 

Jackson, 1979; Gordon et al. 1992). Ryan (1991) concluded that a 12 to 17% increase in 

interstitial fine sediment is associated with a 16 to 40% reduction in the total abundance 

of invertebrates. Insect community structure may change with sedimentation of the 

substrata. As a result of which habitat alterations that favor burrowing insects tolerant of 

low oxygen levels have been noticed (Ryan, 1991). Wagener and LaPerriere (1985) 

reported that sedimentation decreased density and biomass in benthic macro-invertebrate 

communities, and stated that turbidity was the strongest descriptor related to such 

reductions. Tsui and McCart (1981) found that densities and standing stocks of lotic 

insects were inversely related to levels of sedimentation. 

Rosenberg and Wiens (1978) and Culp et al.(1986) have shown increase in 

macroinvertebrate drift with increases in substrata sedimentation. Rosenberg and Snow 

(1975) detected increased drift of macrobenthos with an increase in suspended sediment.

Fairchild et al. (1987) observed changes in drift patterns and community structure of 

benthic insect communities with weekly exposure to sediment concentrations of 1700 

mg/L, and reported that recovery from this type of exposure was gradual. Gammon 

(1970) found that shifts in benthic invertebrate communities were characterized by 

increases in silt tolerant genera such as the mayfly Tricorythodes. These shifts were 

observed by Gammon, (1970) at suspended sediment concentrations as low as 

approximately 53 mg/L.

According to Newcombe and MacDonald, (1991) increases in sediment deposition that 

affect the growth, abundance, or species composition of the periphytic (attached) algal 

community will also have an effect on the macroinvertebrate grazers that feed 

predominantly on periphyton. A change in particle-size distribution in the stream bed can 

alter the habitat and make it unsuitable for certain species of invertebrates. There is no

transition here to the next sentence.

Gammon (1970) noticed that an increase in suspended sediments from 40 to 120 mg/L 
20

resulted in a 25 to 60% decrease in the density of stream macroinvertebrates. Likewise, 

Slaney et al. (1977) found that a 16 hour pulse of suspended sediments (2500 to 3000 

mg/L) led to a 75% reduction of invertebrate biomass within the most affected areas.

In a study in Allen Brook it was shown that excess sedimentation from the surrounding 

watershed has various negative impacts on the aquatic biota. Macroinvertebrates, which 

have been studied for decades in this area and have evolved as acceptable indicators of 

stream health, are the main ecological receptors. Their specific assemblages in streams 

can be referenced and tracked over time to indicate changes in water quality. Indicator 

macroinvertebrates for stream health were Ephemeroptera, Plecoptera, and Trichoptera 

(EPT). At the testing reach, EPT scoring results failed EPA standards for water quality. 

In another study in Chattooga River watershed accelerated sedimentation has been 

identified by Pruitt et al (2001) to be the leading determinant in loss of habitat and 

reduction in bed form diversity within the study area. A significant correlation was 

observed between aquatic ecology and normalized total suspended solids (TSS) data. 

Based on overlaying the biological index on TSS normalized to discharge/mean 

discharge, TSS concentrations greater than 284 mg/l adversely affected aquatic 

macroinvertebrate community structure. However, based on historic regional suspendedsediment concentrations, a normalized TSS concentration of 58 mg/l or less during storm 

flow provides an adequate margin of safety and is protective of aquatic macroinvertebrates in the Blue Ridge physiography. Corresponding turbidity limits of 69 and 

22 NTU established the threshold of biological impairment and margin of safety, 

respectively. According to Hanley etal, (1995) previously, a similar turbidity of 25 NTU 

has been recommended for stream restoration management plans. 

Benthic fauna possess behavioral and morphological adaptations 

which limit them from being displaced in a unidirectional flow environment (Hynes 

1973). Invertebrate drift is a continuous redistribution mechanism that occurs in most 

stream ecosystems. It is an important factor in the regulation of population density 

(Williams and Hynes 1976), in the dispersion of aggregations of young larvae (Anderson 

and Lehmkuhl 1967), in the abandonment of unsuitable areas (Williams and Hynes 

1976), and in the recolonization of areas after disturbance (Barton 1977).

Rosenberg and Weins (1978) reported that invertebrate drift may be induced by elevated 

suspended sediment levels. Rosenberg and Snow (1975) observed that increased rates of 

downstream drift by macrobenthos can be induced by concentrations as low as 23 mg/L. 

Drifting affords invertebrate taxa that are sensitive to increased sediment loads the 

opportunity to avoid areas which become unsuitable as a result of high suspended 

sediment levels. Conversely, Williams and Hynes (1976), Barton (1977), and Young 

(1986) argued that invertebrate drift is considered to be the most important 

component of ecosystem recovery following stream disturbances. Waters (1964) 

supported the view and mentioned that this is especially true in areas of swift-flowing 

waters. Minshal (1984) observed the importance of benthic macro-invertebrates for 

monitoring sediment conditions in streams because substrate is thought to be the most 

important factor regulating invertebrate distribution and abundance at the local or reach 

scale. Southwood (1977) argued that biodiversity is controlled by the habitat structure 

within the stream and thus increased sediment load alters the streambed resulting in an 
21

altered habitat. Southwood (1977) suggested that changes in stream bottom conditions 

(sedimentation) due to urbanization, agricultural runoff, removal of riparian buffers, and 

stream channelization is detrimental to aquatic habitat.

Newcombe (1996) developed a model for the effects of sediment on invertebrates. The 

effects of sedimentation on aquatic habitat and biota are dependent on measured levels 

and persistence of sediment load (Cairns, 1990). High and sustained levels of sediment 

may cause permanent alterations in community structure, diversity, density, biomass, 

growth, and rates of reproduction and mortality. Impacts on aquatic individuals,

populations, and communities are expressed through alterations in local food webs and 

habitat. Miller et al (2002) worked on the detrimental effects of sedimentation on marine 

benthos and suggested that their results are intended to address the ecological aspects of 

dredge materials placement and site selection. According to these authors quantifying 

natural sedimentation rates and the susceptibility of macrofaunal functional groups is one 

approach towards predicting environmental impacts. If materials placement can be 

planned to be analogous to natural events, then community responses will follow natural 

seasonal and successional trends. When sedimentation exceeds natural thresholds, 

ensuing impacts will likely involve total loss of the community and subsequent 

colonization by pioneer species. Thus an entirely different suite of ecological processes 

will drive impacts and recovery, potentially leading to dramatically altered benthic 

communities.

Invertebrates (such as mayflies) use gills for respiration and in silt-laden water their gills 

become clogged and less effective. Some invertebrates filter food from the water using 

nets spun from silk, (such as caddisfly larvae) or other filtering devices, (e.g., blackfly 

larvae have modified antennae that can filter water). These devices become coated or 

clogged with silt, limiting their effectiveness. Invertebrate species have adapted to life in 

running water by developing highly complex methods for securing themselves to the 

stream bed. Leeches, snails and some fly larvae are equipped with small suction devices. 

Caddisfly larvae construct cases from materials found on the sediment bedded stream and 

attach themselves with silk when they pupate. Blackfly larvae use large silk-producing 

salivary glands to construct a tangled mat on the substrate to which they cling by way of 

hooks. All these methods prevent the animal from being washed away, but they rely on 

attachment to clean surfaces. These complex methods of holding on are ineffective on 

silt-covered surfaces. Many invertebrate species utilise spaces within the actual stream 

bed. The stream bed, particularly in rocky streams, forms a matrix of particles with 

water-filled spaces in between. Some species spend most of their life in these spaces 

within the stream bed, others use it as juveniles to escape predators, or during floods to 

avoid being washed away.

Fine sediment effects upon aquatic ecosystems are not universally negative.

Some fine sediment is important to aquatic ecosystems as habitat for particular groups

of invertebrates (Pennak 1978; Murphy and Meehan 1991; Thorp and Covich 1991). In

larger lotic ecosystems, fine sediment particles in interstitial spaces provide surfaces for

the development of microbial populations important in the processing of dissolved

organic carbon (DOC) into fine organic carbon (FOC) (Fiebig and Marxen 1992). Yet,
22

even in stream and river systems where microbes are important to organic matter

processing, excess amounts of fine sediment have negative consequences to

macroinvertebrates (Brunke and Gonser 1999). Brunke and Gosner (1999) reported

the strongest predictor of invertebrate density was the ratio of particulate organic carbon

to total fine sediment particles. As fine inorganic particles increased, the ratio of organic

carbon to total fine sediment particles decreased leading to a decrease in invertebrate

density. Therefore, it was not the presence of fine sediment particles, but excess

amounts, that had negative effects upon invertebrate density.

Effect on Mussels: 

Aldridge et al. (1987) found that increased levels of suspended sediment have adverse 

effect to ingestion rates of freshwater mussels in laboratory studies. Tester and Turner 

(1988), Aruda et al. (1983) and Sherk et al. (1976) reported reduced feeding activity in 

some species of mussels as a response to increased levels of suspended sediments. 

Aldridge et al. (1987) in an attempt to determine the mechanism by which sediment 

exposure leads to mortality of mussels, showed that exposures of 600 to 750 mg 

sediment/L reduced clearance rates and nitrogen excretion and increased 0:N ratios for 

the pimpleback (Q. pustulosa), southern pigtoe (F. cerina), and Mississippi pigtoe 

(Pleurobema beadleanum). The authors hypothesized that the effect of increased 

sediment exposure on these species was starvation due to decreased filtration rates. 

However, most studies dealing with the effects of sediment on freshwater mussels only 

have addressed survival after exposure to sediment, and placed little emphasis on the 

cause of mortality.

Deposited and suspended sediments have negative affect on the survival of freshwater 

mussels (Unionidae). However, it has been suggested that survival may be species 

specific. Ellis (1936), for example, showed that 0.6 to 2.5 cm of deposited silt resulted in 

significant mortality of yellow sandshell (Lampsilis teres), while threehorn wartyback 

(Obliquaria reflexa), mapleleaf (Quad&a quadrula), and monkeyface (Q. metanevv-a) 

were more resistant. The giant floater (Pyganadon grandis) had better survival than either 

Wabash pigtoe (Fusconaiaflava) or black sandshell (Ligumia recta) when buried in 

detritus, sand, mucky sand, or silt (Imlay, 1972), and F. Java appeared to be more 

sensitive to silt and sand than either the plain pocketbook (L. cardium) or fatmucket (L. 

siliquoidea) (Marking and Bills, 1980). Imlay (1972) reported that P. grandis, a species 

often exposed to deposited sediment in lenthic habitats, experienced high morality (85 to 

100% mortality) when exposed to 45 cm of silt. Box and Mossa, (1998) observed that

eastern elliptio (Elliptio complanata), a species of freshwater mussel commonly thought 

to be very tolerant of sedimented conditions, showed depressed growth rates in muddy 

substrates. They noted that many of the >30 extinct species of North American freshwater 

mussels resided in riffle and run habitats, including 15 species of the genus Epioblasma. 

It is probable that the decline of this genus is due to the loss and degradation of such 

habitats.

Several studies have been performed by marine biologists on the effects of sedimentation 

on bivalves that can be used to draw inferences about effects on freshwater species. 
23

Bayne et al. (1993) showed that the blue mussel, Mytilus edulis, compensates for 

increased levels of suspended sediment by (1) increasing filtration rates, (2) increasing 

the proportion of filtered material that is rejected, and (3) increasing the selection 

efficiency for organic matter. Newell and Jordan (1983) demonstrated that the Eastern 

oyster, Crassostrea virginica, also compensates for increased sediment loads by sorting 

particles and selective feeding, such that pseudofecal material produced by these animals 

contains less energy, nitrogen, and carbon/mg than ingested food. Vahl, (1980), Robinson 

et al. (1984), and Ward and MacDonald (1996) suggested that selective feeding in 

sedimented conditions does not appear to be common to all species of bivalves. The surf 

clam (Spisula solidissima), Iceland scallop (Cblamys islandica), and Atlantic pearl-oyster 

(Pinctada imbricata) have bulk pseudofecal production with limited selective rejection of 

unwanted food particles. Consequently, dietary absorption efficiencies for these species 

tend to decrease during exposure to sediment. Bricelj and Malouf (1984) concluded that 

mussel species-specific responses to sediment are likely adaptations to sediment levels in 

the local environment, such that species inhabiting turbid environments are better able to 

select between organic and inorganic particles.

Hneley et al (2000) suggested that many of the endangered freshwater mussel species 

have evolved in fast flowing streams with historically low levels of suspended sediment, 

such species may not be able to actively select between organic and inorganic particles in 

the water column. Therefore, even low levels of sediment may reduce feeding and, in 

turn, have sublethal effects on unionid growth and reproduction. For example, Box and 

Mossa (1998) demonstrated that fine sediment may interfere with feeding rates and 

indirectly affect growth through a reduction of photosynthetic production. The effects of 

sedimentation and turbidity on ubiquitous taxonomic groups such as crayfish and snails 

are not extensive in the scientific literature. According to Hneley et al (2000) crayfish are 

presumed to be environmentally tolerant of episodic perturbations to habitat quality, but 

no testing of sedimentation effects has been conducted. Similarly, snails have not been 

tested for sediment tolerance, although effects on their food resources, such as benthic 

algae, undoubtedly affect gastropod populations. As dominant grazers in streams, any 

factors that reduce the abundance and availability of algae, or the suitability of substrata 

for feeding and mobility, will be detrimental to gastropod populations. Buck, (1956) 

demonstrated that reproduction of warm water species such as largemouth bass 

(Microptews salmoides), bluegill (Lepomis macrochirus), and redear sunfish (Lepomis 

microlopbus), can also be inhibited by higher levels of suspended sediment. The feeding 

success of aquatic species that rely on visual search strategies can be affected in 

sedimented and turbid waters. 

Effects on Coral:

The earliest coral reef researchers recognized that coral reefs were strongly inhibited 

wherever muddy freshwater enters the sea by realizing that gaps in continuous fringing 

and offshore reefs faced the river mouths. Negative impacts of rivers include effects of 

freshwater and sediments. Freshwater can cause tissue bleaching, but excessive sediment 

smothers and kills coral tissue and reduces light levels and food supplied to the coral by 

symbiotic algae. Corals differ greatly in their ability to resist sedimentation, with most 
24

species being highly intolerant of even small amounts while a minority are able to 

tolerate extremely muddy conditions, and a few are even able to live directly in muddy

bottoms. The sediment tolerant corals are able to push sediment off their surface through 

a variety of mechanisms, but these all require expenditure of metabolic energy and when 

sedimentation is excessive they eventually reach the point where they can no longer spare 

the energy to keep themselves clean, and the affected tissue dies back.

Rogers (1990) observed that excessive sedimentation can adversely affect the structure 

and function of the coral reef ecosystem by altering physical and biological processes. 

Jowett (2001) reffered some classic studies of the impact of sedimentation on coral reefs 

include the work of Jorge Cortes in Costa Rica, Gregor Hodgson at El Nido, Palawan 

Island, the Philippines, and Tim McLanahan and David Obura in East Africa and 

reported that high sediment loads can smother tissue resulting in bleaching in the short 

term and death in the long-term. Rogers (1983,1990) and Torres (2001) reported that 

increased sedimentation resulting from coastal development is a major source of coral 

reef degradation. According to Rogers (1990) excessive sedimentation can affect the 

complex food web associated with coral reefs, killing not only corals but other reef 

dwelling organisms (e.g., sponges) which serve as food for commercially important fish 

and shellfish. Rogers (1983) and Dahl (1985) reported that declines in tropical reef 

fisheries in the Carribean and the Pacific are believed to be partially due to increased 

sedimentation rates. Increased sedimentation is also one of several factors, which effect 

coral recruitment. Coral larvae will not settle and establish themselves in shifting 

sediments. Rogers (1990) concluded that increases in sedimentation rates can alter the 

distribution of corals and their associated reef constituents by influencing the ability of 

coral larvae to settle and survive.

Cortes and Risk (1985) reported reduced growth rates in Montastraea annularis living in 

waters with average sedimentation rates between 20-1,000 mg cm-2d

-1

. Reduced growth 

rates and temporary bleaching in M. annularis were also reported by Dodge et al. (1974). 

In a subsequent study Torres (2001) showed that growth rates of M. annularis were 

significantly lower (P<.05) and negatively related with sediment deposition rates and

percentages of terrigenous sediments reaching a coral reef on the south coast of Puerto 

Rico. Nemeth and Nowlis (2000) reported bleaching of coral colonies at sediment 

deposition rates between 10 and 14 mg cm-2d

-1

(r2=0.92). Their study indicated that 

stress from sedimentation may lead to a decline in living coral. The indirect effect of 

increased suspended sediment load also resulted in increased turbidity and a 

corresponding decrease in light penetration limiting the photosynthetic capacity of 

symbiotic zooxanthellae and furthering the decline in coral populations. 

Effects on Wildlife:

There seem to be very few published reports on the effects of SABS on aquaticdependent wildlife (e.g., birds and seals). Reeder (1951) and Fritzell et al. (1979) 

suggested that wildlife habitat is the important component of wetland and an important 
25

function of wetlands is to provide suitable adaptations to wildlife. Krapu (1974a), 

Swanson et al. (1974, 1985) reported that alteration of vegetative cover and aquatic 

invertebrate communities has a direct impact on all wetland wildlife. Euliss and Harris 

(1987), Miller (1987), and Euliss et al. (1991) observed that aquatic invertebrates are 

important dietary items of waterfowl and other wetland-dependent birds. Krapu (1979) 

demonstrated that protein-rich invertebrate foods are needed to meet the physiological 

demands of egg-laying for hens during the breeding season and they provide essential 

amino acids for other seasonal changes such as feather molt. Thus according to 

Heitmeyer (1988) the impact of sedimentation on wetland wildlife is likely indirect, 

involving habitat changes in response to siltation events.

For the most part, aquatic-dependent wildlife are more mobile than the fish and 

invertebrates discussed above, so it stands to reason that aquatic-dependent wildlife can 

avoid most the direct effects of increased SABS. A heron or an osprey, for example, can 

simply avoid more SABS affected turbid areas, and choose areas of clearer water. If and 

when the water clears in the area, the bird can return. If increases in SABS are widespread and long-term, however, they might cause a problem for aquatic-dependent 

wildlife that consume aquatic life. For example McIntyre (1988) reported that a bear may 

have to abandon part of its range if there is failure of a salmon run and Loons are thought 

to require clear water for fishing, and may avoid nesting areas with inadequate water 

clarity.

Most of the studies of the relationship between SABS concentration, turbidity and 

aquatic-dependent wildlife involve field studies with birds. Von Eeerden and Voslamber 

(1995) describe a mass (group) fishing behavior of cormorants, which was apparently 

developed as a response to an increase in the turbidity of a lake in the Netherlands. 

Stevens et al. (1997) found that waterbirds were most abundant on the clear and variably 

turbid segments of the Colorado River and least abundant on the more turbid lower 

segment, providing evidence that turbidity makes it difficult for birds to forage 

effectively. Another study in British Columbia ponds, however, found that the 

abundance of dabbling ducks was positively correlated with turbidity and total dissolved 

nitrogen, and negatively correlated with percent of forested shoreline, percent of marsh, 

and chloride. However, there are no sufficient explanations for these relationships, and it 

is felt that their results highlighted the problems associated with interpreting correlativetype studies, especially the difficulties in assessing the biological significance of the 

observed correlations.

The impact of sedimentation on microbes has not been studied. Adamus and Brandt 

(1990) reported that sediment fallout may cover microbes, or organic matter needed for 

microbial processes, or alter redox profiles important in the performance of water quality 

processes. Finally, the ability of wetlands to remove and retain sediments is a basic 

concept of improved water quality, but many PPR wetlands are closed systems that can 

totally fill with sediments and hence lose their capacity to function properly.

Effects on Aquatic Plants:
26

SABS in the water reduces the penetration of light for aquatic plants to photosynthesize 

and grow. Hilton and Phillips (1982) indicated that reduced light penetration due to 

increased SABS concentration may cause a decrease in macrophyte growth. Submerged 

aquatic plants will not survive in the absence of light needed for photosynthesis. The 

spatial distribution and productivity of submerged aquatic macophytes is regulated in part 

by the availability of light. Stern and Stickle (1978) reported that suspended sediments in 

water interfere with the transmission of light through a liquid medium. Goldsborough and 

Kemp (1988) suggested that decreased water clarity due to SABS causes mortality of 

macrophytes at varying rates depending on the species and their tolerance. 

The effects of reduced primary production on aquatic invertebrates and fishes at higher 

trophic levels are compounded when sediment settles on remaining macrophytes. SABS 

directly limits the growth and distribution of aquatic plants (NHAB, 2000). Ryan (1991) 

suggested that not only is primary production reduced by sedimentation and turbidity, but 

macrophyte quality also is reduced as a food source. These impacts can occur with small 

increases in inorganic-based SABS. Lloyd et al. (1987) found that an increase in turbidity 

of only 5 NTU decreased primary production by 3 to 13%, and increases of 25 NTU 

decreased primary production up to 50%. According to these authors turbidity decreases 

available food sources for aquatic animals as well as a reduction in phytoplankton also 

translates to a reduction in zooplankton. Knight et al. (1998) from NHAB (2000) 

observed that suspended sediment concentrations as low as 100 mg/L can significantly 

reduce primary production. Consequently, a reduction in phytoplankton, and therefore 

zooplankton, may create cascading effects at higher trophic levels via a reduction in 

available food energy. 

Steinman and McIntire (1990) reported that sediment in transport can have an abrasive 

quality and can reduce the quantity of periphyton that grows on stream substrata. Hoetzel 

and Croome (1994) suggested that increases in lotic turbidity can limit light penetration 

and therefore reduce phytoplankton production. Thus periphyton communities are likely 

the most susceptible to the scouring action of suspended particles or burial by sediments. 

Many macrobenthic organisms are, in turn, used as a food source by fish. Decreases in 

zooplankton can occur with very low levels of suspended sediment. McCabe and O’Brien 

(1983) found that feeding efficiency of Dapbniapulex was decreased by about 25% at 6 

NTU. Kirk (1985) and Ryan, (1991) observed as turbidity increases, a reduction in plant 

growth, including algae, occur because less light is available for photosynthetic 

production. 

Areas where sediment suspension is more have been found to be areas with limited 

aquatic macrophyte growth. Best et al. (2001) reported some of aquatic macrophytes have 

experienced dramatic losses over the past two decades, a decline largely attributed to 

changes in underwater light climate brought on by increases in suspended sediment 

concentrations. Aquatic plants and algae are important components of aquatic ecosystem 

food chains. Sedimentation can affect aquatic populations by altering the available food 

supply. Aquatic plants provide structural habitat for invertebrate, vertebrate as well as 

fish population. Murkin (1989) suggested that it also provides substrates for colonization 

by algae and microbes that are important foods of aquatic invertebrates. Increased rates of 
27

sedimentation and enhanced concentrations of suspended sediments can reduce the 

primary productivity of the impacted area. Mann (1988) reported that once aquatic plants 

senescence, they contribute litter for colonization by microbes which provide additional 

food resources for aquatic invertebrates. Murkin (1989), Neill and Cornwell (1992) 

suggested that aquatic plants are major sources of carbon in aquatic habitats and are 

important food resources. Sedimentation has potential to suppress primary production 

and alter natural food chain interactions. Aquatic plant communities are likely the most 

susceptible to the scouring action of suspended particles or burial by sediments. Ellis 

(1936), Robel (1961), and Dieter (1991) demonstrated that increased sediment in the 

water column generally reduces the depth of the photic zone and hence reduces the light 

available for primary production by aquatic plants and algae. Singleton (1985) reported at 

concentrations exceeding 115 mg/L, suspended sediments can reduce light penetration 

and primary productivity. Newcombe and Macdonald (1992) suggested a reduction in 

primary productivity has the potential to appreciably decrease the food supply of 

macrobenthos that graze on periphyton. Many macrobenthic organisms are, in turn, used 

as a food source by fish.

Batiuk et al (1992) and Dennison et al. (1993) observed the large-scale declines of 

submerged aquatic vegetation (SAV) in Chesapeake Bay. It is believed to be directly 

related to increasing amounts of nutrients and sediments entering the Bay. The U.S.EPA 

Chesapeake Bay Program office established water clarity criteria to address the 

unacceptable Bay-wide decline in SAV. Water clarity criteria are based on the light 

requirements for SAV growth and survival. The criteria take total suspended solids 

(particulate matter and chlorophyll a) into account, as well as epiphytic growth and 

salinity regime. The assumption behind the use of water clarity criteria in the 

Chesapeake is that if the SAV are protected, the rest of the ecosystem will also be 

protected, because the SAVs represent the component of the Chesapeake Bay ecosystem 

that is most sensitive to increases in sediments. A detailed explanation of the derivation 

of Chesapeake Bay water clarity criteria can be found in U.S. EPA (2000b).

Bedded sediment can smother plants living on the bottom of the river or stream. SABS 

from agricultural, sewage and some industrial processes runoff can also lead to increased 

amounts of nitrates and phosphates in the water. This can create problems of excessive 

plant growth and lead to blooms of algae and weeds. Asplund (1997) argued that 

excessive increases in SABS concentration levels may create an environment beyond the 

acceptable level of tolerance for the living components of an aquatic ecosystem. For 

example uprooting of macrophytes and angular sedimentation and other related activity 

causes a significant decline in productivity and biomass. Johnson (1994) demonstrated 

that emergent aquatic plants have been disturbed by sedimentation causing for less cover 

along the shoreline and more erosion. 

The magnitude and timing of sedimentation may influence structure and recolonization of 

aquatic plant communities. Under natural conditions aquatic plant communities are 

dynamic and undergo cyclic changes in response to short- and long-term water-level 

fluctuations and salinity. The influence of increased sedimentation on lotic food webs is 

pervasive and begins at primary trophic levels. Increased rates of sedimentation and 
28

increased concentration of suspended sediments can reduce the reproductive productivity 

of the impacted area. Rybicki and Carter (1986), Hartleb et al. (1993), Jurik et al. (1994), 

and Wang et al. (1994) reported as sediment falls out of suspension, deposition may be 

adequate to bury algae, macrophytic photosynthetic substrates, and seed banks. Jurik et 

al. (1994) and Wang et al. (1994) demonstrated that sediment depths of 0.25 cm can 

significantly reduce species richness, emergence, and germination of wetland 

macrophytes. Jurik et al. (1994) also found that the greatest decreases in germination 

occurred for species with the smallest seeds. Hartleb et al. (1993) showed that seed 

germination of water milfoil (Myriophyllum spicatum L.) was significantly reduced when 

buried by more than 2 cm of sediment. Rybicki and Carter (1986) found that survival of 

water-celery (Vallisneria americana MICHX) tubers declined 90% when buried by 10 

cm, and by 100% when buried by 25 cm of sediment. Although, these studies 

demonstrated the relationship between sedimentation and germination, the causative 

agent that inhibits germination or survival is poorly understood. For example, covering of 

seeds with varying depths of sediment may alter light and/or redox conditions that inhibit 

seed germination or the sediment may create a physical barrier to emergence. 

Van der Valk and Davis (1976) identified four prairie wetland cyclic conditions, which 

includes dry marsh, regenerating marsh, degenerating marsh, and lake. They reported that 

during the dry marsh or drawdown phase, sediments and seed banks are exposed and 

mudflat annuals and emergent plant species germinate and recolonize the wetland. Since 

recolonization is dependent on viable seed banks, the covering of seed banks with 

sediment has potential to impede the recolonization process (Jurik et al. 1994). 

Additionally, loss of wetland volume from acclerated sedimentation makes wetlands 

shallower, which allows monodominant stands of cattails, normally restricted to water 

depths <60 cm (Bellrose and Brown 1941) to expand. Linz et al. (1996) suggested such 

stands of vegetation contribute little to biological diversity and exacerbate problems with 

agricultural interests because they provide roost sites for blackbirds that depredate cereal 

crops.

Robert et al (1999) reported that indirect impact sediment exerts on water quality through 

its influence on aquatic plants, organic exchange substrates, and microbial populations. 

Reduction of light available for photosynthesis due to turbidity and the burial of 

macrophyte seed banks are obviously negative impacts of excessive sediment entering 

wetlands from adjacent fields. Aquatic macrophytes and algae are important in the 

uptake, short-term storage, and cycling of nutrients in aquatic environment. Negative 

impacts on plants from sediments may alter water quality functions. Martin and Hartman 

(1987), and Gleason (1996) indicated that increased input of allochthonous inorganic 

matter to aquatic eosystem would reduce the availablity of organic exchange surfaces 

important for sorption of contaminants, especially on the thin aerobic zone at the soilwater interface.

Andrew et al (2002) demonstrated that sedimentation was partially responsible for 

differences between habitat types. It acted as a negative disturbance on the abundance of

algae and survivorship of invertebrates. When combined with differences in light 

intensity, high rates of sediment accumulation had slight negative effects under natural 
29

light, but under shaded conditions these negative effects were substantially increased, 

causing changes to the structure of the whole assemblage. Thus sedimentations have 

profound impact on the aquatic plants at the individual, population and community levels. 

In such an environment, reductions in food availability (caused by reduction of 

photosynthesis process), environmental quality, and habitat can directly affect growth and 

mortality rates of aquatic plant species and algae, which cause major impact at multiple 

trophic levels. In sedimented environments, reductions in species density, biomass, and 

diversity throughout a trophic level are translated into reductions in energy input to the 

next trophic level. Berkman and Rabeni (1987) suggested decreases in plant population 

result into decrease in population of zooplankton, and insect abundance and biomass 

initiate reductions in herbivore, omnivore, and, consequently, predator classes of fish. 

Biotic adaptations to sediment perturbation also may lead to changes in local community 

composition. Lowrance et al. (1985) reported a further synergistic effect when lotic 

communities are impacted by pesticides and other toxins entering the river or stream with 

eroded material.

Other Effects:

Most of the Sedimentation is caused by erosion. Erosion within streams is a natural and 

mechanical process and is affected by parameters such as stream flow, channel structure 

and stability, streambed composition, and disturbance within the watershed such as fire, 

landslide or ice scour. Human activity effects increased turbidity in lakes and rivers 

through a number of avenues. Kerr (1991), and Liddle and Scorgie (1980) and Anderson 

(1996) reported that erosion from nearby agricultural lands, runoff as a result of increased 

water proofing of the land, industrial effluents, urban development, agriculture, mining, 

forestry, stream channel alteration (such as dams and channelization), and instream, 

construction associated with developments (such as bridges, roads or pipeline and 

transmission crossings), flooding, dredging, thermal stratification and water currents are 

causes of sedimentation. The disturbance of lands accelerates erosion and increases the 

delivery of sediment to stream systems. Any activity that is undertaken within a 

watershed that disturbs land surfaces has the potential to increase sediment delivery to 

streams. Human activity (such as motorboat and personal watercraft use) is the subject of 

increasing concern in lakes and rivers all over the United States and Europe. Liddle and 

Scorgie (1980) suggested that physical forces associated with recreational activity on 

lakes include wash, turbulence, propeller action (cutting effects), direct contact with 

bottom sediment and disturbance by sight and sound. 

Changes in Stream Morphology:

In response to changes in the environment, ecosystems often undergo alteration in 

community composition and structure. The entry of SABS to water bodies from upland 

sources is a natural and mechanical process. According to Brett (1958), and Kohen and 

Bayne 1989) organisms respond to environmental change in order to avoid or minimize 

effects on fitness. Klein (1984) reported that elevated sediment loads can change channel 

geometry in addition to altering stream bed composition. When the SABS rate entering a 
30

water bodies exceeds the capacity of the watercourse to transport or assimilate the 

sediment stress may occur to the aquatic community. If an organism cannot compensate 

for a change in the environment and suffers a reduction in fitness, the environmental 

change is termed as “Stress”. Grime (1989) suggested that a stress limits either the rate of 

resource acquisition or growth and reproduction so that fitness is reduced. Selye (1950) 

defined the stress as “the sum of all the physiological responses by which an animal tries 

to maintain or re-establish a normal metabolism in the face of a physical or chemical 

force”. At this point the suitability/ adaptability and productivity of aquatic habitats may 

be altered. Anthropogenic inputs of SABS into stream and river environments can 

cause stress to aquatic systems and thereby, directly and indirectly impact upon aquatic 

organisms behavior and health. Anderson et al. (1996) suggested that increased 

concentrations of SABS can have direct effects on aquatic species behavior, physiology 

and populations.

Elevated levels of sediment deposition can reduce the depth of water reservoir and 

produce a net reduction in riffle areas as a result available habitat can reduce. For 

example, Bjornn et al. (1977) indicated deposition of sediments in pools and other areas 

of instream cover can cause a decrease in the aquatic species holding capacity of a water 

bodies. Excess sediment in rivers and streams has been identified as the largest and most 

pervasive water pollution problem faced by aquatic systems in North America. Smith and 

Saunders (1965) found that decreased aquatic biota populations were related to infilling 

of available cover. Alexander and Hansen (1992) also noted that a decrease in sand 

bedload sediment was associated with an increase in aquatic organism populations. 

Alabaster and Lloyd (1982) observed that changes in physical morphology of the stream 

can also inhibit the movement of aquatic organism or change the distribution of fish.

Conclusion:

The impact of SABS on aquatic habitat is significant. Evidence for the conclusion exists 

in the numerous studies that have focused on the impacts to fish, invertebrates, aquatic 

plants, water chemistry, erosion and habitat. Aquatic habitats face several challenges 

from SABS and best practice management is needed in ecologically sound manner. 

Water bodies are especially valuable to preserve the ecosystem as well as for human 

kinds for drinking and recreational (e.g. sports, fishing, and boating industries) activities. 

Therefore the huge aquatic resource of nature must be protected in order to ensure they

remain valuable for drinking and recreational use. SABS are significant contributors to 

decline in populations of aquatic organisms. The current literature review examined the 

effects of increased SABS load on aquatic ecosystems emphasizing various natural and 

mechanical sources of elevated sediment loads. The effect of sediment on receiving water 

ecosystems are complex and multi-dimensional and further compounded by the fact that 

sediment flux is a natural and vital process for aquatic systems. The review included the 

effects of SABS on fish (behavioral, physiological and population effects) and their 

habitats (including spawning, rearing, food production, summer and overwintering 

habitats). A habitat effects relationship was presented which related the concentration and 

duration of SABS exposure events to the alteration of fish habitats. This relationship 

allows for post-disturbance evaluation of the potential effects on fish habitat. Effects on 

aquatic flora and fauna may be expressed through pervasive alterations in local food 
31

chains beginning at the primary trophic level. Decreases in primary production are 

associated with increases in SABS concentration and turbidity. This cause negative 

cascading effects through depleted food availability to zooplankton, insects, freshwater 

mollusks, and fish. Direct effects of SABS at each trophic level are mortality, reduced 

physiological function, and avoidance. However, decreases in available food at trophic 

levels also result in depressed rates of growth, reproduction, and recruitment. Imbalance 

of suspended solids and bedded sediments (SABS) cause all these problems in water 

bodies. In many water bodies, SABS are severely out of balance due to human activities 

within the watershed. In most cases the problem is excessive sediments, but in some 

cases the problem is too little sediment. Within recent years, the States and EPA have 

identified sedimentation issues that are causing significant water quality problems for 

which water quality criteria for sedimentation have yet to be established. 

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